Human Impacts on Salt Marshes: A Global Perspective 9780520943759, 9780520258921

Table of contents :
CONTENTS
ACKNOWLEDGMENTS
Introduction. Are Salt Marshes at Risk?
PART I Invasions in North American Salt Marshes
1 Spartina Introductions and Consequences in Salt Marshes: arrive, survive, thrive, and sometimes hybridize
2 Changes in Community Structure and Ecosystem Function Following Spartina alterniflora Invasion of Pacific Estuaries
3 Invasive Animals in Marshes: biological agents of change
4 Phragmites australis in Eastern North America: A Historical and Ecological Perspective
PART II Human Inputs and Consumer Effects
5 Opportunistic Herbivores, Migratory Connectivity, and Catastrophic Shifts in Arctic Coastal Systems
6 Top-Down Control and Human Intensification of Consumer Pressure in Southern U.S. Salt Marshes
7 Alligator Hunters, Pelt Traders, and Runaway Consumption of Gulf Coast Marshes: a trophic cascade perspective on coastal wetland losses
PART III Land Use and Climate Change
8 Shoreline Development and the Future of New England Salt Marsh Landscapes
9 Tidal Restrictions and Mosquito Ditching in New England Marshes: case studies of the biotic evidence, physical extent, and potential for restoration of altered tidal hydrology
10 Impacts of Global Climate Change and Sea-Level Rise on Tidal Wetlands
11 Potential Impacts of Elevated CO2 on Plant Interactions, Sustained Growth, and Carbon Cycling in Salt Marsh Ecosystems
PART IV Die-off, Loss, and Conservation
12 From Climate Change to Snails: potential causes of salt Marsh Dieback along the u.s. eastern seaboard and gulf coasts
13 Patterns of Salt Marsh Loss within Coastal Regions of North America: presettlement to present
14 The Use of Science in the Restoration of Northeastern U.S. Salt Marshes
15 Conserving the Diverse Marshes of the Pacific Coast
PART V International Perspectives
16 Human Modification of European Salt Marshes
17 Human Impacts and Threats to the Conservation of South American Salt Marshes
18 Anthropogenic Threats to Australasian Coastal Salt Marshes
CONCLUSION Salt Marshes under Global Siege
CONTRIBUTORS
INDEX

Citation preview

TH E STE P H E N B ECHTE L F U N D I M P R I NT I N EC O LO GY AN D TH E E NVI R O N M E NT

The Stephen Bechtel Fund has established this imprint to promote understanding and conservation of our natural environment.

The publisher gratefully acknowledges the generous contribution to this book provided by the Stephen Bechtel Fund.

HUMAN IMPACTS on SALT MARSHES

HUMAN IMPACTS on SALT MARSHES A Global Perspective

EDITED BY

Brian R. Silliman Edwin D. Grosholz Mark D. Bertness

UNIVERSITY OF CALIFORNIA PRESS Berkeley Los Angeles London

University of California Press, one of the most distinguished university presses in the United States, enriches lives around the world by advancing scholarship in the humanities, social sciences, and natural sciences. Its activities are supported by the UC Press Foundation and by philanthropic contributions from individuals and institutions. For more information, visit www.ucpress.edu. University of California Press Berkeley and Los Angeles, California University of California Press, Ltd. London, England © 2009 by The Regents of the University of California Library of Congress Cataloging-in-Publication Data Human impacts on salt marshes : a global perspective / edited by Brian R. Silliman, Edwin D. Grosholz, and Mark D. Bertness. p. cm. Includes bibliographical references and index. ISBN 978-0-520-25892-1 (cloth : alk. paper) 1. Salt marsh ecology. 2. Salt marshes—Effect of human beings on. 3. Salt marshes—Effect of human beings on—North America. I. Silliman, Brian R., 1972– II. Grosholz, Edwin. III. Bertness, Mark D., 1949– QH541.5.S24H86 2009 577.69'27—dc22 2008048366 Manufactured in the United States of America 16 15 14 13 12 11 10 09 10 9 8 7 6 5 4 3 2 1 The paper used in this publication meets the minimum requirements of ANSI/NISO Z39.48-1992 (R 1997) (Permanence of Paper). Cover illustration: Above: Sprague Marsh, seen from Morse Mountain, in Phippsburg, Maine, at high tide in summer 2007. The primary river channel, dug in 1958, is now a sinuous secondary branch. Photograph courtesy Keryn Bromberg Gedan. Below: A crab feeding on Spartina alterniflora near San Antonia in Patagonia, Argentina. Photograph by Brian R. Silliman.

TO ALL THOSE WORKING TO CONSERVE THE WORLD’S SALT MARSHES

CONTENTS

Acknowledgments / ix Introduction: Are Salt Marshes at Risk? / xi

Part I • Invasions in North American Salt Marshes 1 • SPARTINA INTRODUCTIONS AND CONSEQUENCES IN SALT MARSHES: ARRIVE, SURVIVE, THRIVE, AND SOMETIMES HYBRIDIZE / 3

D. R. Strong and D. R. Ayres 2 • CHANGES IN COMMUNITY STRUCTURE AND ECOSYSTEM FUNCTION FOLLOWING SPARTINA ALTERNIFLORA INVASION OF PACIFIC ESTUARIES / 23

Edwin D. Grosholz, Lisa A. Levin, Anna C. Tyler, and Carlos Neira 3 • INVASIVE ANIMALS IN MARSHES: BIOLOGICAL AGENTS OF CHANGE / 41

James E. Byers 4 • PHRAGMITES AUSTRALIS IN EASTERN NORTH AMERICA: A HISTORICAL AND ECOLOGICAL PERSPECTIVE / 57

Laura A. Meyerson, Kristin Saltonstall, and Randolph M. Chambers

Part II • Human Inputs and Consumer Effects 5 • OPPORTUNISTIC HERBIVORES, MIGRATORY CONNECTIVITY, AND CATASTROPHIC SHIFTS IN ARCTIC COASTAL SYSTEMS / 85

Hugh A. L. Henry and Robert L. Jefferies 6 • TOP-DOWN CONTROL AND HUMAN INTENSIFICATION OF CONSUMER PRESSURE IN SOUTHERN U.S. SALT MARSHES / 103

Brian R. Silliman, Mark D. Bertness, and Mads S. Thomsen 7 • ALLIGATOR HUNTERS, PELT TRADERS, AND RUNAWAY CONSUMPTION OF GULF COAST MARSHES: A TROPHIC CASCADE PERSPECTIVE ON COASTAL WETLAND LOSSES / 115

Paul A. Keddy, Laura Gough, J. Andy Nyman, Tiffany McFalls, Jacoby Carter, and Jack Siegrist

Part III • Land Use and Climate Change 8 • SHORELINE DEVELOPMENT AND THE FUTURE OF NEW ENGLAND SALT MARSH LANDSCAPES / 137

Mark D. Bertness, Brian R. Silliman, and Christine Holdredge

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9 • TIDAL RESTRICTIONS AND MOSQUITO DITCHING IN NEW ENGLAND MARSHES: CASE STUDIES OF THE BIOTIC EVIDENCE, PHYSICAL EXTENT, AND POTENTIAL FOR RESTORATION OF ALTERED TIDAL HYDROLOGY / 149

Caitlin Mullan Crain, Keryn Bromberg Gedan, and Michele Dionne 10 • IMPACTS OF GLOBAL CLIMATE CHANGE AND SEA-LEVEL RISE ON TIDAL WETLANDS / 171

J. Court Stevenson and Michael S. Kearney 11 • POTENTIAL IMPACTS OF ELEVATED CO2 ON PLANT INTERACTIONS, SUSTAINED GROWTH, AND CARBON CYCLING IN SALT MARSH ECOSYSTEMS / 207

Jordan R. Mayor and Caitlin E. Hicks

Part IV • Die-off, Loss, and Conservation 12 • FROM CLIMATE CHANGE TO SNAILS: POTENTIAL CAUSES OF SALT MARSH DIEBACK ALONG THE U.S. EASTERN SEABOARD AND GULF COASTS / 231

David T. Osgood and Brian R. Silliman 13 • PATTERNS OF SALT MARSH LOSS WITHIN COASTAL REGIONS OF NORTH AMERICA: PRESETTLEMENT TO PRESENT / 253

Keryn Bromberg Gedan and Brian R. Silliman

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14 • THE USE OF SCIENCE IN THE RESTORATION OF NORTHEASTERN U.S. SALT MARSHES / 267

John Teal and Susan Peterson 15 • CONSERVING THE DIVERSE MARSHES OF THE PACIFIC COAST / 285

J. C. Callaway and J. B. Zedler

Part V • International Perspectives 16 • HUMAN MODIFICATION OF EUROPEAN SALT MARSHES / 311

A. J. Davy, J. P. Bakker, and M. E. Figueroa 17 • HUMAN IMPACTS AND THREATS TO THE CONSERVATION OF SOUTH AMERICAN SALT MARSHES / 337

Cesar S. B. Costa, Oscar O. Iribarne, and Jose M. Farina 18 • ANTHROPOGENIC THREATS TO AUSTRALASIAN COASTAL SALT MARSHES / 361

Mads S. Thomsen, Paul Adam, and Brian R. Silliman

Conclusion: Salt Marshes under Global Siege / 391 Contributors / 399 Index / 401

ACKNOWLEDGMENTS

We thank thirty-five anonymous reviewers for their time and help in editing the chapters of this book. We also thank two reviewers for their comments on the entire volume. While editing this book, we received support from the Mellon Foundation; Georgia, Rhode Island, and California Sea Grants; the National Oceanic and Atmospheric Administration; the University of

Florida; and the National Science Foundation. This is a contribution of the Georgia Coastal Ecosystems LTER and the University of Georgia Marine Institute on Sapelo Island, Georgia. Thanks to the David H. Smith Conservation Research Fellowship, The Nature Conservancy, Stephanie Wear, and Geri Silliman for providing the inspiration for this book.

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introduction

Are Salt Marshes at Risk?

The impetus for this book was generated from a symposium organized by Brian Silliman and Mark Bertness at the 2003 meeting of the Ecological Society of America in Savannah, Georgia, entitled “Anthropogenic Modification of North American Salt Marshes: Causes, Consequences, and Recommendations.” The goal of the session was to bring together leading marsh ecologists from North America to (1) introduce the causes and consequences of anthropogenic impacts on marsh structure and function and (2) facilitate discussion of the ways these threats can be addressed with conservation measures. This successful symposium produced two fundamental conclusions: (1) human activities, such as shoreline development and invasive species, severely threaten marsh structure and function over the entire continental coastline; and (2) ecologists, conservation biologists, and resource managers worldwide would benefit greatly from an edited volume that summarized and synthesized our findings, other applicable recent research, and management recommendations. The resultant edited volume deals primarily with case studies of human impacts in North American salt marshes and includes selected proceedings of the symposium, additional invited chapters, and two examples that describe how scientific findings have been recently integrated

into ecoregional marsh conservation plans on the eastern and western coasts of the United States. These chapters consider the most pressing and extensive human-induced threats to salt marshes in North America: invasive species, overfishing, nitrogen eutrophication, sea-level rise, rising temperatures, increasing atmospheric carbon dioxide, altered hydraulic and sedimentation regimes, drainage, reclamation, and shoreline development. To provide a global perspective on the North American focus, the last section of the book contains three chapters providing comparative analyses with salt marshes in Europe, South America, as well as Australia and New Zealand. We conclude with a synthesis chapter that highlights distribution of human-generated threats across North America, describes commonalities and contrasts among regions in threats and conservation, and discusses general ecological and conservation lessons learned. This book has the following specific objectives: • Elucidate causes and consequences of anthropogenic impacts on the organization of North American salt marshes and the ecosystem services they provide. • Provide coastal resource managers with information necessary to identify and

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ameliorate human-induced threats by presenting concrete recommendations rather than simply “suggesting experimental results be considered in management decisions.” This information is provided at the end of each chapter. • Provide a rare example in which regional ecological findings are integrated on a continental scale, thereby allowing for creation of cohesive, multisite conservation efforts. • Begin discussion and synthesis of the general ecological lessons learned from the study of human impacts on salt marsh communities on a global scale. Our goal in the intellectual design and organization of this book is to have the volume be useful in multiple contexts: (1) as a textbook example of how basic ecological research provides knowledge necessary to detect and predict human impacts for professional academics in the field of ecology and conservation science; (2) as a scientific and technical resource for scientists, advanced graduate and undergraduate students, and professional research workers; and (3) as a resource for marsh mangers to facilitate identification of pressing human impacts and the necessary steps to ameliorate those threats using regionally based and multiplethreat abatement approaches. This book represents the first effort to bring together the major research findings on the effects and potential consequences of human activities on the ecology of North American coastal salt marshes. It is unique in that it represents one of the few books regarding any natural community to focus on and integrate, on the continental scale, results of local ecological research on human impacts. Many of the chapters describe important results already obtained; thus, the book is a synthesis of several scientific papers and years of work. However, many of the perspectives and conclusions generated from this compilation undoubtedly are new and distinctive, and they provide valuable ecological and conservation lessons for all those with an interest in salt marsh ecosystems.

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We have divided the book into five major sections based on both a conceptual and geographic basis: (1) “Invasions in North American Salt Marshes”; (2) “Human Inputs and Consumer Effects”; (3) “Land Use and Climate Change”; (4) “Die-off, Loss, and Conservation”; and (5) “International Perspectives.” Examples in the first four sections derive from research in North American salt marshes. The last section includes three chapters that compare human impacts in salt marshes between North America and Europe, South America, and Australia–New Zealand. The final chapter discusses regional and global distribution of human impacts and synthesis and conclusions. Part I, “Invasions in North American Salt Marshes,” focuses on the immense threat that invasive plants and animals pose to the ecology of West and East coast salt marshes in the United States. The first two chapters of this section cover plant invasion in West Coast salt marshes. In these coastal communities, the East Coast plant dominant, Spartina alterniflora, has aggressively invaded in both Washington and California, where it has hybridized with a native Spartina species. This invasion has displaced native marsh vegetation and converted intertidal marsh flats (which act as critical feeding habitats for migratory birds) to impenetrable stands of clonal grass, with dramatic impacts on benthic invertebrate communities as well as the storage and cycling of carbon and nitrogen. More has been revealed about mechanisms of this invasion than for Phragmites, and chapter 1 is devoted entirely to this topic. The second chapter in this section focuses on what we know about the consequences of Spartina invasion on the West Coast, steps taken to eradicate the resultant hybrid, and success of those measures. On northeastern and mid-Atlantic shorelines, Phragmites australis, a large, threemeter tall reed, is displacing native marsh plant assemblages, thereby converting once-diverse plant communities into monospecific stands. The structure and function of East Coast marshes are completely changed following invasion of Phragmites, and there is the real

possibility that this plant could eventually dominate marshes along the entire eastern seaboard. The third chapter in this section reveals the spatial extent of Phragmites invasion, mechanisms underlying the invasion, consequences for marsh structure and function, and recommendations for control. Chapter 4 focuses on the extent and impacts of invasive animals in tidal salt marshes on both coasts, including oysters, snails, and crabs. Invasive animals in marshes have received far less research and conservation attention in comparison to invasive plants. Importantly, this chapter shows that this bias is not due to a lack of threat from animal exotics but rather to a lack of research focus and study—an intellectual void that should be addressed rather quickly by conservation biologists, marsh scientists, and managers. Part II, “Human Inputs and Consumer Effects,” is devoted to recent research from Hudson Bay, Gulf Coast, and East Coast salt marshes that strongly challenges the entrenched, bottom-up paradigm of marsh ecology by demonstrating that grazers and their predators exert intense and expansive top-down control over marsh plant growth. Ecologists have commonly invoked salt marshes as the textbook example of a community regulated by physical factors. For over sixty years, top-down forces, such as grazing and predation, have been considered irrelevant to marsh primary production and large-scale spatial patterning of plants. Because of this bias, almost all conservation efforts to date have been framed around the marsh bottom-up paradigm and have focused almost entirely on the maintenance of natural physical processes and nutrient regimes. These chapters deal with three examples showing that at high population densities, plant-grazing snails, nutria, and snow geese all independently have the capability of completely destroying salt marshes. Populations of these animals are both directly and indirectly regulated by human activities (e.g., snow geese populations have exploded due to increased food availability in abandoned agriculture fields, and snail and nutria numbers are kept in check by human-

harvested predators—blue crabs and alligators). All three chapters (1) argue that both marsh ecologists and resource managers have greatly underestimated consumers’ role in regulating marsh plant structure and function and (2) discuss how human land use, hunting, and fishing practices have escalated top-down control in salt marshes, fueling runaway consumption of marshes in many areas along the eastern North American seaboard. Part III, “Land Use and Climate Change,” includes four chapters covering a wide range of issues. These chapters summarize long-term marsh research on key, forcing variables of salt marsh structure, especially those that humans have significantly altered. These include nitrogen loading, sea-level rise, shoreline development, human changes in sedimentation and hydraulic regimes, marsh reclamation, shoreline armoring/hardening, flow obstruction, and global warming and associated increases in temperature and carbon dioxide. For example, chapter 10 discusses the interactive effects of sea-level rise on marsh plant productivity, nutrient cycling, and elevation. Chapter 11, in contrast, summarizes marsh field experiments that have been running for seventeen years in Maryland examining impacts of rising atmospheric carbon dioxide on salt marsh ecosystem processes. In each of these chapters, the authors identify the role of humans in modifying forcing variables and summarize results of research that demonstrate and predict effects of human impacts on marsh structure and function. Part IV, “Die-off, Loss, and Conservation,” includes two chapters that document and discuss recent and long-term patterns in marsh loss on the North American continent and two case studies detailing integration of marsh science into ecoregional conservation plans. The first chapter summarizes what is known to date about the unparalleled marsh dieback events that have occurred over the past ten years in Spartina-dominated U.S. East Coast marshes, killing more than 250,000 acres. Only one study has experimentally demonstrated significant contributing factors to marsh die-off—drought,

introduction

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pathogenic fungi, and overgrazing by snails— but many other forces are hypothesized to be involved as well, including crab herbivory, metal toxicity, nematodes, changing pH, and pollution. The second chapter in this section summarizes marsh loss studies over the entire continent, spanning back two hundred years, to provide an unprecedented continent-wide baseline for marsh coverage before massive human impacts began. The next two chapters deal with how government and nongovernment agencies (e.g., The Nature Conservancy, private consulting firms) have incorporated results and recommendations from ecological research directly into marsh conservation efforts. These chapters are unique for conservation biology books in that they actually highlight how science has been integrated into conservation practice by providing real-life examples of salt marsh conservation planning and action. Part V presents three chapters comparing human impacts on salt marshes in North America to those in Europe, South America, and Australia–New Zealand. These chapters are authored by internationally recognized salt marsh ecologists from England, the Netherlands, Spain, Chile, Brazil, Argentina, Australia, and New Zealand. These authors have summarized human impacts in their homeland marshes and compared and contrasted those results with findings in this book. In some cases, summary of human impacts in their homeland salt marshes goes back more than two thousand years (the Netherlands) and includes impact summaries from both colonial and native societies (Australia–New Zealand). In the final chapter, we conclude by summarizing the distribution of human impacts on marshes across North America, discussing the current state and predicted future of North American salt marshes, and presenting the general ecological and conservation lessons learned from integration of work across such large spatial scales. Why is this book important? First, salt marshes play a critical role in the ecology and

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economy of North American shorelines, but marsh structure, function, and persistence are severely threatened by human activities over the entire continental coastline. To protect these valuable natural resources, we must clearly identify human-generated threats and devise a cohesive plan to abate them, a process we hope begins and is well under way with the publication of this book. Second, a long-held myth holds that salt marshes are resilient systems and thus do not need protection and study in regard to human impacts. Specifically, most nonscientists, managers, and ecologists have long viewed salt marshes as systems that are relatively resilient to human threats in comparison to “more sensitive” systems such as coral reefs and tropical rain forests. Indeed, salt marshes have most often been regarded as the quintessential systems that act as natural buffers to protect adjacent communities such as estuaries and forests from the ill effects of natural and humaninduced threats (e.g., hurricanes, flooding, toxins in groundwater, nutrient overloading, bacterial overloading in groundwater, fishery collapse, increasing atmospheric carbon dioxide). Third, because marsh reclamation is now believed to be severely reduced in North America, some arenas of thought hold that the threat to the long-term persistence of North American salt marshes has subsided. In this book, we present a body of evidence that dispels the myth that salt marshes are currently not at great risk from human activities and show that they are severely threatened over the entire continent and, indeed, world. Most North American salt marsh ecologists will be surprised and enlightened to realize that virtually all mainland European salt marshes have been actively farmed and managed since they began accreting after the last ice age, making the notion of a pristine marsh nonsense in Europe. The findings of this compilation show that only through an integration and synthesis of recent research can we begin to recognize and identify the diversity, extent, and impact of human-induced threats to marshes throughout the world.

Finally, we hope to use the work revealed in this volume to generate a loud call to arms for integrated conservation efforts that incorporate the most up-to-date science carried out in salt marshes both within and outside the marsh area of concern. This volume shows that ecologists and managers have likely paid significantly in the form of further marsh degradation for only focusing on one or a few threats at a time and ignoring all those that do not fit into current paradigms. By using this volume as a

guide to identify and map these human impacts over the entire continent and to highlight their real and potential devastating interactions, we hope to generate a powerful wave of discussion, research, and conservation action that will reverse the tide of current fortune for salt marshes under global siege. BRIAN R . SILLIMAN EDWIN D . GROSHOLZ MARK D . BERTNESS

introduction

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PART I

Invasions in North American Salt Marshes

1

Spartina Introductions and Consequences in Salt Marshes arrive, survive, thrive, and sometimes hybridize D. R. Strong and D. R. Ayres Maritime Spartina species define and maintain the shoreline along broad expanses of temperate coasts where they are native. The large Spartina species grow lower on the tidal plane than other vascular plants; tall, stiff stems reduce waves and currents to precipitate sediments from turbid estuarine waters. With the right conditions, roots grow upward through harvested sediments to elevate the marsh. This engineering can alter the physical, hydrological, and ecological environments of salt marshes and estuaries. Where native, Spartinas are uniformly valued, mostly for defining and solidifying the shore. The potential to terrestrialize the shore was the rationale of many of the scores of Spartina introductions. In a time of rising sea levels, these plants are valued as a barrier to the sea in native areas and in China and Europe where they have been cultivated. In contrast, in North America, Australia, Tasmania, and New Zealand, nonnative Spartinas are seen as a bane both to ecology and to human uses of salt marshes and estuaries. Four of the seven large-scale invasions involved interspecific hybrids between introduced and native Spartinas, or intraspecific hybridization between formerly allopatric populations. Rapid evolution driven by selection of genotypes particularly adapted for invasive behavior could be the cause of observed high spread rates of hybrid cordgrass. The study of Spartina introductions is a rich mixture of social and basic sciences, with interaction of human values, ecology, and evolution.

Spartina species, cordgrasses, are powerful ecosystem engineers and grow over a great range of elevations in the intertidal zone (see chap. 16 in this volume). Their tall, stiff stems Editor’s Note: “Arrive, Survive, and Thrive” as a rubric for species introduction and invasion was coined by Kevin Rice, Richard Mack, and Spencer Barrett.

reduce wave energy and cause sediments carried in estuarine waters to precipitate (see chap. 2); the plants then grow into these sediments with the result of marsh elevation. The motivation for many early Spartina introductions was to stabilize shorelines and terrestrialize intertidal lands. Storm and tidal defenses in

3

a time of rising sea levels are more recent values for these plants on the low coasts of Europe and China, as well as where the maritime Spartina species are native. In other places, such as North America, Australia, Tasmania, and New Zealand, the potent ecosystem engineering of nonnative and hybrid Spartinas is seen as a bane both to ecology and to human uses of salt marshes and estuaries. Many of the scores of Spartina introductions for which we have records have been purposeful, and a few have come about as inadvertent hitchhikers on other human activities. Most introductions failed, some spread little, and a few have spread widely by dint of the ability of the seed of these plants to float on the tide. Seedladen inflorescences can disperse great distances within rafts of wrack, which form in the fall when the aboveground parts of the plant senesce. The four most extensively introduced species are the maritime species S alterniflora, S. patens, S. densiflora, and the allopolyploid of recent origin, S. anglica. Growing near docks, cordgrasses were an abundant and convenient source of cushioning for ballast. The oldest known introduction, S. densiflora, from South America to the Gulf of Cadiz, Spain, in the sixteenth century (Castillo et al. 2000), may well have been ballast packing. The first records of this practice are from seventeenth-century England (Ranwell 1967). On the East Coast of North America, bales of S. patens and S. alterniflora were packed as cushions among heavy items in the holds of ships (Civille et al. 2005). However, the large numbers of voyages from the Atlantic to the Pacific from the sixteenth century onward resulted in no known introductions of Spartina. All of the Pacific introductions came with the twentieth century, and none are known to have resulted from ballast. The first recorded introductions of S. alterniflora were to France (1803) and England (1816) from North America. In both countries, the invasions led to stands of the plant that were used by people. For example, early in the nineteenth century, residents used S. alterniflora to thatch roofs on the Itchen River, England 4

(Marchant 1967). These invasions initially thrived then receded. Patches of dieback are not unusual for S. alterniflora (Mendelssohn and McKee 1988) and S. anglica (Goodman, Braybrooks, and Lambert 1959). S. alterniflora apparently did not spread much beyond its introduction sites in Brittany and the southwest of France (Baumel et al. 2003), and it is now extremely rare in the United Kingdom (Gray, Marshall, and Raybauld 1991). In both countries, early ranges of S. alterniflora overlapped with S. maritima in multiple places, giving great potential for hybridization. Marchant (1967) speculates that these S. alterniflora introductions were from ships’ ballast from America. While it is possible that the European introductions were inadvertent, it is also reasonable that European visitors to North America purposefully carried the plant home. S. alterniflora is a much taller and more robust plant than the European S. maritima, and the American species had obviously greater potential to produce fodder and fiber. It is also reasonable to assume that people who used the plant would encourage its growth and spread. The contrast between salt marshes along the Atlantic Coast of North America, the center of diversity for Spartina, and those where cordgrasses have been introduced illustrates the influences of these plants. In Atlantic North American estuaries, there are no invasive cordgrasses, and the natives are distinctly valuable in maintaining the shoreline habitat (Warren et al. 2002). Only at the fringes of the range of the genus—in Europe, the South Atlantic, and the Pacific—do native and invading Spartinas coexist. In estuaries of the Pacific, Australia, Europe, and China, nonnative Spartinas bring large changes (see chap. 16 in this volume). Human introductions of Spartinas have led to hybrids, some of which spread rapidly and are powerful ecological engineers. Four of the seven largest invasions are by hybrids between introduced and native Spartinas (S. anglica in Europe, China, and Puget Sound, Washington; and S. alterniflora ⫻ S. foliosa in San Francisco Bay, California) and another (S. alterniflora to

invasions in north american salt marshes

China) was by intraspecific hybrids selected for vigor and fertility. The other two large-scale invasions are S. alterniflora in Willapa Bay, Washington, which is not hybridized, and S. densiflora in Humboldt Bay, California, which could bear a remnant of hybridization with S. alterniflora (Baumel et al. 2002). Spartina is a small genus of halophytic species in the Chloridoideae, a monophyletic lineage of the Poaceae (Hsiao et al. 1999). Most species are native to the maritime, north temperate Atlantic, New World (Mobberley 1956; Baumel et al. 2002). S. alterniflora is native to the Atlantic shore of South America, and S. densiflora is native to both the Atlantic and Pacific of South America. Before S. anglica, only one species was native to Western Europe (S. maritima); S. anglica is an allopolyploid hybrid species that arose in the nineteenth century, in Europe, shortly after one of its parents, S. alterniflora, was introduced there from the New World. S. maritima was the European, male parental species (Ferris, King, and Gray 1997). North America is home to most species. Many maritime salt marshes of the Atlantic and Gulf coasts of North America are dominated by S. alterniflora, the smooth cordgrass, which ranges from Canada through the Gulf of Mexico and along the Atlantic coast of South America. The second-most widespread and abundant North American species is S. patens, the salt hay grass, which ranges from Canada through the Gulf and into Mexico and into the Caribbean. Pacific maritime estuaries have only two natives, S. foliosa in California from San Francisco Bay south through Baja California, and S. densiflora in the Pacific of Chile and the Atlantic of South America. The Argentine S. longispica is inferred to be a hybrid of the native S. densiflora and S. alterniflora there (Orensanz et al. 2002). While no evidence exists that such species as S. alterniflora and S. densiflora are hybridized, they are hexaploids. Research will possibly reveal evidence of hybridization in their history, such as that hinted at in S. densiflora in Humboldt Bay, California (Baumel et al. 2002). In this review,

we focus on Spartina species that have been introduced beyond their native ranges.

SPARTINAS ARE ECOSYSTEM ENGINEERS Spartinas can be particularly influential to other species especially when they affect marsh elevation of the marsh. In one well-studied marsh in New England, the accumulation and loss of mineral sediment and organic matter have maintained equilibrium of the marsh surface with sea level for four thousand years (Redfield 1972). Modification and regulation of the environment that is mediated biologically and large relative to abiotic influences is termed ecological or ecosystem engineering (Jones, Lawton, and Shachak 1994, 1997). Although ecological engineering by organisms is neither a simple nor a single phenomenon (Reichman and Seabloom 2002), wide-ranging discussion over the past decade has verified ecological engineering as widely important in nature; the concept comprises a multifaceted set of notions with evolutionary, ecosystem, and community implications as well as shortertermed autecological ones (Bruno 2000). Spartina plants are the head engineers in temperate marshes where they occur, just as are mangroves in the tropics. Spartinas epitomize organisms with powerful reciprocal influences between biotic and physical environmental features and have been called “foundation species” (Pennings and Bertness 2001). The ecosystem engineering prowess of Spartinas comes from their erect, stiff stems, which create drag, dissipate hydrodynamic forces, and reduce wave height and current velocity (see fig. 1.1). Stiffer stems are more costly metabolically than more flexible stems and can be seen as an adaptation to harvest sediment, which increases plant fitness (Bouma et al. 2005). Sediments that are delivered to marshes by the tides, storms, rainfall, and riverine input (Allen and Pye 1992; Neumeier and Ciavola 2004) are trapped within the Spartina canopy. Roots grow up through the harvested sediment, elevating the marsh. With these influences, cordgrasses can affect the ecology of the marsh

spartina introductions and consequences

5

100

Spartina Zostera

Dissipated wave height, δh (mm/m)

Stiff strips Flexible strips

SF SG

10

SF ZF SG

1 100

1000

10,000

100,000

FIGURE 1.1 Drag as a function of wave energy for Spartina anglica and Zostera noltii. The stiff culms of S. anglica give a higher slope of drag with increasing wave energy than the flexible stems of Z. noltii. The high drag lows water movement, which results in sediment deposition. From Bouma et al. 2005, fig. 6.)

well above and below where they grow on the tidal plane. With the appropriate conditions, the sediment harvesting of S. alterniflora can maintain broad, flat monospecific salt marsh platforms close to the mean high tide. In a South Carolina estuary, primary productivity of S. alterniflora was greatest toward the lower tidal elevation tolerance of Spartina, at a depth between 40 and 60 centimeters below mean high tide (see fig. 1.2). This was where most sediment was trapped and salinity lowest. Hypoxia limited productivity at depths below maximum productivity, while hypersaline pore water owing to evaporation limited growth at higher elevations; stunted, short-form S. alterniflora grew at these higher elevations. This processes created a potentially hump-shaped function of primary productivity with position along the intertidal gradient (Morris et al. 2002). It is these relationships that allow S. alterniflora marshes to maintain equilibrium with sea-level changes over short and very long time scales. 6

The coast of the Mississippi Delta was built on sediments deposited by the river and elevated in this manner by S. alterniflora as sea levels rose during the Holocene (Redfern 1983). Decreasing sediment supply has led to loss of coast areas in the delta (Stockstad 2005). In the Netherlands, a declining sediment budget could affect many aspects of salt marshes and S. anglica dynamics in the future (Riese 2005). In San Francisco Bay, a deficit of sediment will hamper restoration of the massive South Bay Salt Ponds. Sediment supply in the south arm of the estuary is less than 1 million cubic yards (MCY) per year, while the restoration project will require well over 100 MCY (Siegel and Bachland 2002). In New Zealand, both introduced S. alterniflora and S. anglica accreted so much sediment as to negatively affect mangroves and salt marshes (Lee and Partridge 1983). In Ireland (Hammand and Cooper 2002), the United Kingdom (Goss-Custard and Moser 1988), Tasmania (Kriwoken and Hedge 2000), and San Francisco Bay (Stralberg et al.

invasions in north american salt marshes

Stable region

Unstable

2800

Aboveground production (g.m–2.yr–1)

2400

2000

1600

1200

800

400

0 0

10

20

30 40 50 Depth below MHT, D (cm)

60

70

80

FIGURE 1.2 The curved relationship between net productivity of Spartina alterniflora as a function of position on the intertidal plane (depth below mean high tide [MHT]). Highest productivity occurs near the lowest elevations where salinities are lowest; at slightly lower elevations, hypoxia reduces productivity so much that S. alterniflora is excluded. At elevations above the peak productivity, evapotranspiration increases sediment salinity, which limits growth. Open circles are data from low marsh; closed circles, from high marsh. From Morris et al. 2002, fig. 2.

2004), nonnative Spartinas and hybrids cover the soft sediments and exclude the invertebrate food of shorebirds.

RATIONALES, HISTORIES, AND EVOLVING VALUES IN SPARTINA INTRODUCTIONS And that old man asked me to think of United States Marines in a Godforsaken swamp. “Their trucks and tanks and howitzers are wallowing,” he complained, “sinking in stinking miasma and ooze.” He raised a finger and winked at me. But suppose, young man, that one Marine had with him a tiny capsule containing a seed of icenine, a new way for the atoms of water to stack and lock, to freeze. If that Marine threw that Seed into the nearest puddle. Kurt Vonnegut, Cat’s Cradle

Many Spartina introductions were the product of high, if vague, hopes to solidify soft mud and increase the economic value of salt marshes. The “most obvious economic application of Spartina is to use it for the reclamation and stabilization of muddy foreshores. There is no plant in the world better fitted for this particular purpose” (Oliver 1925, 84). Though muted, early concerns about Spartina introductions anticipated those of today: overly optimistic scenarios, adverse effects on other economic uses of estuaries and salt marshes (e.g., oyster culture and navigation), limited agricultural value of reclaimed salt marshes, decreased recreational and aesthetic value of beaches invaded by the plant, and threats to the ecology and wild life of salt marshes (Ranwell 1967). “Whether the result will in the end be beneficial or to the

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contrary will depend greatly on local conditions. In any case it will be a change worth watching and studying” (Stapf 1908, 34). Herbicide applications to introduced Spartina were under way in the United Kingdom by the mid–twentieth century (Ranwell and Downing 1960). EUROPE

Today in Iberian and Mediterranean regions, S. maritima gives desired protection against erosion during storms; unlike the more robust species and hybrids forming the bulk of this review, the small, diffusely growing S. maritima does not accrete sediment during calm periods (Neumeier and Ciavola 2004). The distribution of S. maritima may have been influenced by humans; it “has been known for a long time (since 1629), and is beyond doubt truly indigenous to Europe” (Stapf 1908, 33). At the same time, this species is found along the west coast of Africa and in South Africa (Mobberley 1956). It is possible that S. maritima is a Southern Hemisphere (“tropical”) species long ago introduced to Europe by early shipping (Marchant 1967, fig. 3). A report of Spartina pollen in early Holocene sediments of South Africa (Meadows and Baxter 2001) raises the possibility of fragmented S. maritima populations strung along the Atlantic coasts of southern Europe and Africa. Such a distribution would mirror the disjunctive populations of S. alterniflora that range along the South American Atlantic coast. S. patens, first detected in Europe in 1849, was thought until recently to be native there (Mobberley 1956). It has long been widespread and harvested for hay and fodder on the Atlantic and the Mediterranean coasts of Europe as well as in its native range (Ainouche, Baumel, and Salmon 2004). Reports of the invasive S. patens in Spain (Javier et al. 2005) do not mention any desirable features of the species, such as protection against erosion. Forming dense monocultures, it is now seen as a threat to native high marsh vegetation in natural areas of the Mediterranean and Iberian coasts (Castillo et al. 2005). S. densiflora, a native of Chile and Argentina, is believed to have been introduced, 8

either accidentally or purposefully, in the sixteenth century to the southwestern corner of Spain on the Atlantic. In the Gulf of Cadiz, S. densiflora has spread to eight estuaries and a number of shoreline habitats including dunes, high marsh, salt pans, and intertidal flats. A North African invasive population is presumably derived from the Spanish introduction (Castillo et al. 2005). S. anglica is presumed to have invaded France by floating without human aid across the English Channel. It was first detected in France in 1906, at Baie des Veys, Normandy (Baumel, Ainouche, and Levasseur 2001). The invasion proceeded during the remainder of the twentieth century from Normandy southward through Brittany mainly without human intervention. S. anglica now occupies nearly all suitable habitats along this shore of France. The prodigious abilities to accrete sediment of S. anglica can greatly change intertidal elevations and ecology where it has invaded in France (Guénégou et al. 1991). The Netherlands and Germany were invaded by S. anglica during the 1920s (Gray et al. 1991). It grew lower than the native vegetation and trapped volumes of sediment there. The influence of S. anglica is now greatest in areas with most available sediment, and its influence decreases to the north where colder winters hinder the plant (Bakker et al. 2002). In the southern Netherlands, S. anglica was important for holding sediments and building elevated land that is now being restored with native vegetation (de Jonge and de Jong 2002). S. anglica spread to Ireland in the 1930s and is now seen as a bane to natural areas and conservation. The intentions are to eradicate it from Ireland, and only a shortage of funds has prevented attainment of this goal (Hammond and Cooper 2002). The twentieth century saw a shift from agriculture to sea defenses as the major human value of salt marshes in both the Netherlands and the United Kingdom. Rising sea levels and increasing storm strength maintain the value of S. anglica in parts of Europe as a protection from coastal

invasions in north american salt marshes

flooding (Bouma et al. 2005). Productivity of S. anglica in the United Kingdom could increase with rising temperatures and higher atmospheric CO2 concentrations, but other factors, such as competition with other marsh species, complicate this picture (Gray and Mogg 2001). In the Netherlands, human influences on salt marshes have long rivaled, if not exceeded, those of weather, geology, and hydrology. In the late twentieth century, terrestrialization of the shore has vied for limited fine sediments with ecologically and economically valuable intertidal mud flats—a phenomenon called “the Wadden Sea squeeze” (Delafontaine, Flemming, and Mai 2000). The sediments from large rivers that would feed the marshes of shallow coastal areas are now shunted offshore down deep shipping channels into the North Sea (Riese 2005). Sediment loss via shipping channels is also a threat to coastal marshes of Louisiana in North America (Redfern 1983). The dredged, deepwater passage of the Mississippi River carries sediments offshore into the Gulf of Mexico. Six hundred square kilometers of coastal Louisiana salt marsh have washed into the sea in the last decade, and the loss is greater during hurricanes (Stockstad 2005). Were these sediments to be delivered over the fresh- and saltwater marshes of Louisiana, as before dredging of the Mississippi channels, the rate of coastal loss would be much less (Committee on the Restoration and Protection of Coastal Louisiana 2006). Vast monospecific stands of S. alterniflora define this coast, and the prodigious sediment-holding and -elevating capacities of this plant would contribute substantially to the maintenance of the land there. Protection from the sea is the value of S. anglica in the southeastern United Kingdom, where the Earth’s crust is subsiding while sea level rises. However, multiple values of restoration, conservation, mariculture, and sea defense combine—and even come into conflict—in salt marsh issues in the United Kingdom. For example, the removal of S. anglica in marsh restoration and conservation has led to lawsuits

based on allegations that liberated sediment harmed nearby oyster culturing (Kirby 1994). Complications for managing S. anglica in European Union countries arise from needs to meld traditional with modern uses of the shoreline differently among regions (Pethick 2002). Moreover, the human values of salt marshes are evolving rapidly. “Until the last decade, salt marsh has often been conceived as coastal wasteland with minimal economic value, which has led to considerable loss through land reclamation for use as agriculture, caravan sites, industrial developments and marinas.” (King and Lester 1995, 181). In recent years, sea defenses that include S. anglica have risen to the highest levels among these multiple values of salt marshes (King and Lester 1995), but this can conflict with conservation. In parts of the United Kingdom, dense monospecific swards of S. anglica replaced mud flats and excluded native invertebrates and the shorebirds that feed on them (Goss-Custard and Moser 1988; Frid, Chandrasekara, and Davey 1999). Subsidence of the southeastern UK coast means a sediment deficit in the long run, and even S. anglica can afford little protection to rising sea levels under these conditions. NEW ZEALAND

Early in the twentieth century, there was unabashed enthusiasm for the potential of nonnative Spartina in New Zealand. “For thousands of years tidal salt mud flats the world over have made entrances to harbours unsightly and treacherous and have remained as vast areas of waste flats. . . . In the past they have provided an almost unconquerable challenge to man. . . . Now such mud can be conquered, and . . . reclaimed to form useful and stable farmlands. This plant which has such an important role is . . . Spartina townsendii” (Harbord 1949, 507). By 1950, the slow growth of introductions of S. anglica and S. ⫻ townsendii on the North Island of New Zealand shifted attention to S. alterniflora, albeit with a soft counterpoint of caution. “Extensive areas of tidal flats round New Zealand’s coastline, usually difficult and

spartina introductions and consequences

9

costly to develop, have become the subject of renewed interest with the introduction of the maritime grass Spartina alterniflora, which will enable many farmers to capitalize on these naturally fertile soils. . . . However, farmers are warned that the adverse influences may not always be readily apparent” (Blick 1965, 275). In the subsequent decade or so, attitudes reversed from conquering to preserving salt marshes. “In some places the problems caused by its spread are virtually insurmountable. With renewed appreciation of estuarine wetlands in their natural states, planting of any species of Spartina around the coast of New Zealand should not be allowed to take place. Suitable control and eradication measures need to be developed where Spartina is already present” (Partridge et al. 1987, 567). During the last sixteen years, new herbicides and new methods of application have eradicated all meadows and patches of S. anglica on the South Island (Miller 2004). AUSTRALIA AND TASMANIA

The motivations for introduction of S. anglica into Australia were vague and mostly the product of little more than curiosity among European colonists about what might grow there. Unlike the United Kingdom, the Netherlands, and New Zealand, where farmland from marshland was a focused objective, Australia had no general engineering or agricultural problems to be solved by the plant (Boston 1981). The successful plantings were clustered in southwestern Australia, where the last recorded planting was made in 1962. In Tasmania, the objective for S. anglica was clear: “to stabilize the mudflats so that they would eventually be above high water level and become relatively useful land . . . [and] . . . force stream flow into the central part of the river, creating a scouring effect and keeping the main channel free of mud” (Wells 1995, 12). Plantings of S. anglica were done from 1930 until 1968. Values changed and the negative results of the plant came into focus: reduction of areas of soft sediments where shorebirds forage, harm to native animal communities, large unwanted changes to the appearance of 10

the shore and beaches, and lack of access to the water across the S. anglica sward, which had been an open shore (Hedge and Kriwoken 2000). S. anglica was also seen as a threat to the Tasmanian oyster industry (Hedge and Kriwoken 2000). During the early 1990s, S. anglica was seen as undesirable and was removed with herbicides and other means in both Australia and Tasmania (Wells 1995). In recent years in Tasmania, opinion has shifted to retaining S. anglica as a habitat and food source for native species that have lost habitat to land clearing and industrial activity (M. Sheehan, personal communication) CHINA

Shorelines in China have been densely occupied for thousands of years, and shoreline dynamics have long been affected as much by human activities as by geology, weather, and hydrology (Li et al. 1991). In the last half of the twentieth century, several species of Spartina were introduced. While future coastal sediment loads will decrease with new dams (Chen et al. 2005), recent decades have seen staggeringly large volumes of sediment discharged from the vast, densely populated watersheds of the Yangtze, Yellow, and Pearl rivers. From two to three square kilometers of new intertidal lands appeared on these sediments annually (Han et al. 2000). Some botanically inclined authors have praised introduced Spartinas in China. S. anglica, introduced in 1963, was planted in about a hundred locations over 2,700 kilometers of coastline and spread to 33,000 hectares by the 1980s. Accreting sediment at high intertidal levels, S. anglica was touted as protecting dikes from erosion during typhoons and contributing to the creation of new pastureland (as part and parcel of polderizing: diking, freshwater flooding, and drainage). Spartina was harvested for green manure, animal fodder, fish food, and cellulose for paper and rope (Chung, Zhuo, and Xu 1983, 1993). The taller S. alterniflora was planted widely in China in 1975 and grew lower on the tidal gradient than S. anglica. It accreted sediment at a prodigious rate in a

invasions in north american salt marshes

band two hundred kilometers long and ninety kilometers wide by 1997 (Chung et al. 2004; Zhang et al. 2004). Brief accounts of concern that S. alterniflora displaces native marsh species have been published recently (Zhang et al. 2004; Xie et al. 2001; An et al. 2004). At least some view nonnative Spartina to be distinctly undesirable in China (Ding et al. 2008). A different perspective comes from the literature of physical geography, where Spartina comprises no more than the most minor of footnotes. Coastal areas of China are extremely vulnerable to sea level rise (Han et al. 2000). Agriculture by diking has been practiced on the shores of the Pearl River Delta since the Han Dynasty, beginning around 200 BCE. Age-old human effects on the shoreline accelerated with industrial development when China opened to the outside world. Dikes have always been the main means of reclamation of intertidal lands, and building of dikes does not need Spartina. Dikes line almost the entire Pearl River Delta, and any appearance of a tidal flat is immediately diked, drained, and flushed of salts. Wetlands, mangroves, and tidal flats have been completely eliminated. The entire region is on very low, muddy, subsiding ground that is vulnerable to freshwater floods from inland and typhoon flooding from the sea. SAN FRANCISCO BAY, CALIFORNIA

The salt marshes of San Francisco Bay are a casualty of 150 years of “reclamation,” in which the vast majority of intertidal and supertidal habitat, both brackish and salty, were diked and drained for agriculture, urban development, and industry (Williams and Faber 2001). The remaining 125 square kilometers, 5 percent of those found by Euro-Americans in this huge Pacific Coast estuary, are the foundation of conservation efforts with shorebirds of the Pacific flyway and of three federally endangered species (a mammal, a bird, and a plant). These tenuous salt marshes, and brackish and freshwater wetlands upstream, form a narrow buffer for wildlife and ecosystem services between San Francisco Bay and surrounding human population and agriculture.

Introduced Spartina played virtually no role until 1975, when the U.S. Army Corps of Engineers planted S. alterniflora, which hybridized with the native S. foliosa soon afterward (Faber 2000). Ironically, native S. foliosa was a poster child for the destruction of salt marshes. “This species is useful in reclaiming salt marshes, and in several places about San Francisco Bay it has modified the coast line and increased the acreage of many farms. The town of Reclamation received its name from the fact that in that vicinity many acres of land have been reclaimed chiefly by the use of this grass” (Merrill 1902, 6). Actually, Reclamation was not a town but rather a “locality” with few inhabitants on San Pablo Bay, upstream on the Sacramento River from San Francisco Bay (Durham 1998). A post office was established at Reclamation in 1891 and discontinued in 1903 (Salley 1977). S. foliosa is a modest ecosystem engineer, small, diffusely growing, and shallow rooted. It is but a minor league sediment trapper compared to S. anglica, S. alterniflora, and the massive S. alterniflora ⫻ S. foliosa hybrids now spreading through San Francisco Bay. Reclamation, California, originally salt- and freshwater marshes and now extensive hay and oat fields, was formed by diking, dredging, and draining the wetlands. While S. foliosa probably made little contribution to the reclamation, it is playing a role in restoration of a small fraction of the shore. Conservation interests opened 340 hectares to the sea during the 1990s, and S. foliosa will grow to form a band at the lowest tidal elevation on at least part of erstwhile Reclamation, California (Marcus 2000).

INVASIVE SPARTINAS ON THE PACIFIC COAST OF NORTH AMERICA Perhaps the first Pacific introduction was S. densiflora to Humboldt Bay, California, speculated to have been introduced there as early as the 1850s from Chile (Spicher and Josselyn 1985; Kittleson and Boyd 1997) and identified in 1984 (Faber 2000). In 1999, it was found in 329 hectares and 94 percent of the salt marshes

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of Humboldt Bay (Pickart 2001). S. densiflora was dominant over a wide tidal range, from low Salicornia virginica marsh to the species-rich high marsh (Eicher 1987). It is the object of control by herbicides and cutting, as it threatens two rare plant species classified as endangered under State of California and federal regulations, Menzies’ wallflower, Erysimum menziesii, and the beach Layia, Layia carnosa (Pickart 2005). S. densiflora from Humboldt Bay was planted at least twice in San Francisco Bay in the 1970s and 1980s (Faber 2000) and had spread from Marin County to about five hectares at three sites by 1999 (Ayres et al. 2004). In 2002, a few plants of unknown origin were found in Gray’s Harbor, Washington (Murphy 2004); and in 2005, it was discovered in Vancouver, Canada (G. Williams, personal communication). The first known Pacific introduction of S. alterniflora is to Willapa Bay, Washington, an estuary forty-four kilometers long and twenty kilometers wide just north of the Columbia River. The plant probably arrived as a hitchhiker in live oyster shipments rather than from shipping ballast (Civille et al. 2005). The first Euro-American settlements on Willapa Bay exploited native oysters and began export to San Francisco in the 1850s. The oyster industry grew rapidly, but by the 1880s, native oysters were in decline owing to overexploitation in Willapa Bay. For a few years, Atlantic oysters, Crassostrea virginica, were imported from populations cultivated in San Francisco Bay. If California cordgrass, S. foliosa, was introduced in these shipments, it was never detected and has become extinct in Willapa Bay. In 1893, eighty barrels of Atlantic oysters were imported to Willapa Bay directly from Atlantic marshes on the new transcontinental railroad. The origins of the oysters were from areas in New York Harbor and Long Island, where native S. alterniflora flourished. Upon arrival at Willapa Bay, after the nine- to thirteen-day rail journey, the contents of the barrels were spread widely in the intertidal. More than three hundred railcars of eighty to one hundred barrels each of oysters were 12

imported to Willapa Bay from New York between 1893 and 1919. Cordgrass seed and plants could easily have been placed into the barrels with the oysters. Most likely, the introduction was inadvertent. We know of no economic value for cordgrasses at that time and place, and the abundant discussion in the press during the early twentieth century about introductions from the Atlantic to the Pacific Coast of oysters and other species with commercial potential lacks any mention of Spartina. Whether the large colonies of smooth cordgrass that appeared in several places in Willapa Bay a few decades later were introduced and spread inadvertently or had some intentional component is not known. Oystermen, who were the biggest users of the bay in the early twentieth century, viewed the plant as undesirable (Sayce 1988). The earliest written record is an anecdotal account that implies but does not say that S. alterniflora was in the bay in 1911 (Sheffer 1945). The earliest photographs of it show large plants in several widely spaced areas in Willapa Bay (Civille et al. 2005). A Sheffer photograph of 1940 (Civille et al. 2005), the earliest known of smooth cordgrass in Willapa Bay, shows a plant or group of plants about 42 meters in diameter, covering nearly 1,385 square meters. Aerial photographs from 1945 show similarly large plants at seven widely separated sites on the bay, some twenty kilometers apart from one another. Present-day growth rates of S. alterniflora in Willapa Bay suggest that the plants in the photos were about fifty years old. This implies that S. alterniflora had dispersed, or was spread by humans, widely in Willapa Bay and thrived very soon after the earliest oyster trains dumped oysters there from the Atlantic, in 1893 (see fig. 1.3). Maritime Spartina seed disperses long distances quickly without human help by floating on tides and currents. There is no evidence that maritime Spartina has a seed bank, and all recruitment comes from seeds less than a year old. Floating seed was responsible for virtually the entire invasion of Willapa Bay, and rhizome fragments have made virtually no contribution to the spread of S. alterniflora there (Civille et al.

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M B

C

B

C

H

D

D

K E

E

G

G F

A

F

2000

1945

L

FIGURE 1.3 The relative abundance of the earliest known colonies of Spartina alterniflora in Willapa Bay, Washington, in 1945 (left) and the growth of this plant by 2000 (right). Note the proximity of the earliest known colonies of S. alterniflora to the approximate locations of oyster beds at the beginning of the twentieth century (Jensen Spit, North Cove, Kindred Slough, Palix River, Nemah River, and Seal Slough).

2005). New areas have consistently been invaded at densities so low that young individual plants grew separated by meters of open intertidal mud. These isolated recruits grew rhizomatously into isolated circular plants, each composed of a single genet. The lack of other emergent plants on the open mud makes new colonies of cordgrass obvious on aerial photographs, reminiscent of bacterial colonies growing on agar. Over several decades, isolated plants grew rhizomatously into continuous meadows of S. alterniflora. These meadows completely cover the intertidal mud (Davis, Taylor, Civille, et al. 2004). Isolated plants comprised a very large fraction of S. alterniflora in Willapa Bay through the twentieth century and contributed little to the spread of the invasion because they set very little seed (Davis, Taylor, Civille, et al. 2004).

The meadows produced most of the seed and were responsible for driving the invasion. The seed set of meadow plants was nearly tenfold that of isolated plants, 20 percent compared to 2 percent of florets set seed, respectively. While 92 percent of meadow plants produced at least some seed, only 37 percent of isolated plants produced any seed at all. The lower reproductive rate of isolated plants than of those in meadows was an Allee effect. Because the lowest densities of isolated recruits had high survival, grew rhizomatously, and did set some seed, the Allee effect was weak (Taylor et al. 2004). S. alterniflora is self-incompatible, outbreeding, and therefore requires pollen from another plant in order to set seed (Daehler 1998). The cause of the Allee effect was sparse pollen among the isolated plants. Only in the much older meadows was the density of the

spartina introductions and consequences

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wind-borne pollen sufficient for substantial seed set in Willapa Bay (Davis, Taylor, Lambrinos, et al. 2004). The coverage of S. alterniflora in Willapa Bay grew at a remarkably constant 12 percent or so annually over the fifty-five-year history of aerial photographs. In 2000, nearly a hundred years after the invasion began, the invader covered about 1,670 hectares, or 27 percent of the 6,000 hectares of the intertidal habitat of Willapa Bay. Without the Allee effect, the invasion would have covered virtually the entire bay long ago (Taylor et al. 2004). Other small infestations of S. alterniflora appeared in Washington, Oregon, and northern California during the twentieth century. Duck hunters introduced S. alterniflora to Dike Island in Padilla Bay, Puget Sound, Washington, between 1940 and 1946 (Riggs and Bulthuis 1994). The seed was provided by a nursery in Wisconsin, but there is no record of the maritime area from which the nursery obtained the seed. The Padilla Bay introduction had spread to cover about 1.4 hectares by 1979 and about 4.9 hectares by 1991. Flowering stems were first seen in October 1992. Some of this seed was viable, and seedlings had appeared around Dike Island by the late 1990s. Control efforts are greatly reducing S. alterniflora and S. anglica in Padilla Bay (Riggs 2005). In the 1990s, several small infestations were discovered: in Grays Harbor, Washington, twenty-five kilometers to the north of Willapa Bay; in Conner Creek farther north; and in the Copalis River yet farther to the north (Murphy 2004). These presumably arose from seed that floated northward in S. alterniflora wrack on currents from Willapa Bay, without human help. The environmental community in Washington is aware of several other small infestations of S. alterniflora in Puget Sound (Riggs 2005; Murphy 2004), and we infer that the lack of reporting indicates that these either have been eradicated or are not spreading. In Oregon, S. alterniflora from Georgia on the Atlantic Coast of North America was purposefully planted in the 1970s, in the 14

Siuslaw River at the entry to Coos Bay, Oregon (Frenkel and Boss 1988). The patch had grown to cover about two hectares by 1994, when it was sprayed with herbicides and dug up. With no visible growth, it was declared to be eradicated in 1997. In 2005, five culms of S. alterniflora were found on the same site, and another patch of this species was discovered downstream in Coos Bay, presumably brought there as rhizome material in dredge spoils from the first site (Howard et al. 2007). This indicates potential for local dispersal by rhizomes. In Humboldt Bay of northern California, a single large patch of S. alterniflora was presumably eradicated by the efforts of the California Department of Fish and Game during the 1980s or early 1990s (Cohen and Carlton 1995). The Pacific Coast of North America had two known introductions of S. anglica, both recent. That in Puget Sound was planted in 1961, perhaps with seed from England (Hacker et al. 2001). During the intervening forty years or so, the infestation spread to more than seventy sites over a roughly linear course of about 160 kilometers to affect 3,300 hectares and to cover solidly nearly 400 hectares. The San Francisco population of S. anglica was detected in the 1980s and by 2004 had grown to only twentyfour individuals covering 360 square meters. The largest plant was eight meters in diameter, and what was inferred to be the oldest plant was six meters in diameter (Ayres et al. 2004). In August 2003, S. anglica was discovered in southwest British Columbia, Canada, in the Fraser River estuary, the largest estuary on the Pacific Coast of Canada (Williams et al. 2004). Seeds presumably originated from the S. anglica populations in Puget Sound, which extend as far north as Orcas Island, about twelve miles south of the Fraser River mouth. By October, mapping had been completed, and manual removal began. Surveys in November discovered another invaded site in Boundary Bay. In 2004, a multiagency committee was established to conduct removal, begin outreach and education of naturalists, and enlist volunteer support.

invasions in north american salt marshes

S. patens, salt marsh hay, has been introduced to the Pacific Coast of North America, as well as to Europe and China (described earlier). It was first known in Oregon as three small patches in a 1939 aerial photograph of Cox Island in the Siuslaw River. Fifty years later, it had grown to ninety large monospecific patches (Frenkel and Boss 1988). In San Francisco Bay, two plants of S. patens were found in 1970 at the mouth of Suisun Bay near the town of Benicia, in the Southampton marsh (Ayres et al. 2004).

SPARTINA HYBRIDIZATION In 1974, the U.S. Army Corps of Engineers conducted test plantings of S. foliosa and Salicornia virginica on unconfined dredge spoils deposited along Coyote Hills Slough (also known as New Alameda Creek) near Fremont, California (37⬚33⬘58.62⬙ N by 122⬚07⬘44.52⬙ W) as part of a larger study to quantify and ameliorate the impact of dredging and dredged material disposal in San Francisco Bay (U.S. Army Corps of Engineers 1976). The immediate goal of these plantings was to determine the feasibility of using native marsh species to create a salt marsh on dredge spoils confined in a former salt pond (pond 3) adjacent to New Coyote Hills Slough, while the eventual goal was to evaluate whether this approach had potential utility as a means to both dispose of dredge spoil and create new salt marsh. After determining that native species were able to establish from sown and tidally borne seed, the Corps of Engineers inexplicably planted S. alterniflora in pond 3 from seed it had obtained from an environmental consulting firm in Maryland (Faber 2000). According to personal communication in 2005 between Jun Bando of the University of California, Davis, and L. Hunter-Cario, the firm’s nursery manager, the S. alterniflora seed came from marshes in Maine and Virginia in the 1970s. A genetic survey of plants growing along Coyote Hills Slough in 1994 found equal numbers of S. alterniflora and hybrids between S. alterniflora and S. foliosa, and a single S. foliosa (out of forty-five specimens) (Ayres et al. 1999).

In 1978, the Corps of Engineers planned to use both native and exotic smooth Spartina to control shoreline erosion at Alameda Island, fifteen miles to the north of pond 3 (U.S. Army Corps of Engineers 1978). When this population was genetically surveyed in 1998, only hybrid and smooth cordgrass was found; the native species was absent (Ayres et al. 1999). Similarly, the native species was absent from an introduced population at San Bruno marsh that contained equal numbers of hybrid and S. alterniflora plants in 1994. In recent surveys, hybrid cordgrass is spreading rapidly in the San Francisco estuary, while the native and nonnative parents are now becoming rarer (Ayres, Baye, and Strong 2003; Sloop 2005). The hybrid swarm hinders access to shorelines, blocks flood control channels, overgrows intertidal foraging areas of shorebirds, and without control could lead to the extinction of the native S. foliosa in San Francisco Bay by means of competition and pollen swamping (Ayres et al. 2003). The potential spread of these hybrids southward through the range of S. foliosa, in southern California and Baja California, carries the risk of global extinction of S. foliosa (Ayres et al. 2003). A large control program is trying to eradicate hybrid cordgrass from San Francisco Bay (www.spartina.org). Hybridizations resulting from human introductions loom large in Spartina biogeography and in the huge influence that nonnative cordgrasses have had in salt marshes around the world. The history and incidence of Spartina hybridizations are incompletely known. For example, the S. densiflora in Humboldt Bay, California, has genetic features that suggest introgression from S. alterniflora (Baumel et al. 2002). S. ⫻ townsendii is a sterile F1 homoploid hybrid species that formed in Southampton Water in southeastern England in the nineteenth century (Gray et al. 1991). The parental species were S. alterniflora, introduced from America, and S. maritima, presumably native to Europe. The first notice of these homoploid hybrids noted was in Southampton Water, England, in 1870. It was given the name S. ⫻ townsendii (Marchant 1967),

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while other hybridizations of these parental species occurred in France and are named S. neyrautii (Ainouche et al. 2003). Only in Southampton Water did subsequent chromosomal doubling occur. It gave rise to the fertile, dodecaploid, allotetraploid species, S. anglica, around 1890 (Gray et al. 1991). S. anglica was spread and dispersed on its own to salt marshes throughout Britain (Raybould et al. 1991), France (Gray et al. 1991), and beyond. The introduced S. alterniflora is now extremely narrowly restricted in Britain, at Southampton Water at the hybridization site (Gray et al. 2001). S. alterniflora maintains small, vigorous populations at the three sites in France to which it was introduced in the early twentieth century (Baumel et al. 2002). The lack of spread could be due to pollen scarcity as S. alterniflora is largely self-incompatible (Davis, Taylor, Civille, et al. 2004; Davis, Taylor, Lambrinos, et al. 2004). S. ⫻ townsendii is rare today. It was not fully discriminated from S. anglica until the 1960s, after many years of exportation from Poole Harbor. The Spartina nursery at Arne had both forms in the sward in the 1950s. S. ⫻ townsendii grew together with S. anglica in the south and southeast of England and on the Isle of Wight (Goodman et al. 1969, fig. 3). While little information exists on export to most places except New Zealand, sterile S. ⫻ townsendii would have been outcompeted by fertile S. anglica in mixed swards. East Anglia marshes had both forms through the 1930s and 1940s (Gray et al. 1991). In a world survey, Ranwell (1967) grouped S. ⫻ townsendii with S. anglica under the rubric S. townsendii sensu lato (s.l.) and noted that by 1870, one or both of these hybrids had been spread about the United Kingdom. While seed of the allotetraploid S. anglica disperses on the tide, the sterile diploid sets no seed and could be spread only by humans or, rarely, by means of floating vegetative fragments that might break loose from an eroding bank. Globally, both were spread widely and covered more than twentyeight thousand hectares around the world when Ranwell (1967) recorded twenty-two successful 16

introductions and twenty-two failures. Specifically, for 1967, Europe saw thirteen success and no failures; Australia and New Zealand, eight successes and three failures; and Puget Sound, one success, no failures. The rest were apparent failures: India and Asia, seven; South Africa, two; the Red Sea, one; the Mediterranean, one; the western Atlantic, five; Greenland, one; British Columbia, one; and Hawaii, one (Ranwell 1967, fig. 1). The higher success rate in Europe was matched by wider spread there (in maximum estimated hectares in 1967): the United Kingdom, 12,000; France, 8,000; Netherlands, 5,800; Germany, 800; Denmark, 500; Ireland, 400; Tasmania and New Zealand, each 40; Australia, 20; and Puget Sound, less than 1. The second known hybridization of S. alterniflora with a native species occurred in San Francisco Bay after introduction of this Atlantic native in 1976 by the U.S. Army Corps of Engineers (Faber 2000). Most introductions of S. alterniflora to the Pacific were beyond the ranges of native Spartina species and posed no possibility of hybridization. The northern limit of the only native north temperate species in the Pacific, S. foliosa, California cordgrass, is Drake’s Estero, forty kilometers north of San Francisco Bay. Introductions north of San Francisco Bay, to Washington, Oregon, and northern California, were into regions with no native Spartina. California cordgrass, S. foliosa, was the native parent species of these hybrids. Although not known at the time, many of the hybrids were purposefully spread during the 1980s from the earliest site of hybridization, which was in pond 3, adjacent to Coyote Hills Slough (New Alameda Creek) at the southeastern end of San Francisco Bay. Seed floating on the tide spread the hybrids to many other marshes over eastern and western shores of the sixty kilometers of shoreline in the southern arm of the bay. A few hybrid colonies established in salt marshes of Marin County, on the north side of the Golden Gate. In the earliest published record of this invasion (Callaway and Josselyn 1992), cordgrasses assumed to be

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S. alterniflora, but probably hybrids, were found to be competitively superior to, and to have a wider tidal range than, the native S. foliosa. Hybrids were detected in the mid-1990s when plants were found that contained genetic material unique to each parent (Daehler and Strong 1997). The highly diverse nuclear and cytoplasmic composition together with chromosome numbers equal to or close to the parents’ suggested that these plants comprised a swarm of backcrossing hybrids rather than an allopolyploidization, as per S. anglica (Ayres et al. 1999; Anttila et al. 2000). Much of the shoreline and

most creeks flowing into the southern arm of San Francisco Bay were invaded by hybrid cordgrass by 2004 (see fig. 1.4). Native California cordgrass, S. foliosa, is shorter, with shallower roots, and grows less densely than vigorous hybrid genotypes (Daehler and Strong 1997). The vigorous subset of hybrid genotypes is transgressive; it grew larger and produced more inflorescences, pollen, and seed than either parent species (Ayres et al. 1999, 2003). Transgressively vigorous hybrids are now found at the two leading edges of the hybrid invasion. One leading edge is in the native marshes

Spartina patens Spartina anglica Spartina alterniflora/hybrid Spartina densiflora Coast line

N

0

12 km

FIGURE 1.4 Distribution of exotic cordgrasses in San Francisco Bay, 2001. From Ayres et al. 2004.

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of dense S. foliosa growing above mean sea level. In native marshes, hybrids spread by vegetative competitive displacement of the native species and by swamping native stigmas with pollen, which leads to hybrid seed. Where large hybrids are growing, most S. foliosa flowers set hybrid seed. The other leading edge of the invasion is below mean sea level on the vast, open, intertidal mud flats of the bay. While very few recruits of native California cordgrass appeared in open areas in the San Francisco estuary in recent years (Sloop 2005), the hybrid swarm increased tremendously in numbers and coverage in both native marshes and previously open mudflats. Both vegetative expansion and multiple episodes of seedling recruitment contribute to the invasion by hybrids. The area of San Francisco Bay covered by hybrids in 1975 was about two hectares. By 1990, about 650 circular plants expanding vegetatively were noted on aerial photos, and we estimated this to amount to five hectares. In 1993, the number of such hybrids had increased to one thousand, many had coalesced to form meadows, and we estimated the cover to be nearly ten hectares. In 2001, cover of hybrids was 190 hectares. While exponential growth in cover gives a semilog straight line, the semilog plot of the data for hybrid cover from these years gives a convex line. This indicates that the rate of spread of hybrid cordgrass has increased. The transgressive traits caused by hybridization undoubtedly contributed greatly to the very high rate of spread in San Francisco Bay. In China, intraspecific hybridization of multiple S. alterniflora populations may have created genotypes with a propensity to spread rapidly. Seeds and cutting of Spartina alterniflora from Morehead City, North Carolina, Altamaha Estuary, Georgia, and Tampa Bay, Florida, were sent to C. H. Chung at Nanjing University in China in 1979 (Chung et al. 2004). They were grown in the Botanical Garden their first year, and their growth was carefully monitored. There were striking differences among the populations in phenotypic characters; for example, the maximum height 18

of Georgia plants was almost three meters, while the largest Florida plants were half that height (An et al. 2004). Differences were also found in isozyme banding patterns, indicating genetic as well as phenotypic variation among the populations. In 1981, rooted cuttings from the nursery population were planted into a 1,300-square-meter field site at Luouyuan Bay. Ecotypic differences persisted in the field plantation. In 1985, a nursery was established in a village paddy near Chengmengkou using mixed seed collected from the Luouyuan Bay field site. Plants and/or seeds (references don’t say which) from the nursery were outplanted into three field sites for experimental monitoring. The allopatric S. alterniflora provenances cultured together in China could have readily hybridized. They were mixed within a small 1,300square-meter plot. The potentially hybrid seed was sown into a common paddy, and then a mixture of genotypes was selected, with preference for Georgia provenances, in the first large field trials. Adding to this genetic farrago, at some point 0.5 kilogram of seeds from North Carolina was introduced, and one source claims that it was this source that led to most of the salt marshes in coastal China (An et al. 2004). It is not unlikely that hybrids and the most vigorous progeny were spread widely. Intraspecific hybridization could well have played a role in the rapid spread of S. alterniflora thorough Chinese marshes. The high planting densities in the Chinese planting could have overcome Allee effects (Davis, Taylor, Civille, et al. 2004; Davis, Taylor, Lambrinos, et al. 2004). It is also possible that the rapid spread is a product of evolution of selfcompatibility, which occurred in the hybrids of S. foliosa ⫻ S. alterniflora in San Francisco Bay (Sloop 2005). No interspecific hybridization is known for Chinese S. alterniflora, but it is possible that the intraspecific hybridization of the three North American provenances produced self-compatible genotypes. Acknowledgments. We thank Daniel Goldstein of the University of California, Davis, library for help with the history of Reclamation, California; Tjeerd

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Bouma and Alan Gray for advice about S. anglica on the coast of the Netherlands and the United Kingdom, respectively; Kevin Rice, Richard Mack, and Spencer Barrett for coining “Arrive, Survive, and Thrive”; and three anonymous reviewers for helpful suggestions on the manuscript. T. B. C. Shaw provided inspiration. REFERENCES Ainouche, M. L., A. Baumel, and A. Salmon. 2004. Spartina anglica C. E. Hubbard: A natural model system for analysing early evolutionary changes that affect allopolyploid genomes. Biological Journal of the Linnean Society 82: 475–484. Ainouche, M. L., A. Baumel, A. Salmon, and G. Yannic. 2003. Hybridization, polyploidy and speciation in Spartina (Poaceae). New Phytologist 161: 165–172. Allen, J. R. L., and K. Pye. 1992. Coastal salt marshes: Their nature and importance. Pages 1–18 in J. R. Allen and K. Pye (eds.), Salt Marshes, Morphodynamics, Conservation and Engineering Significance. Cambridge: Cambridge University Press. An, S., C. Zhou, Z. Wang, Z. Deng, Y. Zhi, and L. Chen. 2004. Spartina in China: Introduction, history, current status, and recent research. Paper presented at the Third International Conference on Invasive Spartina, San Francisco, November 8–10. Anttila, C. K., R. A. King, C. Ferris, D. R. Ayres, and D. R. Strong. 2000. Reciprocal hybrid formation of Spartina in San Francisco Bay. Molecular Ecology 9: 765–770. Ayres, D. R., P. Baye, and D. R. Strong. 2003. Spartina foliosa—A common species on the road to rarity? Madroño 50: 209–213. Ayres, D. R., D. Garcia-Rossi, H. G. Davis, and D. R. Strong. 1999. Extent and degree of hybridization between exotic (Spartina alterniflora) and native (S. foliosa) cordgrass (Poaceae) in California, USA determined by random amplified polymorphic DNA (RAPDs). Molecular Ecology 8: 1179–1186. Ayres, D. R., D. L. Smith, K. Zaremba, S. Klohr, and D. R. Strong. 2004. Spread of exotic cordgrasses and hybrids (Spartina sp.) in the tidal marshes of San Francisco Bay, California, USA. Biological Invasions 6: 221–231. Bakker, J. P., P. Esselink, K. S. Dijkema, W. E. van Duin, and D. J. de Jong. 2002. Restoration of salt marshes in the Netherlands. Hydrobiologia 478: 29–51. Baumel, A., M. L. Ainouche, R. J. Bayer, A. K. Ainouche, and M. T. Misset. 2002. Molecular phy-

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Pethick, J. 2002. Estuarine and tidal wetland restoration in the United Kingdom: Policy versus practice. Restoration Ecology 10: 431–437. Pickart, A. 2001. The distribution of Spartina densiflora and two rare plant species in Humboldt Bay, California 1998–1999. Arcata, CA: U.S. Fish and Wildlife Service. ———. 2005. Control of Invasive Spartina densiflora in a High-Elevation Salt Marsh, Mad River Slough, Humboldt Bay National Wildlife Refuge. Arcata, CA: U.S. Fish and Wildlife Service. Ranwell, D. S. 1967. World resources of Spartina townsendii (sensu lato) and economic use of Spartina marshland. Journal of Applied Ecology 4: 239–256. Ranwell, D. S., and B. M. Downing. 1960. The use of Dalapon and substituted urea herbicides for control of seed-bearing Spartina (cord-grass) in intertidal zones of estuarine marsh. Weeds 8: 78–88. Raybould, A. F., A. J. Gray, M. J. Lawrence, and D. F. Marshall. 1991. The evolution of Spartina anglica Hubbard, C.E. (Gramineae)—Genetic-variation and status of the parental species in Britain. Biological Journal of the Linnean Society 44: 369–380. Redfern, R. 1983. The Making of a Continent. Toronto: Fitzhenry & Whiteside. Redfield, A. C. 1972. Development of a New England salt marsh. Ecological Monographs 42: 201–237. Reichman, O. J., and E. W. Seabloom. 2002. The role of pocket gophers as subterranean ecosystem engineers. Trends in Ecology and Evolution 17:44–49. Riese, K. 2005. Coast of change: Habitat loss and transformations in the Wadden Sea. Helgoland Marine Research 59: 9–21. Riggs, S. 2005. Survey and Control of Spartina in Padilla Bay in 2005. Mount Vernon: Washington State Department of Ecology, Shorelands and Environmental Assistance Program. Riggs, S., and D. A. Bulthuis. 1994. Estimated Net Aerial Primary Productivity and Monitoring of Selected Characteristics of Spartina alterniflora in Padilla Bay, WA, April 1992–May 1993. 94-176, Technical Report no. 11. Mount Vernon: Washington State Department of Ecology, Shorelands and Environmental Assistance Program. Salley, H. E. 1977. History of California Post Offices, 1849–1976. La Mesa, CA: Postal History Associates. Sayce, K. 1988. Introduced Cordgrass, Spartina alterniflora Loisel, in Saltmarshes and Tidelands of Willapa Bay, WA. USFWS-87058 (TS). Ilwaco, WA: Willapa National Wildlife Refuge, U.S. Fish and Wildlife Service. Siegel, S. W., and P. A. M. Bachland, eds. 2002. Feasibility Analysis: South Bay Salt Pond Restoration San Francisco Estuary, California. San Rafael: California Wetlands and Water Resources. 22

Sloop, C. M. 2005. Reproductive and recruitment dynamics of invasive hybrid cordgrasses (S. alterniflora ⫻ S. foliosa) in San Francisco Bay tidal flats. Unpublished PhD diss., University of California, Davis. Spicher, D., and M. Josselyn. 1985. Spartina (Gramineae) in northern California: Distribution and taxonomic notes. Madroño 32: 158–167. Stapf, O. 1908. Spartina townsendii. Gardener’s Chronicle 43: 33–35. Stockstad, E. 2005. Louisiana’s wetlands struggle for survival. Science 310: 1264–1266. Stralberg, D., V. Toniolo, G. W. Page, and L. E. Stenzel. 2004. Potential Impacts of Non-native Spartina Spread on Shorebird Populations in South San Francisco Bay. Stinson Beach, CA: Point Reyes Bird Observatory. Taylor, C. M., H. G. Davis, J. C. Civille, F. S. Grevstad, and A. Hastings. 2004. Consequences of an Allee effect in the invasion of a pacific estuary by Spartina alterniflora. Ecology 85: 3254–3266. U.S. Army Corps of Engineers. 1976. Dredge Disposal Study San Francisco Bay and Estuary: Appendix K, Marshland Development. San Francisco: Author. ———. 1978. Shoreline Erosion Control Demonstration Program, Alameda, California: Preconstruction Report. San Francisco: Author. Warren, R. S., P. E. Fell, R. Rozsa, A. H. Brawley, A. C. Orsted, E. T. Olson, V. Swamy, and W. A. Niering. 2002. Salt marsh restoration in Connecticut: 20 years of science and management. Restoration Ecology 10: 497–513. Wells A. 1995. Rice grass in Tasmania: An overview. Pages 11–13 in J. E. Rash, R. C. Williamson, and S. J. Taylor (eds.), Proceedings of the Australasian Conference on Spartina Control. Melbourne: Victorian Government Publications. Williams, G., J. Baumann, D. Buffett, and et al. 2004. Discovery and management of Spartina anglica in the Fraser River estuary, British Colombia, Canada. Paper presented at the Third International Conference on Invasive Spartina, November 8–10, San Francisco. Williams, P., and P. Faber. 2001. Salt marsh restoration experience in San Francisco Bay. Journal of Coastal Research 27: 203–211. Xie, Y., Z. Y. Li, W. P. Gregg, and L. Dianmo. 2001. Invasive species in China: An overview. Biodiversity and Conservation 10: 1317–1341. Zhang, R. S., Y. M. Shen, L. Y. Lu, S. G. Yan, Y. H. Wang, J. L. Li, and Z. L. Zhang. 2004. Formation of Spartina alterniflora salt marshes on the coast of Jiangsu Province, China. Ecological Engineering 23: 95–105.

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Changes in Community Structure and Ecosystem Function Following Spartina alterniflora Invasion of Pacific Estuaries Edwin D. Grosholz, Lisa A. Levin, Anna C. Tyler, and Carlos Neira One of the most pervasive human impacts to salt marshes around the world is the introduction of nonnative species. Plant introductions to salt marsh systems have resulted in significant changes ranging from species replacement to broad-scale alteration of ecosystem properties. In this chapter, we examine the changes produced by the invasive Atlantic smooth cordgrass Spartina alterniflora in two wetland ecosystems, one in San Francisco Bay, California, and one in Willapa Bay, Washington. We compare and contrast the impacts of Spartina invasion on a range of processes, focusing primarily on benthic invertebrate communities and the sediment environment. Our work shows that the structure of the plant itself and the changes it produces in the physical and chemical environment strongly influence benthic communities. The aboveground structure shades the substrate, reducing photosynthesis of benthic microalgae and restricts water flow, which contributes to decreased recruitment and slower growth of benthic suspension feeders. At the same time, the plant structure increases sedimentation of fine-grained particles and leads to the accumulation of detritus and low-quality organic matter. These changes, together with increased belowground plant biomass and peat production, influence sediment chemistry and metabolism. In addition, belowground plant biomass can preempt substantial amounts of belowground habitat, directly reducing benthic abundance and diversity. These physical and chemical changes can result in dramatic shifts in benthic communities including substantial reductions in larger, surfacefeeding taxa concurrent with increases in smaller, subsurface detritivores. Such shifts in the functional identity of benthic species may negatively affect feeding of both invertebrate and vertebrate predators. The magnitude and direction of these various changes depend on the habitat invaded and are generally more substantial where Spartina has invaded open mudflats and less where Spartina invades areas of native vegetation. We discuss the consequences of these changes in the context of plant effects in other coastal systems in order to develop a set of predictions regarding future plant invasions. The ultimate consequences of plant invasion on benthic communities involve a dynamic balance of positive and negative effects. 23

By understanding the mechanisms that determine the magnitude and direction of these effects, scientists and managers will be better able to understand the consequences of human-induced plant invasions in salt marsh systems and to predict the impacts of future introductions.

Salt marshes are among the many habitats worldwide experiencing rapid change as the result of human-mediated stressors (Vitousek et al. 1997; Chapin et al. 2000; Mooney and Hobbs 2000). Among the most important changes experienced by salt marshes is the introduction of nonnative plants. Wetlands generally are susceptible to invasion, and introduced plants can foster changes acting broadly across multiple spatial and temporal scales, producing long-lasting and system-wide effects (Zedler and Kercher 2004). These changes may include a host of physical and chemical alterations that can influence populations and communities of plants and animals as well as alter the storage and recycling of nutrients (Mack et al. 2000). Introduced plants can also directly affect the abundance, diversity, and functional identity of communities via the provision or preemption of physical habitat and as well as changes in the basis of trophic support. The invasion of Spartina alterniflora in U.S. West Coast estuaries provides an excellent demonstration of the extent to which plant invasions can rapidly and extensively alter community and ecosystem dynamics in salt marsh systems. In this chapter, we will compare and contrast the invasions of the Atlantic smooth cordgrass Spartina alterniflora in two estuaries in western North America. The first is San Francisco Bay, California, where S. alterniflora hybridized with the native Spartina foliosa, and the resulting hybrid has spread rapidly throughout the central and southern portions of San Francisco Bay. The second is Willapa Bay, Washington, where in the absence of any native Spartina, S. alterniflora has colonized habitats throughout most of this bay. We will address the impacts of Spartina on benthic systems in these two regions with a par-

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ticular focus on infaunal and epifaunal communities and the changes in physical and chemical processes in the sediment environment. The changes brought about by Spartina are ultimately the result of the extensive above- and belowground growth of the plant. But the overall effects on community and ecosystem processes depend on the magnitude of change in a range of properties, including above- and belowground vascular plant biomass, production and decomposition of detritus, microalgal productivity, and sediment physicochemistry, all of which can have variable effects on the native plants and animals. The consequences of Spartina invasion strongly depend on the age of the invasion, the habitat being invaded, the presence or absence of native vegetation, and many site-specific characteristics including elevation, sediment grain size, and hydrodynamic regime. The primary goal of this chapter is to synthesize recent studies of Spartina invasions in Washington and California to better understand and predict the impacts of Spartina and other marsh plants in the future. We do this by comparing our results with other studies of invasive Spartina in this western region and other regions, as well as with systems where Spartina is native. An additional goal is to use this understanding to assist ongoing eradication programs in California and Washington that aim to eliminate invasive Spartina from this region. By quantifying the full extent of the changes being brought about by the Spartina invasion, as well as knowing the mechanisms underlying the changes, we hope to assist resource managers in determining when, where, and how eradication of Spartina will be most effective.

invasions in north american salt marshes

CONTRASTING INVASIONS The invasion of Spartina alterniflora in San Francisco Bay, henceforth “SFB,” began in 1975 with its intentional introduction from the Atlantic Coast of the United States by the Army Corp of Engineers for salt marsh restoration (Ayres, Strong, and Baye 2003; Ayres, Strong, et al. 2004). Following the initial introduction, a hybrid developed from S. alterniflora and the native S. foliosa (hereafter referred to as hybrid Spartina) (Daehler and Strong 1997). This hybrid grows more vigorously than either parent plant and has a higher and lower tidal range than native plants (Ayres et al. 2003; Ayres, Smith, et al. 2004; Ayres, Zaremba, and Strong 2004). The hybrid has successfully colonized nearly eight hundred hectares of the central and southern portions of SFB, where it has invaded open mudflat at the lower end of its tidal distribution. Although three other species of Spartina have been introduced into SFB (S. densiflora, S. anglica, and S. patens), and another hybrid (S. densiflora ⫻ S. foliosa) has been found (D. Ayers, personal communication), none have spread significantly (California Coastal Conservancy 2006). The hybrid has displaced the native S. foliosa as well as the original parent S. alterniflora in central SFB through hybridization and overgrowth (Ayres et al. 2003; Ayres, Smith, et al. 2004; Ayres, Zaremba, and Strong 2004) and competes with native plants such as Sarcocornia pacifica, Distichlis spicata, and Jaumea carnosa. The invasion of Willapa Bay, henceforth “WB,” began with the accidental introduction of S. alterniflora around 1890 (Feist and Simenstad 2000; Davis et al. 2004; Civille et al. 2005). Throughout most of the century since its initial introduction, the spread of Spartina has been fairly gradual. However, from 1995 to 2005, it rapidly spread to cover 1,600 hectares in WB for unknown reasons. Because there are no native Spartina species in the Pacific Northwest with which to hybridize, this invasion is entirely the result of the spread of S. alterniflora. As in SFB, S. alterniflora has rapidly colonized open mud-

flat and has been found among native plants in the upper intertidal zone, where it has the potential to compete with and even exclude species such as Sarcocornia pacifica, Distichlis spicata, and Deschampsia caespitosa, although this has not so far been demonstrated. While both systems involve invasion of mudflats at the lower tidal range, noteworthy differences may further influence the impact of Spartina. In WB, the invasion is much older, the aerial cover is much greater, and other ecosystems, such as oyster reefs, are at risk of displacement. In addition, the invasive Spartina does not resemble any native marsh plant forms, and since there were no native Spartina species present prior to the invasion, none of the native plants or animals have coevolved with this genus. In contrast, the SFB invasion is relatively recent and involves a hybrid that became a superior competitor that threatens to eliminate the pure native S. foliosa (Ayres et al. 2003). The system that is adapting to the invasion in SFB is replete with birds, fish, and invertebrates that have evolved with native Spartina in upper intertidal elevations. SFB is also heavily invaded, and because many of these introductions came from the Atlantic Coast, these invaders have a long evolutionary history with the presence of Spartina across a broad tidal range (Cohen and Carlton 1998). These include bivalves such as Geukensia demissa, Mya arenaria, and Gemma gemma as well as gastropods such as Urosalpinx cinerea and Ilyanassa obsoleta; these are among the most common species in areas of SFB with invasive Spartina.

CONSEQUENCES OF INVASION VASCULAR PLANT PRODUCTION

The productivity of Spartina, both above- and belowground, was greater than that of native vascular plants at both the WB and SFB sites. The magnitude of the differences between native and nonnative habitats varied depending on the initial conditions (vegetation type and substrate) and the age of the invasion. In SFB, the

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TABLE 2.1 Summary of studies to date examining changes in primary production, sediment respiration, and sediment organic C and N in invaded western salt marshes

san francisco bay

AG biomass (g m⫺2)1,2 BG biomass (g m⫺2)1,2 Benthic Chl a (mg m⫺2)3 GPP (mg Cm⫺2 d⫺1)3 Respiration (mg Cm⫺2 d⫺1)3 %N4 %C4 C:N4

willapa bay

Mudflat

Spartina

Mudflat

Spartina

(mean ⫾ SE) ——— ——— 184 ⫾ 51 459 ⫾ 92 179 ⫾ 34 0.10 ⫾ 0.03 1.22 ⫾ 0.30 14.8 ⫾ 2.1

(mean ⫾ SE) 1823 ⫾ 283 4197 ⫾ 773 97 ⫾ 15 258 ⫾ 52 752 ⫾ 91 0.17 ⫾ 0.02 2.70 ⫾ 0.29 21.8 ⫾ 2.2

(mean ⫾ SE) 3⫾1 55 ⫾ 8 117 ⫾ 9 ——— ——— 0.12 ⫾ 0.00 1.35 ⫾ 0.04 12.9 ⫾ 0.1

(mean ⫾ SE) 1695 ⫾ 245 878 ⫾ 91 247 ⫾ 23 ——— ——— 0.23 ⫾ 0.02 3.36 ⫾ 0.34 16.0 ⫾ 0.6

NOTE: Aboveground (AG) and belowground (BG) biomass were measured at the end of the growing season (September–October). SOURCES: 1Tyler et al. 2007 (Spartina values); 2mudflat biomass values are for the invasive intertidal seagrass Zostera japonica from A. C. Tyler, unpublished data, 2002; 3Tyler and Grosholz (n.d.), except for Willapa Bay chlorophyll a values, which are from A. C. Tyler, unpublished data, 2003); 4A. C. Tyler, unpublished data, fall 2002.

aboveground biomass of invasive Spartina was significantly higher than the biomass of native S. foliosa and Sarcocornia sp. (see table 2.1; Brusati and Grosholz 2006; Tyler, Lambrinos, and Grosholz 2007). The belowground biomass of invasive Spartina was also generally higher than native plants including S. pacifica, although the difference was not as great as for aboveground biomass (table 2.1; Brusati and Grosholz 2006, Tyler et al. 2007). Also, the biomass of invasive Spartina was obviously much greater than the micro- and macroalgae on historically “unvegetated” mudflats. The belowground biomass of hybrid Spartina in SFB (top six to ten centimeters) changes with invasion stage. For example, at one site it increased from unvegetated tidal flats (less than one hundred grams per square meter) to 348 grams per square meter on the “growing edge” (about one meter inside vegetation—i.e., young invasion), to over 720 grams per square meter in the central area (thirty-year-old invasion) (Neira et al. 2007). In WB, the introduced eelgrass Zostera japonica overlaps considerably in the tidal zone colonized by S. alterniflora (fig. 2.1). This eelgrass apparently produces only about a third of the biomass produced by S. alterniflora (Ruesink et al. 2005). The long history of invasion in WB 26

FIGURE 2.1 Introduced Spartina alterniflora seedlings competing for space with introduced eelgrass Zostera japonica in Willapa Bay, Washington.

has allowed us to determine the dynamics of Spartina production by using the different-aged meadows as a chronosequence to represent the trends in biomass through time. Based on a series of eighteen S. alterniflora meadows ranging from less than one year to more than thirty years old (age was determined based on the date when the areal extent of cordgrass in each region was greater than 50 percent), we found that aboveground biomass reached a peak within five to ten years and then declined as the meadow aged. In contrast, the buildup of a dense mat of roots and rhizomes required a longer time (ten to

invasions in north american salt marshes

fifteen years) but then remained relatively constant (A. C, Tyler, unpublished data). This aboveground biomass of S. alterniflora in WB was similar to that of hybrid Spartina in SFB; however, the belowground biomass in WB was considerably less than in SFB (table 2.1). PHYSICAL PROCESSES

The large amount of aboveground biomass produced by invasive Spartina necessarily results in changes in the physical regime of the invaded mudflats. The plant reduces both the light reaching the mudflat surface and tidal energy. In both SFB and WB, the aboveground canopy of invasive Spartina significantly reduced light penetration to the sediment surface by up to 83 percent (Neira, Levin, and Grosholz 2005; Neira et al. 2006; Tyler and Grosholz, forthcoming). Also, the plant canopy significantly restricted water flow and reduced velocity by up to 76 percent and bed stress relative to unvegetated areas (Neira et al. 2005, 2006). These reductions in light and flow produced by the invasive Spartina canopy were greater than those of native S. foliosa (Brusati and Grosholz 2006; Tyler and Grosholz, forthcoming), but they were comparable to those produced by Sarcocornia pacifica, which can result in more substantial reduction in light than Spartina (Neira et al. 2005; Tyler and Grosholz, forthcoming). The aboveground structure of Spartina invasion also led to increased sediment accretion rates and reduced sediment grain size (Neira et al. 2006). Together with the buildup of belowground biomass and peat, this has resulted in increased tidal elevation of invaded areas. These changes are common to both SFB and WB. In areas of even moderate flow, unvegetated mudflats experience rapid changes in elevation of up to four centimeters over periods as short as a few weeks in comparison with changes usually much less than one centimeter in vegetated areas. By reducing flow, Spartina attenuated these short-term changes in elevation resulting in a pattern of gradual increase in elevation relative to the adjacent unvegetated areas (E. D. Grosholz, unpublished data).

MICROALGAL PRODUCTION

The Spartina invasion has a substantial impact on benthic microalgal photosynthesis. In many intertidal habitats, benthic microalgal production comprises a substantial proportion of overall system primary production and is an important contributor to local food webs (Zedler 1980, 1984; Page 1995; Deegan and Garritt 1997; Kwak and Zedler 1997; Page 1997). This is especially true on Pacific Coast mudflats and in Sarcocornia-dominated marshes, where microalgae are the primary carbon source for many secondary consumers (Kwak and Zedler 1997; Moseman et al. 2004). The impact of Spartina invasion on benthic microalgae was variable across sites and is dependent to some extent on the nature of the uninvaded habitat. Where Spartina had invaded mudflats in both SFB and WB, benthic chlorophyll a concentration (a proxy for microalgal biomass) was often, but not always, higher in Spartina meadows than in the adjacent uninvaded mudflats (table 2.1; Neira et al. 2005, 2006; Tyler and Grosholz, forthcoming; A. C. Tyler et al., unpublished data). However, microalgal production was consistently higher on the mudflats of SFB than in the adjacent Spartina marshes. Factors such as light availability and grazing (see “Food Web Structure”) may create a disparity between chlorophyll a–based standing stock and overall productivity. Overall, the Spartina invasion had a negative impact on microalgal productivity with important ramifications for overall food web support. NUTRIENTS AND ORGANIC MATTER

The invasion of Spartina resulted in a shift in the overall cycling of organic matter and nutrients because of the refractory nature (higher carbon:nitrogen) of its tissues (A. C. Tyler et al., unpublished data). Following the invasion of WB, we observed an extensive accumulation of detritus associated with S. alterniflora litter that occurs primarily at the end of the growing season when the aboveground portion died back. This detritus accumulated in massive heaps in the high intertidal, where it slowly decomposed

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and had detrimental effects on native vegetation plant biomass, aboveground productivity, and species richness (A. C. Tyler et al., unpublished data; Lambrinos, forthcoming). In SFB, because hybrid Spartina decomposes more slowly than native S. foliosa, more wrack build-up can be attributed to the hybrid (A. C. Tyler et al., unpublished data). In addition to the high aboveground litter production, belowground production generates large volumes of roots, rhizomes, and peat, leading to an increase in belowground organic matter and particulate organic carbon and nitrogen relative to the native systems in both SFB and WB (table 2.1; Neira et al. 2006; Levin, Neira, and Grosholz 2006; A. C. Tyler et al., unpublished data). We observed a subsequent increase in meadow sediment respiration rates relative to mudflats in SFB that fostered a shift from a net autotrophic system, dominated by microalgal production, to a new heterotrophic system, dominated by belowground microbial decomposition (table 2.1; Tyler and Grosholz, forthcoming). However, we found higher sediment respiration rates in native vegetation than in hybrid Spartina, most likely as a result of the more refractory nature of the Spartina detritus that is a less desirable substrate for microbial decomposers (Tyler and Grosholz, forthcoming). In all cases, the turnover of carbon and nutrients is slowed down, resulting in the accumulation of massive quantities of highly refractory organic matter. In contrast to the dramatic increase in sediment particulate nitrogen upon invasion of unvegetated mudflats, we found a significant decrease in available porewater ammonium in invaded meadows compared to native mudflats in both SFB and WB (A. C. Tyler, unpublished data). Possible causes for this decrease are plant uptake, increased microbial immobilization, or increased coupling of nitrification-denitrification in the oxidized root zone of Spartina. On the Atlantic coast, the invasion of Phragmites australis also caused distinct changes in nitrogen cycling in sediments depending on the invaded community (Windham and Meyerson 28

2003). Even though hybrid Spartina is able to increase the redox potential in the vicinity of its roots, we observed an overall decrease in redox potential (Neira et al. 2005). Accordingly, our results from SFB also show that porewater sulfide concentrations are nearly one hundred times greater in Spartina meadows relative to uninvaded mudflats (meadow, 1,016 ⫾ 121 ␮M; mudflat, 12 ⫾ 1 ␮M; mean ⫾ SE; A. C. Tyler, unpublished data; Neira et al. 2007), which is likely due to the increased organic matter pool and subsequent increase in both microbial oxygen demand and sulfate reduction. BENTHIC COMMUNITIES

The altered physical and chemical environments can also strongly influence the recruitment, survival, growth, and reproduction of benthic invertebrates in the invaded areas. Both epifaunal and infaunal organisms respond to the changing light, flow, and sediment conditions in ways that affect not only the biomass and diversity of these organisms, but also the functional identity of many groups (Neira et al. 2006). Our work has shown significant changes in the recruitment and growth of common taxa in the hybrid Spartina meadows relative to uninvaded areas. Measurements of barnacle (Balanus glandula) recruitment on standardized recruitment substrata (mussel shells) placed in the growing edge of hybrid Spartina meadow and on adjacent mudflat areas showed recruitment was ninefold higher on the mudflats (Neira et al. 2006). Experiments using marked Macoma petalum transplanted into hybrid Spartina meadows, and adjacent mudflat areas showed growth was significantly reduced by up to twofold in the hybrid Spartina treatments (Brusati and Grosholz 2007). Our results suggest reductions in flow; thus, reduced delivery of propagules (recruitment) and seston (for growth) likely contribute to these patterns. Our work has also shown that both biomass and diversity of infauna can decline strongly in invaded areas relative to the previously

invasions in north american salt marshes

unvegetated control areas (Neira et al. 2005, 2007; Levin et al. 2006; Neira et al., forthcoming). In SFB, invertebrate densities decline by as much 75 percent relative to unvegetated mudflat (Neira et al. 2005), although densities at some sites are not significantly different from those in native Sarcocornia pacifica areas. Species richness also showed a significant 25 percent decline in Spartina invaded areas compared with unvegetated controls, but it was elevated relative to Sarcocornia-vegetated habitat (Neira et al. 2005). Similar patterns were measured in WB, where invertebrate species richness was significantly higher in mudflat/Zostera areas than in S. alterniflora areas (mudflat, 11.75 ⫾ 2.62; Spartina, 6.0 ⫾ 3.12; mean ⫾ 1 SE; per 19.6 square centimeters). Species diversity was also higher in the mudflat/Zostera areas relative to the Spartina meadow (Shannon-Wiener H⬘ [log2] ⫽ 3.92, mudflat; 2.79, Spartina), and the density of invertebrates was slightly greater in mudflat/Zostera areas, although not significantly so. In SFB, we experimentally investigated the mechanisms underlying these changes in a series of transplant experiments, plant manipulations, predator inclusion and exclusion studies, and trophic investigations. This work demonstrated that the changes in the density and diversity of infauna were due to the combined effects of preemption of belowground habitat by Spartina, changes in the food supply and predation pressure, as well as the physical and chemical changes to sediments and porewater (Neira et al. 2006; Levin et al. 2006). For these reasons, the effects of Spartina on benthic abundance and diversity are less dramatic in areas previously occupied by native marsh plants (Neira et al. 2005). FACILITATING OTHER INVASIVE SPECIES

The aboveground structure of Spartina can have other positive effects, including facilitating higher densities of other invasive species. In SFB, our data suggest that the invasive hybrid has facilitated several nonindigenous inverte-

brate species, creating invasive hot spots (areas of greater abundance) in the recently colonized Spartina edge areas not present on the open mudflat. At one site in SFB, we found that densities of several clams introduced from the Atlantic (Macoma petalum, Mya arenaria, Geukensia demissa) were two to ten times higher in the growing edge of the hybrid Spartina meadow than on the adjacent open mudflat (E. D. Grosholz et al., unpublished data). Although high Geukensia densities were largely due to the requirement of structure for attachment provided by Spartina (similar structure is virtually absent on open mudflats), densities of Mya and Macoma may be higher as the result of refuge from predation by bat rays (Myliobatus californica). At one site where greater than 40 percent of the substrate was disturbed by bat rays, densities of Mya and Macoma were two to three times greater in the growing edge of Spartina compared with adjacent mudflats. At a second site with less than 1 percent of the area disturbed by bat rays, there was no difference in clam densities between Spartina and mudflats (E. D. Grosholz et al., unpublished data). In SFB, we also found high abundances of introduced Atlantic gastropods Urosalpinx cinerea and Ilyanassa obsoleta along the leading edge of the hybrid Spartina. In experiments where we simulated the aboveground structure of Spartina by adding wooden dowels to mudflat areas, we found twenty- to fiftyfold increases in the density of both Urosalpinx and Ilyanassa in dowel addition treatments compared with adjacent mudflat control areas (E. D. Grosholz et al., unpublished data). Data documenting sediment temperatures in dowel addition areas suggest that lower sediment temperatures may strongly influence these patterns. We have also documented that the hybrid facilitates the European green crab Carcinus maenas in SFB. Our data show that this species is three to five times more abundant, particularly for smaller size classes, in hybrid in comparison with the adjacent mudflat (E. D. Grosholz et al., unpublished data; Neira et al. 2006). This predator, in turn, appears to have a significant

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TABLE 2.2 A comparison of the effects of both above- and belowground structure for invasive versus native Spartina on overall abundance and diversity of benthic invertebrates relative to unvegetated mudflats

invasive spartina

Aboveground Belowground

native spartina

Growing Edge

Developed Meadow

Growing Edge

Developed Meadow

Increase Slightly decrease

Increase Decrease

Increase Minimal

Increase Slightly decrease

NOTE: The growing edge areas for both are areas colonized within the last five years approximately and the developed meadows areas that have had vegetation for at least ten years (from Brusati and Grosholz 2006; Neira et al. 2007).

impact on the invertebrate taxa that are negatively affected by Spartina invasion and so may reinforce the changes brought about by this plant invasion (Neira et al. 2006). The extent to which Spartina facilitates other invasive species in WB is less certain. However, evidence suggests that Spartina fills a similar role facilitating younger size classes of green crabs (Yamada 2001). We summarize the generalized effects of above- and belowground portions of Spartina on invertebrate diversity and abundance in table 2.2. The growing edge of the hybrid Spartina meadow that had been colonized within approximately five years functions similarly to the native S. foliosa and has a similar influence on invertebrate abundance and diversity. In this growing margin, the aboveground structure has the same generally positive effects on invertebrates, but the belowground portion of the hybrid has not developed sufficiently to reduce infaunal densities. In areas farther back from the growing edge, which are approximately five to ten years old, changes in sediment biogeochemistry and belowground biomass have progressed to the point of significantly depressing densities and diversity of infauna (Neira et al. 2005). Therefore, we suggest that the influence of hybrid Spartina on benthic communities varies with the successional stage of the invasion (table 2.2; Neira et al. 2007). FOOD WEB STRUCTURE

The structure of benthic food webs was also affected by the changes in vascular plant and litter 30

availability relative to other food sources. We used stable isotope enrichment experiments to follow consumption of Spartina detritus and microalgae by invertebrates and to determine the broader effects of Spartina invasion on benthic food web structure. Using 15N-labeled Spartina detritus and 13C-labeled microalgae simultaneously, we found that the Spartina invasion in SFB has dramatically shifted the tidal flat infaunal invertebrate community from one dominated by surface feeders that primarily consume microalgae (amphipods, bivalves) to one dominated by belowground feeders that primarily consume plant detritus (oligochaetes, capitellid polychaetes) (Levin et al. 2006). Using two- and three-end-member mixing models, we determined that a significant portion of the diet of these belowground detritivores was the detritus produced by hybrid Spartina. Therefore, this invasion has resulted in a qualitative shift of the entire food web from one based on fresh primary production to one based on detritus. We summarize the overall results of these changes on invertebrate functional groups in table 2.3. When we compare these results with those from native Spartina areas that have not been invaded, we note some important differences. Natural abundance stable isotope studies conducted in native versus hybrid Spartina areas show little incorporation of the added detritus into the food web. Larger epifauna in hybrid areas did not show ␦ 13C values consistent with the added detrital input of Spartina when compared with organisms collected from open mudflat habitats, though there was a slight

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TABLE 2.3 A summary of predicted effects of aboveground and belowground invasive Spartina on densities of different functional groups of benthic invertebrates

Mobile epifauna Surface-feeding infauna Detrital-feeding infauna Bivalves

Aboveground

Belowground

Positive Negative Positive Positive/negative

None Negative Negative Negative

SOURCES: From Neira et al. (2005, 2006); Levin et al. (2006); Brusati and Grosholz (2006, 2007); E. D. Grosholz, unpublished data.

isotopic shift toward Spartina in invertebrates collected in native S. foliosa meadows compared with those on the open mudflat (Brusati and Grosholz 2007). This is also consistent with the idea that the detritus of S. foliosa is less refractory than that of the hybrid Spartina detritus, and benthic invertebrates may consume more detritus of the native than invasive Spartina. However, a comparison of ingestion of 15N-labeled S. foliosa and 15N-labeled hybrid Spartina in SFB indicated mostly similar infaunal uptake (oligochaetes and capitellid polychaetes) or avoidance patterns (Levin et al. 2006). We investigated food web changes in WB using similar isotopic enrichment experiments. We found that the taxonomic range of benthic species that consumed detritus of Spartina alterniflora was similar to SFB (E. D. Grosholz et al., unpublished data). We found that tanaids, bivalves, spionid and cirratulid polychaetes, and chironomid larvae all showed enriched signals in treatments with Spartina detritus in the Spartina zone (E. D. Grosholz et al., unpublished data). Based on elevated ␦ 15N values (30 to 80 per mille), we found significantly more Spartina detritus was consumed in the Spartina zone than in mudflat/Zostera zone treatments. As in the SFB study, we found that ␦ 13C values (15 to 50 per mille) were higher for surface feeders including amphipods, tanaids, and cirratulid polychaetes, with values similar in both Spartina and mudflat/Zostera zone treatments. Although we see some differences between food web impacts in SFB and WB, these do not appear to be due to differences in

palatability since the carbon:nitrogen ratios of S. alterniflora relative to the hybrid are similar (WB ⫽ 46.0 ⫾ 1.9; SFB ⫽ 48.4 ⫾ 1.8, and there was no difference in percent nitrogen; A. C. Tyler, unpublished data). Instead, the differences may reflect the much longer history of invasion in WB (nearly one hundred years) or the presence of extensive Zostera cover and detritus input in what was historically an unvegetated mudflat. We found that the detritus of the introduced Z. japonica was consumed to a much greater extent than Spartina in the WB experiments based on enriched ␦ 15N signatures. A wide range of surface-feeding consumers including several bivalves and terebellid polychaetes showed elevated ␦ 15N values with much greater values (100 to 500 per mille), supporting the higher edibility of Zostera relative to Spartina based on the lower carbon:nitrogen ratio (Z. japonica carbon:nitrogen 17.3 ⫾ 0.7; A. C. Tyler, unpublished data). VERTEBRATE CONSUMERS

The expansion of hybrid Spartina has also influenced grazing of cordgrasses by Western Canada geese (Branta canadensis moffitti) in SFB. Our studies have shown that the geese can distinguish differences in blade toughness between the native and hybrid forms of Spartina and completely avoid the hybrid while readily consuming the native S. foliosa. Data from both field studies and feeding trials using captive geese have shown that geese will graze more than 90 percent of the S. foliosa biomass at some sites while grazing less than 1 percent of the hybrid (E. D. Grosholz et al., unpublished data).

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FIGURE 2.2 Hybrid Spartina colonizing open mudflat and preempting foraging habitat used by wintering populations of migratory shorebirds in San Francisco Bay.

The Spartina invasion may also have important implications for wintering populations of migratory shorebirds. Most shorebirds will not forage in vegetated areas (Page, Stenzel, and Kjelmyr 1999; Stenzel et al. 2002), so the colonization of naturally unvegetated mudflats by invasive Spartina is a de facto loss of foraging habitat (fig. 2.2). Also, our data show that the biomass of invertebrates is greater in tidal flat areas at higher tidal elevations (Christiansen et al., forthcoming). Therefore, as the Spartina invasion expands to occupy a greater portion of these higher-elevation areas, shorebirds will be increasingly forced to forage at lower tidal elevations, which are not only exposed for shorter periods of time, but also contain lower densities of invertebrate prey. The consequences of Spartina invasion for shorebirds apply to WB as well. Although SFB is a more important area for wintering shorebirds, WB and adjacent Gray’s Harbor represent important winter grounds in that region for shorebirds (Page et al. 1999). Nonetheless, the same overall consequences of habitat loss apply, although we do not have similar data on the abundances of benthic invertebrates at different tidal elevations. The Spartina influence 32

in WB is also complicated by the extensive invasion of lower tidal elevations by the Japanese eelgrass Z. japonica, which during the summer months covers virtually the entire tidal range from the lower border of Spartina to the upper border of the native Zostera marina. In contrast to shorebirds that forage on tidal flats at low tide, estuarine fish forage in these same areas at high tide. Fish such as Chinook salmon, California halibut, rainwater killifish, striped bass, and Tule perch could potentially experience the same habitat losses and density effects in SFB and WB as the shorebirds.

PREDICTING IMPACTS OF INVASION We can identify three distinct invasion scenarios where Spartina effects appear to differ. The first is the invasion of unvegetated, higher-energy sand flats, which results in the largest changes in flow rates, sediment accumulation, elevation increase, and litter buildup, as well as the largest changes in plant and animal communities. This type of invasion has occurred in portions of both SFB and WB. A second type involves invasion of existing vegetation in high marsh habitats, which results in less alteration

invasions in north american salt marshes

FIGURE 2.3 Conceptual model of changes following Spartina invasion. The many changes brought about by the Spartina invasion are illustrated in this diagram showing positive (⫹) and negative (⫺) effects on a range of physical, chemical, and biological processes. Spartina reduces light and tidal water flow and increases deposition of fine sediments and sediment organic matter. The increased belowground biomass together with changes in the sediments results in increased levels of sulfide and decreased ammonium levels, which can negatively affect other plants and animals. These changes result in increasing detrital loads and decreasing productivity of benthic microalgae, which favor subsurface deposit feeders and detritivores at the expense of grazers, suspension feeders, and surface deposit feeders. The shift toward smaller, subsurface species may reduce invertebrate food resources for larger consumers such as crabs, birds, and fishes.

of benthos and in some cases leads to enhanced diversity of associated fauna (which are often invasives as well) (Neira et al. 2005). Under these circumstances, displacement of native plants may occur through light/space or nutrient competition (Tyler et al. 2007). A third mode of invasion involves hybridization with native Spartina, threatening the integrity of native Spartina species (Ayers et al. 2003) and altering the above- and belowground habitat structure. Understanding that Spartina invasions can yield different consequences under different circumstances can help identify the most urgent conservation needs (Thompson 1991; Hacker et al. 2001; Hacker and Dethier 2006). The changes caused by Spartina invasion can be viewed within the broader context of ubiquitous plant effects on wetland benthos. Plants have been recognized as ecosystem engineers (or foundation species) with major influence on the structure of marine communities, independent of invasion status (Levin and Talley 2002; Crooks 2002). The generalized

influences of introduced plants on the benthos can be partitioned into two major pathways. We identify the details of these pathways and the targets in figure 2.3 and illustrate these points using the changes produced by Spartina. The first pathway involves the indirect effects of plants on benthic organisms mediated through several changes in abiotic processes. First, reduced flow due to increased aboveground biomass generally reduces system energy; increases sediment stability, sedimentation rates, and organic matter accumulation; and reduces fluxes of reproductive propagules and food particles for consumers. Second, reduced light, also due to increased aboveground biomass, results in reduced evaporation and lower levels of sediment porewater salinity, sediment temperature, and higher water content. Third, plant belowground structures influence soil geochemical conditions (via root transport of oxygen, litter buildup and degradation), exploit space, and mechanically anchor sediments.

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The second pathway involves direct effects on benthic invertebrates mediated by the structural habitat they provide. Introduced plants can provide substrate for sessile epibionts, both plant and animal, simply by their presence in benthic habitats such as mudflats with limited emergent structure. Plants can also provide refugia for mobile predator and prey species, both infaunal and epifaunal, as well as provide living plant tissue for herbivores or litter for detritivores (Levin et al. 2006; Neira et al. 2006).

COMPARISONS WITHIN THE PACIFIC REGION To understand the overall effects of Spartina on benthic communities, we first compare our results with recent work on the benthic impacts of native S. foliosa. Studies comparing invertebrate communities in native versus invasive Spartina meadows in SFB showed that the native Spartina positively affects invertebrate abundance and, to a lesser extent, diversity in comparison to the largely negative effects of the hybrid (Brusati and Grosholz 2006). Invertebrate densities were up to five times greater in S. foliosa at some sites, although the magnitude of the effect was variable. This is largely the result of positive effects of aboveground structure but much lower belowground biomass. The density of roots and rhizomes apparently was not sufficient to reduce infaunal abundance as the hybrid does (see table 2.1). Recent removal and shading experiments conducted in southern California (Whitcraft and Levin 2007) suggest S. foliosa controls macrobenthic abundance and composition through light effects, which moderate soil temperature and salinity, enhance water retention, and promote microalgal growth. This positive effect of S. foliosa on benthic fauna parallels what we see in the initial stages of the hybrid Spartina invasion along the growing edge. Within the first few meters of a hybrid clone, plant densities and belowground biomass are similar to S. foliosa areas resulting in 34

the same facilitation of species abundance. This “edge effect” is similar to the increases in benthic abundance documented at the edges of seagrass meadows relative to the center (Bologna and Heck 2002). However, within five to ten years, the increasingly large above- and belowground biomass of the hybrid starts to reduce the benthic diversity and abundance through habitat changes at the surface and preemption of habitat space belowground. Our results from SFB and WB showing strong effects of Spartina invasion on benthic animal communities are broadly consistent with other studies of Spartina in the western United States. Other studies in WB have shown reduced infaunal diversity and abundance in the middle of Spartina meadows (Zipperer 1996; Dumbauld et al. 1994; O’Connell 2002) and negative effects on bivalve growth (Ratchford 1995). Interestingly, other WB studies have also documented an “edge effect” with increased diversity and abundance of some taxa in the growing edge of the Spartina marsh. Dumbauld et al. (1994) found increased abundances of introduced clams (Mya) within Spartina meadows in WB. Zipperer (1996), also working in the same area, found increased abundances of infaunal invertebrates in the newly growing Spartina border. The facilitative effects of S. alterniflora in WB are again similar to what we find for both S. foliosa and for the growing edge of the hybrid Spartina invasion in SFB. Our results showing dramatic shifts in the trophic mode of infaunal invertebrates following Spartina invasion are also in agreement with results from other studies in the U.S. West Coast region. In WB, Cordell et al. (1998) found a shift from surface-feeding bivalves and amphipods to deposit feeders and insect predators within the Spartina meadows. Also in WB, trophic shifts were documented by O’Connell (2002), who found increases in belowground detritivores in S. alterniflora marshes, consistent with our findings. Also, both Zipperer (1996) and Dumbauld et al. (1994) found reduced densities of bivalves and other surface

invasions in north american salt marshes

feeders within Spartina clones and increasing numbers of larval insects in areas with higher tidal elevation as the community moved toward a more terrestrial assemblage. Finally, many of the changes we have measured in the sediment ecosystem have also been documented following the invasion of Spartina anglica in Australia, England, and the Puget Sound region of Washington State. In Australia, Hedge and Kriwoken (2000) found that both the diversity and abundance of invertebrates were not significantly different between S. anglica–invaded areas and naturally vegetated marsh, but there was reduced diversity and abundance of invertebrates on nearby mudflats. In England, S. anglica has lowered densities of bivalves and Corophium and enhanced densities of tubificid oligochaetes relative to mudflat settings (Jackson 1985), with an overall effect of reduced species richness (Frid and James 1989). Studies of S. anglica invasion in Washington by Reeder and Hacker (2004) and Hacker and Dethier (2006) at sites with varying substrate and salinity found above- and belowground biomass varied substantially with habitat type, as did sediment accretion rates, sediment water content, salinity, and redox potential. In their mudflat and high-salinity marshes (most similar to our studies), sediment salinity was lower and water content was higher in the presence of invasive S. anglica. However, they found a predictable increased in redox potential in areas invaded by S. anglica, in contrast to our data showing lower redox in areas with hybrid Spartina. This difference is expected based on studies by Maricle and Lee (2001), who showed that S. anglica transports oxygen to their roots much more effectively than S. alterniflora.

COMPARISONS WITH OTHER REGIONS Studies from the native range of S. alterniflora have also shown that Spartina strongly influences the diversity and abundance of benthic species. Several studies from the Atlantic and Gulf coasts of the United States have generally

shown a positive effect of S. alterniflora on the abundance of invertebrates relative to unvegetated sediments (Rader 1984; LaSalle, Landin, and Simms 1991; West and Williams 1986). Similar results have been shown in studies of native Spartina alterniflora effects on infauna marshes in Brazil (Lana and Guiss 1991; Netto and Lana 1991). Shifts in trophic modes have also been measured in studies from the native range of S. alterniflora. Declines in the abundance of surface-feeding bivalves has been demonstrated in the native range of Spartina alterniflora in the eastern United States (Capehart and Hackney 1989), as have changes in trophic modes of infauna with successional stage of Spartina areas (Kneib 1984). Declines in suspension feeders in Spartina areas in comparison with open mudflats have also been documented in the native South American range of S. alterniflora (Lana and Guiss 1991). Finally, other studies have also shown impacts at higher trophic levels involving vertebrate consumers. In perhaps the bestdocumented study of impacts of invasive Spartina on shorebirds, Goss-Custard and Moser (1988) showed significant declines in the numbers of Dunlin (Calidris alpina) in estuaries in the United Kingdom that had been invaded by S. anglica. Overall, the conclusions from studies of Spartina in its native range suggest that it generally has a positive effect on the diversity and abundance of native benthos. This provides an interesting contrast with the largely negative effects of Spartina in the introduced range. One theory for opposing effects of Spartina on the two coasts is that predation pressure is considered much greater in Atlantic coastal systems (e.g., due to the presence of blue crabs, horseshoe crabs); thus, Spartina stands are more likely to function as refugia on the Atlantic coast. The much higher animals densities on Pacific than Atlantic mudflats supports this hypothesis (Levin, Talley, and Hewitt 1998). These findings are broadly consistent with our finding that Spartina tends to facilitate many

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species that have evolved within its native range and thrive in the microenvironment produced by this plant, but it negatively affects many species that have evolved in systems devoid of large vascular plants like Spartina and that generally do not tolerate the changed conditions.

PROSPECTS FOR ERADICATION AND RECOVERY Extensive Spartina eradication programs are currently being carried out in SFB and WB (fig. 2.4). In both bays, the primary means of eradication is aerial and boat-based spraying of the commonly used herbicide Imazypyr (known as “Arsenal”). The expectation is that complete eradication is a reality with sufficient effort invested over a long enough period of time. The goal in each case is restoration to the preinvasion condition. It is clear that some of the changes can be rapidly reversed. By removing the aboveground portion of the plant with herbicide, eradication efforts will be able to locally reverse the immediate physical modification of flow, light, and sedimentation. Our data suggest the aboveground contribution to added detrital buildup following eradication will be relatively short-lived. In areas that were previously vegetated by native plants, there is the expectation that native species will recolonize successfully, although additional efforts may be needed to jump-start this process in some

areas. The short-lived seed bank of the invasive Spartina provides hope that once eradication from large areas has occurred, it will be maintained notwithstanding dispersal from other areas. However, this will take vigilance in removing Spartina seedlings. Our work in areas that have undergone trial eradication or natural dieback suggests that there are more lasting effects of the roots and rhizomes that may require several years to decompose (Neira et al. 2007; A. C. Tyler and E. D. Grosholz, unpublished data). The rate of breakdown appears to be dependent on rates of flushing, sediment grain size, and other local habitat variables. The decay of belowground biomass and the sulfide and anoxia that result can slow the return of sediment conditions that would permit reestablishment of native plants and animals. Also, the increased tidal elevation created by enhanced accretion in invasive Spartina stands may not be quickly reversed. Because the increased elevation reduces tidal inundation, the mat of roots and rhizomes may continue to maintain the increased tidal elevation and substantially delay recovery. In some areas, if native plants colonize prior to the return of sediment to its preinvasion elevation, there may be a permanent shift from intertidal mudflat to vegetated tidal marsh, as has been documented following eradication of S. anglica in Puget Sound, Washington (Reeder and Hacker 2004; Hacker and Dethier 2006).

FIGURE 2.4 Ongoing eradication of Spartina alterniflora at a higher tidal elevation site in Willapa Bay, Washington.

36

invasions in north american salt marshes

However, there is room for optimism based on our monitoring of sites in WB that had been treated with herbicide seven years earlier. Our data indicate that recovery of sediment characteristics and porewater chemistry as well as invertebrate community indices can be achieved within six years and possibly faster (E. D. Grosholz and A. C. Tyler, unpublished data). However, these treated areas were small, isolated clones in relatively sandy sediments. In large continuous meadows or in muddier sediments, we predict that the recovery rate for native plants will be similar to that following S. anglica eradication (Reeder and Hacker 2004).

OVERALL CONCLUSIONS Introductions of nonnative species are clearly among the most pervasive anthropogenic changes in salt marsh systems and threaten to alter their structure and function on a broad scale. Our work has documented that Spartina invasions in two major estuaries in western North America are rapidly changing a wide range of physical, chemical, and biological processes. First, there are immediate changes to rates of light penetration, water flow, and sedimentation that then more gradually influence sediment physicochemistry as well as recruitment, growth, and survival of benthic organisms. The Spartina invasion has shifted the dominant primary producers from shortstature native plants and benthic microalgae to tall, dense plants that create approximately two kilograms per square meter of refractory aboveground detritus annually. This change results in a less “available” food source and slows down carbon and nutrient recycling. These changes in turn result in dramatic shifts in benthic communities and food web structure. In San Francisco Bay, our studies support the idea that this changing resource base appears to have contributed to a shift in benthic communities from surface feeders to subsurface feeders. The shifts in benthic communities may also have consequences for higher trophic levels (fig. 2.5). Within the Spartina meadow, the

FIGURE 2.5 Federally endangered California clapper rails (Rallus longirostris obsoletus) are impacted by the hybrid Spartina invasion in San Francisco Bay.

reduced densities of larger, surface-feeding invertebrates and increase in smaller infaunal detritivores may reduce the food base available to birds, fishes, and crabs. The presence of the invasive grass itself may also negatively affect wintering shorebirds by reducing the quality and availability of foraging areas on open tidal flats. Reversing the changes brought about by the Spartina invasion will require extensive management resources over a significant period of time. But our results provide reason for optimism regarding the success of recovery following eradication efforts. While complete eradication of Spartina in these systems may require several years of follow-up efforts, the elimination of most of the aboveground biomass and the decay of belowground biomass appear to begin within just a few years of the initial treatment. Our limited evidence suggests that these systems can return to something similar to the preinvasion state within approximately five years and possibly sooner. However, we have now documented that these changes are habitatspecific and that many factors such as tidal exchange, sediment grain size, elevation, and other variables will strongly affect the rate at which the system can recover (Reeder and Hacker 2004). Finally, our work provides examples of the many different kinds of effects that plants can have in estuarine systems. By altering hydrodynamics, sediment geochemistry, productivity,

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and sedimentation regimes—through food web modification and via substrate provision, stabilization, and refuge—invasive cordgrasses are certainly among the foremost agents of landscape-scale change in the coastal zone. Acknowledgments. We would like to thank R. Blake, U. Mahl, C. Whitcraft, N. Christiansen, N. Rayl, E. Brusati, S. Norton, C. Love, A. Carranza, G. Mendoza, C. Whitcraft, D. Chiang, M. Young, S. Maezumi, P. McMillan, and J. Gonzalez for assistance with laboratory and/or fieldwork in San Francisco and Willapa bays. Our thanks to D. Ayres and D. Strong for verifying Spartina hybrid genotypes in San Francisco Bay. We appreciate the timely and accurate stable isotope analyses provided by D. Harris and the University of California, Davis, stable isotope facility. We also thank C. Nordby, R. Blake, and E. Brusati for providing compliance with California clapper rail permit requirements in San Francisco Bay. Thanks also to B. Couch for providing air boat services in Willapa Bay. We greatly appreciate the comments of Brian Silliman and two anonymous reviewers, which helped improve the chapter. This work was provided by the National Science Foundation Biocomplexity Program (DEB 0083583) to E.D.G. and L.A.L. REFERENCES Ayres, D. R., D. L. Smith, K. Zaremba, S. Klohr, and D. R. Strong. 2004. Spread of exotic cordgrasses and hybrids (Spartina sp.) in the tidal marshes of San Francisco Bay. Biological Invasions 6: 221–231. Ayres, D. R., D. R. Strong, and P. Baye. 2003. Spartina foliosa (Poaceae): A common species on the road to rarity. Madroño 50: 209–213. Ayres, D. R., K. Zaremba, and D. R. Strong. 2004. Extinction of a common native species by hybridization with an invasive congener. Weed Technology 18: 1288–1291. Bologna, P. A. X., and K. L. Heck. 2002. Impact of habitat edges on density and secondary production of seagrass-associated fauna. Estuaries 25: 1033–1044. Bruno, J. F., J. J. Stachowicz, and M. D. Bertness. 2003. Inclusion of facilitation into ecological theory. Trends in Ecology and Evolution 18: 119–125. Brusati, E. D., and E. D. Grosholz. 2006. Native and introduced ecosystem engineers produce

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Dethier, M. N., and S. D. Hacker. 2005. Physical factors vs. biotic resistance in controlling the invasion of an estuarine marsh grass. Ecological Applications 15: 1273–1283. Dumbauld, B. R., M. Peoples, L. Holcomb, J. Tagart, and S. Ratchford. 1994. The potential influence of cordgrass Spartina alterniflora on clam resources in Willapa Bay, Washington. Journal of Shellfish Research 14: 228. Feist, B. E., and C. A. Simenstad. 2000. Expansion rates and recruitment frequency of exotic smooth cordgrass, Spartina alterniflora (Loisel), colonizing unvegetated littoral flats in Willapa Bay, Washington. Estuaries 23: 267–274. Frid, C., and R. James. 1989. The marine invertebrate fauna of a British coastal salt marsh. Holarctic Ecology 12: 9–15. Goss-Custard, J. D., and M. E. Moser. 1988. Rates of change in the numbers of dunlin, Calidris alpina, wintering in British estuaries in relation to the spread of Spartina anglica. Journal of Applied Ecology 25: 95–109. Hacker, S. D., and Dethier, M. N. 2006. Community modification by a grass invader has differing impacts for marine habitats. Oikos 113: 279–286. Hacker, S. D., D. Heimer, C. E. Hellquist, T. G. Reeder, B. Reeves, T. J. Riordan, and M. N. Dethier. 2001. A marine plant (Spartina anglica) invades widely varying habitats: Potential mechanisms of invasion and control. Biological Invasions 3: 211–217. Hedge, P., and L. K. Kriwoken. 2000. Evidence for effects of Spartina anglica invasion on benthic macrofauna in Little Swanport estuary, Tasmania. Austral Ecology 25: 150–159. Jackson, D. 1985. Invertebrate populations associated with Spartina anglica salt-marsh and adjacent intertidal mud flats. Estuary and Brackishwater Science Association Bulletin 40: 8–14. Kneib, R. T. 1984. Patterns of invertebrate distribution and abundance in the intertidal salt marsh: Causes and questions. Estuaries 7: 392–412. Kwak, T. J., and J. B. Zedler. 1997. Food web analysis of southern California coastal wetlands using multiple stable isotopes. Oecologia 110: 262–277. Lambrinos, J. L., D. R. Strong, J. C. Civille and J. Bando (in press). Implications of variable recruitment for the management of Spartina alterniflora in Willapa Bay, WA. In: Ayres, D. R., D. W. Kerr, S. D. Ericson and P. R. Olofson, eds. Proceedings of the Third International Conference on Invasive Spartina (San Francisco, CA), San Francisco Estuary Invasive Spartina Project of the State Coastal Conservancy (California), Cambridge Publications Limited, Cambridge, UK.

Lana, P. D., and C. Guiss. 1991. Influence of Spartina alterniflora on structure and temporal variability of macrobenthic associations in a tidal flat of Paranagua Bay (southeastern Brazil). Marine Ecology Progress Series 73: 231–244. LaSalle, M. W., M. C. Landin, and J. G. Simms. 1991. Evaluation of the flora and fauna of a Spartina alterniflora marsh established on dredged material in Winyah Bay, South Carolina. Wetlands 11: 191–208. Levin, L. A., C. Neira, and E. D. Grosholz. 2006. Invasive cordgrass modifies wetland trophic function. Ecology 87: 419–432. Levin, L. A. and T. S. Talley. 2002. Natural and manipulated sources of heterogeneity controlling early faunal development of a salt marsh. Ecological Applications 12: 1785–1802. Levin, L., T. Talley, and J. Hewitt. 1998. Macrobenthos of Spartina foliosa (Pacific cordgrass) saltmarshes in southern California: Community structure and comparison to a Pacific mudflat and a Spartina alterniflora (Atlantic smooth cordgrass) marsh. Estuaries 21: 129–144. Mack, R. N., D. Simberloff, W. M. Lonsdale, H. Evans, M. Clout, and F. A. Bazzaz. 2000. Biotic invasions: Causes, epidemiology, global consequences, and control. Ecological Applications 10: 689–710. Maricle, B. R., and R. W. Lee. 2002. Aerenchyma development and oxygen transport in the estuarine cordgrasses Spartina alterniflora and S. anglica. Aquatic Botany 74: 109–120. Mooney, H. A., and R. J. Hobbs. 2000. Invasive species in a changing world. Washington, DC: Island. Moseman, S., L. Levin, C. Currin, and C. Forder. 2004. Infaunal colonization, succession and nutrition in a newly restored wetland at Tijuana Estuary, California. Estuarine, Coastal and Shelf Science 60: 755–770. Neira, C., E. D. Grosholz and L. A. Levin (in press). Mechanistic processes driving shifts in benthic infaunal communities following Spartina hybrid tidal flat invasion. In: Ayres, D. R., D. W. Kerr, S. D. Ericson and P. R. Olofson, eds. Proceedings of the Third International Conference on Invasive Spartina (San Francisco, CA), San Francisco Estuary Invasive Spartina Project of the State Coastal Conservancy (California), Cambridge Publications Limited, Cambridge, UK. Neira, C., E. D. Grosholz, L. A. Levin, and R. Blake. 2006. Mechanisms generating modification of benthos following tidal flat invasion by a Spartina (alterniflora ⫻ foliosa) hybrid. Ecological Applications 16: 1391–1404. Neira, C., L. Levin, and E. D. Grosholz. 2005. Benthic macrofaunal communities of three sites in San

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Francisco Bay invaded by hybrid Spartina, with comparison to uninvaded habitats. Marine Ecology Progress Series 292: 111–126. Neira, C., L. A. Levin, E. D. Grosholz, and G. Mendoza. 2007. The influence of invasive Spartina growth phases on associated macrofaunal communities. Biological Invasions 9: 975–993. Netto, S. A., and P. C. Lana. 1999. The role of aboveand below-ground components of Spartina alterniflora (Loisel) and detritus biomass in structuring macrobenthic associations of Paranaguá Bay (SE Brazil). Hydrobiologia 400: 167–177. O’Connell, K. A. 2002. Effects of invasive Atlantic smooth-cordgrass (Spartina alterniflora) on infaunal macroinvertebrate communities in southern Willapa Bay, WA. Unpublished MS thesis, Western Washington University, Bellingham. Page, H. M. 1995. Variation in the natural abundance of 15N in the halophyte, Salicornia virginica, associated with groundwater subsidies of nitrogen in a southern California salt-marsh. Oecologia 104: 181–188. ———. 1997. Importance of vascular plant and algal production to macro-invertebrate consumers in a southern California salt marsh. Estuarine, Coastal and Shelf Science 45: 823–834. Page, G. W., L. E. Stenzel, and J. E. Kjelmyr. 1999. Overview of shorebird abundance and distribution in wetlands of the Pacific Coast of the contiguous United States. Condor 101:461–471. Rader, D. N. 1984. Salt-marsh benthic invertebrates: Small-scale patterns of distribution and abundance. Estuaries 7: 413–420. Ratchford, S. G. 1995. Changes in the density and size of newly-settled clams in Willapa Bay, WA, due to the invasion of smooth cordgrass, Spartina alterniflora Loisel. Unpublished MS thesis, University of Washington, Seattle. Reeder, T. G., and S. D. Hacker. 2004. Factors contributing to the removal of a marine grass invader (Spartina anglica) and subsequent potential for habitat restoration Estuaries 27: 244–252. Ruesink, J. L., B. E. Feist, C. J. Harvey, J. S. Hong, A. C. Trimble, and L. M. Wisehart. 2006. Changes in productivity associated with four introduced species: ecosystem transformation of a “pristine” estuary. Marine Ecology Progress Series 311: 203–215.

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Stenzel, L. E., C. M. Hickey, J. E. Kjelmyr, and G. W. Page. 2002. Abundance and distribution of shorebirds in the San Francisco Bay area. Western Birds 33: 69–98. Thompson, J. D. 1991. The biology of an invasive plant: What makes Spartina anglica so successful? BioScience 41: 393–401. Tyler, A. C., and E. D. Grosholz. Forthcoming. Spartina invasion changes intertidal ecosystem metabolism in San Francisco Bay. In Proceedings of the Third Annual Invasive Spartina Conference. Cambridge: Cambridge University Press. Tyler, A. C., J. G. Lambrinos, and E. D. Grosholz. 2007. Nitrogen inputs promote the spread of an invasive marsh grass. Ecological Applications 17: 1886–1898. Vitousek, P. M., H. A. Mooney, J. Lubchenco, and J. M. Melillo. 1997. Human domination of Earth’s ecosystems. Science 277: 494–499. West, D. L., and A. H. Williams. 1986. Predation by Callinectes sapidus (Rathbun) within Spartina alterniflora (Loisel) marshes. Journal of Experimental Marine Biology and Ecology 100: 75–95. Whitcraft, C. R., and L. A. Levin. 2007. Regulation of benthic algal and animal communities by salt marsh plants: Impact of shading. Ecology 88: 904–917. Windham L., and L. A. Meyerson. 2003. Effects of common reed (Phragmites australis) expansions on nitrogen dynamics of tidal marshes of the northeastern US. Estuaries 26: 452–464. Yamada, S. Y. 2001. Global invader: The European green crab. Unpublished manuscript, Oregon Sea Grant, Oregon State University. Zedler, J. B. 1980. Algal mat productivity: Comparisons in a salt marsh. Estuaries 3: 122–131. ———. 1984. The Ecology of Southern California Coastal Salt Marshes: A Community Profile. Washington, DC: U.S. Fish and Wildlife Service. Zedler J. B., and S. Kercher. 2004. Causes and consequences of invasive plants in wetlands: Opportunities, opportunists, and outcomes. Critical Reviews in Plant Sciences 23: 431–452. Zipperer, V. T. 1996. Ecological effects of the introduced cordgrass Spartina alterniflora on benthic community structure in Willapa Bay, Washington. Unpublished MS thesis, University of Washington, Seattle.

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Invasive Animals in Marshes biological agents of change James E. Byers Relative to other marine systems, salt marshes and estuaries are highly susceptible to invasion, and impacts by exotic species in these systems seem particularly pronounced. These impacts range from purely trophic and competitive effects that can lead to replacement of native species by exotics, to physical transformation by exotic species that engineer habitat and alter large-scale abiotic and hydrographic properties of the marsh environment. I discuss several examples of each of these, as well as three mechanisms that promote high establishment rates and strong competitive effects of nonnative species in marshes. Although the problem of exotic species in marsh systems is substantial, marshes’ tractable, discrete boundaries make intervention more successful than in other marine systems, such as the open coast. However, protocols and policies (ideally standardized at a national or international level) need to be in place for eradication or containment of incipient invasions, which often require fast action to be effective. Protocols and monitoring efforts should aim not only to detect newly introduced species but also to quantify the dynamics and impacts of established invaders to enable prioritization of intervention efforts. Estuaries and associated marshes are a heavily invaded habitat that must be well managed to mitigate the increasing ecological impacts of exotic species on native species and the valuable ecosystems services they provide.

Salt marshes are one of the most anthropogenically impacted marine ecosystems (Cairnes 1993; Kennish 2001; Nicholls 2004). Historically, physical and chemical impacts to salt marshes have been the most conspicuous and influential. Physical changes include extensive habitat conversion (e.g., filling and dredging) and altered hydrography that stem, especially in recent times, from marshes occupying valuable coastal real

estate. Kennish (2001) calculates that more than 50 percent of original tidal salt marsh in the United States has been lost through such physical alterations. Chemical impacts on marshes largely result from the close proximity of this marine habitat to humans and the associated pollutants they produce. That is, salt marshes are typically the outlets for watersheds where large volumes of anthropogenic pollutants are deposited (Fox et al. 41

1999; Sanger, Holland, and Scott 1999; Holland et al. 2004). Furthermore, as the large number of marsh Environmental Protection Agency (EPA) Superfund sites attests, marshes historically have directly received large quantities of contaminants because they were perceived as a convenient, valueless dumping ground. While these physical and chemical impacts have affected marsh biota for decades, only relatively recently, with the surge in globalization and consequent transport of nonnative species, have biological forces themselves been recognized as major agents of change (e.g., Ruiz et al. 1999). By skewing optimal environmental conditions away from conditions to which native species are adapted, many physical and chemical impacts may have set the stage for more frequent and successful biological invasions. Biological impacts of nonnative species range from purely trophic and competitive effects that can lead to replacement of native species by exotics, to physical transformation by exotic species that engineer habitat and alter largescale abiotic and hydrographic properties of the marsh environment. Perhaps the bestdocumented example of the severe biological impacts possible from invasive species is San Francisco Bay, where more than 240 nonindigenous species now reside, and 90 to 95 percent of the biomass is exotic in many areas of the bay (Cohen and Carlton 1998; Lee, Thimpson, and Lowe 2003). Clearly, to fully understand modern impacts on both the function and the taxonomic composition of estuaries and salt marshes, impacts of exotic species must be considered. Relative to other marine systems, salt marshes and estuaries are highly susceptible to invasion, and impacts by exotic species in these systems seem particularly pronounced. I begin this chapter by discussing underlying properties of estuaries that may be important in explaining these patterns. I then highlight some prominent impacts driven by invasive animals in salt marshes, using examples of nonindigenous species that exert large engineering effects and ones whose impacts are limited to trophic or competitive effects. I briefly conclude with 42

recommendations for how to best monitor and manage salt marshes against these threats. Although the focus of this chapter is animal invasions, most of the messages here are germane to invasions of all types.

ESTUARIES, INCLUDING THEIR ASSOCIATED MARSHES, ARE THE MOST INVADED MARINE HABITAT Compared to open coasts, a much higher number and proportion of exotic species are found in embayments, including associated marshes (see, e.g., Ruiz et al. 1997, 2000; Reise, Gollash, and Wolff 2002; Nehring 2002). For example, in Elkhorn Slough, California, Wasson, Fenn, and Pearse (2005) documented 526 invertebrates comprised of 443 natives, 58 exotics, and 25 cryptogens (species whose geographic origin is uncertain). The surrounding rocky intertidal open coast contained 588 species, of which only 8 were exotic and 13 cryptogenic. The number and proportion of exotics were significantly higher in the estuary (11 percent) than on the adjacent coast (1 percent). Furthermore, exotic species in the estuary were not only more diverse but also more abundant and conspicuous than on the open coast (e.g., the mudsnail Batillaria attramentaria, the orange sponge Hymeniacidon, the reef-building tube worm Ficopomatus enigmaticus). Similarly, while more than 240 nonnative species are known from San Franscico Bay, fewer than 10 are found on the adjacent outer coast (Ruiz et al. 1997). Perhaps even more enigmatic is the observation that in many cases, even species that are typically open coast residents in their native habitats often remain in embayments and marshes in their introduced range (Griffiths 2000; Robinson et al. 2005; Wasson et al. 2005). For example, the snail Littorina saxatilis, which is almost exclusively coastal in eastern North America and Europe where it is native, has not left the confines of San Francisco Bay where it has been established for more than a decade (Carlton and Cohen 1998; W. Miller, personal communication).

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Clearly this pattern of a higher number of invaders in marshes and estuaries may be due in part to a sampling bias. Embayments and marshes are located nearshore in close proximity to humans, who as a result have examined these habitats more than most other marine habitats (e.g., Ruiz et al. 1997, 2000; Cohen and Carlton 1998; Hewitt et al. 1999). In addition, compared to open marine systems, salt marshes have discrete boundaries and a high proportion of benthic, tractable species. However, three general, spatially variable factors likely contribute to a true pattern of higher numbers of introduced species in estuaries and marshes. First, bays and their associated marshes receive vast quantities of exotic propagules from ballast water. In U.S. ports alone, tens of millions of metric tons of ballast water from ports of origin all over the world are discharged yearly, with each liter containing up to ten zooplankton organisms (Verling et al. 2005). Second, bays are retention zones where larvae often are not advected away. Byers and Pringle (2006) demonstrated that advection typical of open coastlines makes retention, and thus establishment, difficult and may be largely responsible for the dearth of invasive species in those environs. Third, marsh and estuarine habitats best match the habitat from which most nonindigenous species propagules are exported. Two of the biggest vectors for nonindigenous marine species are ballast water and aquaculture, particularly shellfish imports (Ruiz et al. 2000). These sources most often originate from estuaries and bays, and the similarity of source and recipient habitats, especially in the case of intentionally introduced oysters and their associated communities (Ruesink et al. 2005), results in a high rate of successful establishment.

EXOTIC SPECIES MAY HAVE GREATER IMPACTS IN ESTUARIES AND MARSHES THAN OTHER MARINE HABITATS Having more nonindigenous species in estuaries and marshes increases not only their cumulative impact but also the odds of having species

with particularly high per capita impact. That is, having more exotic species essentially creates a sampling effect, whereby more species simply means species with greater impacts have a higher probability of being among those established. However, even after we standardize for the number of introduced species, we may still expect marshes to experience larger impacts from exotic species for at least two additional reasons. First, as emphasized previously (and throughout this book in general), salt marshes are one of the most anthropogenically disturbed habitats. Pollutants, eutrophication, habitat filling, drainage, dredge spoils dumping, and channelization are a few of the many, often severe abiotic alterations humans have imposed on this habitat type (e.g., Kennish 1992, 2001; Valiela, Rutecki, and Fox 2004; see also chaps. 8 and 9, this volume). The novel and sustained environmental changes that anthropogenic disturbances impose may often be enough to move a species out of the parameter space that defined its evolutionary history and to which it was adapted—a process dubbed selection regime modification (SRM) (Byers 2002a). A native marsh species may therefore suddenly find itself in an environment that in key ways is just as novel as it is to a nonindigenous species (Byers 2002a). SRM can thus accentuate competitive impacts of exotics on natives by eliminating a native species’ prior resident effect or “home court advantage.” That is, disturbances increase invader establishment and impact not only by creating new microhabitats, introducing propagules, and decreasing populations of native species that can resist invasion, but also by potentially weakening the per capita ability of the native biota to resist invaders. Because marshes are usually heavily altered by humans, they are a prime environment for selection regime modification and thus large resultant impacts through competition with exotic species (Byers 2000b, 2002a). Second, the retentive environment that contributes to increased exotic establishment in marshes also likely enhances their

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population-level competitive effects (Byers and Pringle 2006; Byers 2009). Within marshes, populations tend to be closed, and impacts of exotic species on natives may therefore be intensified. Specifically, because there is tight coupling of adults and successive generations, local, density-dependent impacts of exotic species may directly effect a lower population growth rate of the resident population. The separation between disparate marshes along a coast helps ensure their insulation and provides little chance that a declining native population will receive a rescue effect from an immigration pulse. Although the closed nature of salt marsh and estuarine habitats increases the propensity for high competitive impacts, it also has positive implications. This tendency toward closed populations should make invasive species there easier to control, mitigate, or eradicate. In fact, nearly all marine eradications are done in embayments. For example, Anderson (2005) describes the successful response of various government and nonprofit agencies responding to incipient invasions of Caulerpa taxifolia in a small southern California lagoon. Similarly, Hutchings, Hilliard, and Coles (2002) discuss the discovery of a problematic invasive mussel, Mytilopsis sallei, in two harbors near Darwin, Australia. The harbors were quarantined, and large doses of sodium hypochlorite and copper sulfate were added to poison the mussel. The process successfully eliminated Mytilopsis despite densities of the mussel that had already reached ten thousand square meters (Kuris 2003). In contrast, few if any eradication attempts on the open coast have been reported, except for one that was successful (Culver and Kuris 2000). An established population of the African shell-boring sabellid polychaete, Terebrasabella heterouncinata, was eliminated from a coastal site in California by removing 1.6 million potential snail hosts in the infected area (Culver and Kuris 2000). Although eradication was successful, it was only possible due to an extremely anomalous, localized distribution of the invader. 44

EXAMPLES OF PROBLEMATIC ESTUARINE/MARSH ANIMAL INVADERS Before we delve into some examples, it is worth stating that the depiction and assessment of impacts by nonindigenous species in marshes (as well as most every ecosystem) is likely conservative. Although there is a growing body of research on impacts of marine invaders (e.g., Grosholz 2002), many early invasions occurred with little notice. Presumably many of their immediate and pronounced impacts could have occurred long ago (Cohen and Carlton 1998), setting a different baseline for benchmarking modern-day changes (Dayton et al. 1998). A further issue that compounds the difficulty of assessing impact is the problem of cryptogenic species, species whose definitive native geographic distribution is unknown. For example, Ruiz et al. (2000) tallied 298 exotic invertebrate and algal species in coastal and estuarine waters of North America; however, this figure excludes cryptogenic species, hundreds of which may in fact be exotic species that just have not been identified as such. Even if impacts by such cryptogenic species are observed, the impacts cannot be definitively ascribed to an exotic species. Carlton (1996) estimates that 37 percent of the total number of known exotic and cryptogenic species in San Francisco Bay are cryptogenic; in Chesapeake Bay, this percentage is far higher. Robinson et al. (2005) report that South Africa has ten marine species that are confirmed as exotic and twenty-two as cryptogenic. As in the ecological literature, the taxa on which my examples are focused are better studied because they are larger and economically or culturally important. The invasion history of many fish, crustaceans, and mollusks, for example, is typically well known due to their importance as a human food source (and for mollusks also their interest to early shell collectors). Ray (2005b) calculates that in the Pacific Northwest and Alaska, 47 of the 162 exotic marine and estuarine animal species are mollusks and 39 are crustaceans, thus combining for over half of the total. Similarly, Ruiz et al. (2000)

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determine that half of the identified exotic invertebrate and algae species in coastal and estuarine waters North America are mollusks and crustaceans. Nonnative fish species are also a large contributor (Ruiz et al. 2000; Ray 2005a). For these species, it is not uncommon to have some background data on their invasion history and sometimes even their population dynamics. In contrast, many microorganisms, including diatoms, protozoans, and fungi, are easily transported unintentionally by humans, yet little historic record exists of these organisms because early natural historians seldom had the interest, let alone the expertise to identify them. Even today, many of these taxa are overlooked. Thus, the ease of incidental introduction and the likelihood of invasion detection are inversely correlated (Ruiz et al. 2000). Therefore, as a preface to the following examples, it is important to keep in mind that we know little about the impacts from what are likely to be the most common invaders. The examples that follow are organized by type of impact: invasive ecosystem engineers that change the structural character of the system and other invasive species that change competitive or trophic relationships between species. While there is a growing list of problematic, conspicuous invaders of North American estuaries and marshes (e.g., Chinese mitten crab Eriocheir sinensis, in San Francisco Bay [Rudnick et al. 2003]; Atlantic ribbed mussel Guekensia demissa, in California and Baja Mexico [Torchin, Hechinger, et al. 2005]), I have selected examples to illustrate some of the bestdocumented, quantified impacts. PROMINENT EXAMPLES OF EXOTIC ECOSYSTEM ENGINEERS

Organisms that create, modify, or destroy structure often disproportionately affect the communities they invade. These so-called ecosystem engineers (Jones, Lawton, and Shachak 1994) essentially alter the entire playing field on which ecological interactions take place by changing habitat structure, refuge availability, and even abiotic processes (e.g., hydrography)

(Wright and Jones 2006; Crain and Bertness 2006). Especially in urban areas, where marshes are often already small fractions of their original extent, any habitat conversion by ecosystem engineers can alter a substantial proportion of remaining marsh habitat. In terrestrial realms, ecosystem engineers are predominantly plant species, whereas in marine environments, it is largely animals that fill this role (e.g., oysters, coral, tube worms) with occasional contributions from vegetation (e.g., macroalgae) (Crooks 2002). As interfaces between terrestrial and marine environments, salt marshes have a mix of plant and animal engineers. Traditionally, the focus in marshes has been on the structural changes caused by introduced angiosperms (see chapters on Spartina and Phragmites in this volume), but plenty of invasive animal species are causing structural change as well. Examples of organisms that have been shown to be important habitat engineers include oysters, the tube worm Ficopomatus enigmaticus, the mussel Musculista senhousia, the burrowing isopod Sphaeroma quoyanum, and nutria (Myocastor coypus). These species engineer habitat in their native range as well; however, in a novel environment where their habitat structure arises de novo or where there are few checks on their abundance and thus the scale or rate at which habitat is altered, they can strongly affect native biota that do not share a common evolutionary history with the invasive engineer (Crooks 2002). FICOPOMATUS ENIGMATICUS

Ficopomatus enigmaticus is a cosmopolitan, reef-building serpulid polychaete introduced to many estuaries worldwide, including marshes of California and coastal lagoons of Argentina (fig. 3.1). In Mar Chiquita Lagoon, Argentina, reefs composed of thousands of calcareous tubes of this species cover roughly 80 percent of the lagoon. Reefs, which can be up to 4 meters in diameter and 0.5 meter high, increase habitat structure, modify the abundance of species that use it for shelter, and change the pattern of distribution of soft-bottom species (Schwindt and

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FIGURE 3.1 Ficopomatus enigmaticus reefs in Argentina. Photo courtesy of Martin Bruschetti.

Iribarne 2000). The reefs built by Ficopomatus in Mar Chiquita also greatly affect the physical environment by altering the bedload sediment transport and water flow (Schwindt, Iribarne, and Isla 2004). In general, ecosystem engineering species, such as F. enigmaticus, that invade soft sediment environments and create hard substrate tend to have large impacts on community composition. This is because hard substrata are novel to most marshes, and thus they often provide exotic species equal opportunity to compete for space because native marsh species are not specifically adapted to this habitat type. Thus, transformation of soft bottom estuarine habitat into hard substratum is a prime example of selection regime modification, and it has been found to be associated with increases in invasive species (Wasson et al. 2005; Tyrrell and Byers 2007). MUSCULISTA SENHOUSIA

Several decades ago, this Asian mytilid mussel invaded marshes and associated mudflats of Mission Bay (San Diego), San Francisco Bay, and Puget Sound among others along the Pacific coast of North America (Crooks 1998).

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The mussel anchors itself through byssal mats it produces that dramatically alter soft sediment habitats. Crooks (1998) noted marked changes to sedimentary properties and to the resident biota. Densities of macrofauna as well as species richness were typically higher inside than outside mussel mats. A tanaid amphipod, Leptochelia dubia, and the gastropod Barleeia subtenuis were particularly enhanced within mussel mats. In contrast, the abundance of native, filter-feeding clams declined, possibly because of competition for food (Crooks 2001). Through a series of experiments that compared community effects of live mussel mats to structural mimics, Crooks and Khim (1999), demonstrated that the physical structure of the mussels far outweighed the effects of living mussels. In fact, they determined specifically that the structure provided by the mussel mats was more influential than the mussel shells. In general, the effects of Musculista and Ficopomatus agree with observed effects of other habitat-modifying exotic ecosystem engineers, illustrating dramatic effects on biota by nonnative species capable of creating and altering physical structure.

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(A)

into mudflat (Talley, Crooks, and Levin 2001) (fig. 3.3). Carlton (1979) estimated that in infested areas of San Francisco Bay, vegetated marsh has been eroded back by dozens of meters since the introduction of Sphaeroma in the late 1800s. This makes Sphaeroma one of the largest agents of shoreline erosion in San Francisco Bay. By eroding vast tracks of marsh habitat, Sphaeroma not only accelerates conversion of marsh habitat into mudflat but also alters hydrography and sedimentation regimes (Talley et al. 2001). Bioeroders are a broad class of recognized ecosystem engineers (Meadows and Meadows 1991) that seemingly have potential for great impact in marshes where shallow root systems and persistent hydrological exposure can heighten their effects. MYOCASTOR COYPUS

(B) FIGURE 3.2 X-radiographs of marsh bank sediment containing (A) low and (B) high densities of Sphaeroma quoyanum. White areas are plant roots and rhizomes (A) or Sphaeroma burrows (B). Picture is from San Diego Bay, 1998. Photos courtesy of Theresa Talley and Springer.

SPHAEROMA QUOYANUM

Sphaeroma quoyanum is a burrowing Australasian isopod introduced to salt marshes of San Diego Bay and San Francisco Bay, among others. This isopod forms dense, anastomosing burrow networks (fig. 3.2). These burrows typically cut into the edges of marsh banks, reducing sediment stability and causing erosional loss in excess of one meter of marsh edge per year, dramatically reducing the extent of vegetated marsh and accelerating its conversion

Coypu, also known as nutria, were introduced from South America to almost every continent to farm for their fur (Carter and Leonard 2002) (fig. 3.4). In the southeastern United States, their population size is now estimated at twenty to thirty million (Maryland Department of Natural Resources 2004). Through their burrowing and rooting activity, nutria decrease aboveground biomass, belowground production, soil elevation, and the expansion of the root zone. Collectively these effects depress soilbuilding processes. Marshes with low sediment deposition are particularly susceptible to nutria impacts and may be destroyed without substantial human remediation (Ford and Grace 1998). Through destruction of some of the important vegetative physical structures of the marsh, coypu exert far-reaching consequences for the marsh community. EXAMPLES OF EXOTIC SPECIES WITH PURELY TROPHIC AND COMPETITIVE IMPACTS

In addition to structural changes, invasive estuarine and marsh species may also exert important impacts through their biotic interactions (e.g., competition, predation, parasitism, and apparent competition).

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(A)

FIGURE 3.3 (A) Extensive burrowing by Sphaeroma quoyanum in vertical marsh banks, San Diego Bay. (B) Such burrowing loosens sediment, increasing erosion, undercutting banks, and releasing chunks of marsh surface and reducing marsh habitat. Corte Madera Marsh, San Francisco Bay. Photos courtesy of Theresa Talley and Springer.

(B)

BATILLARIA ATTRAMENTARIA

An exotic snail, Batillaria attramentaria, has successfully invaded several salt marshes and mud flats in northern California, Washington, and British Columbia (Byers 1999) (fig. 3.5). In California, populations of the native mud snail, Cerithidea californica, have declined precipitously. Experimental manipulations demonstrated that the snail species competed exploitatively for epipelic diatoms colonizing the surface of the marsh mud. Although the two species did not differ in their effect on resource levels at any experimental snail density, the introduced snail was always more efficient at

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converting limited resources to tissue growth. Batillaria’s enhanced resource conversion efficiency provides a sufficient explanation for its successful invasion and subsequent exclusion of Cerithidea (Byers 2000a). Byers and Goldwasser (2001) subsequently combined these detailed, quantitative field data on Batillaria and its interactions with Cerithidea in an individual-based model. In empirically parameterized simulations, the native snail was driven extinct within fifty-five to seventy years after the introduction of Batillaria, closely matching direct field estimates of Cerithidea’s time to displacement (Byers 1999).

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FIGURE 3.4 Nutria (Myocastor coypus) in marsh. Photo courtesy of Guerry O’Holm.

pass through a second intermediate host and then on to a final host during their life cycles. Depending on the trematode species, mollusks, crustaceans, or fishes may serve as second intermediate hosts for the metacercarial cysts. The trematode life cycle is completed when a second intermediate host is eaten by the final host, typically a shorebird. Cerithidea californica hosts at least eighteen native trematode species throughout its range in California (Martin 1972). Batillaria, however, hosts just a single trematode species—itself a nonnative species (Cercaria batillariae) (Torchin, Byers, and Huspeni 2005; fig. 3.6). In Elkhorn Slough, California, hundreds of cysts of C. batillariae were found in all individuals of the three fish species examined, including within physiologically sensitive regions like the pericardium. Because trematode parasites are highly specific for the species of snail they infect, in marshes where C. californica becomes extirpated by Batillaria, Cerithidea’s parasites will also become locally extinct. The removal of

FIGURE 3.5 The exotic mudsnail, Batillaria attramentaria (the two snails pictured on the right), has successfully invaded several salt marshes and mud flats in northern California, where it contributes to the exclusion of the native confamilial mudsnail, Cerithidea californica (the two snails pictured on the left). Photo by Jeb Byers and John Meyer.

Replacement of one mud snail for another might seem like a superficial biotic change. However, because of its superior resource conversion efficiency, Batillaria achieves higher densities than Cerithidea, thus suppressing the diatom standing stock to very low levels. Such resource depression is likely to dramatically affect other benthic grazers within the marsh, such as ghost shrimp and other snails. But the largest impact stemming from the replacement of native Cerithidea was uncovered only recently. Both Batillaria and Cerithidea are first intermediate hosts for trematode parasites. All but one of these trematode species must obligately

FIGURE 3.6 Cercaria of the multihost trematode parasite, Cercaria batillariae. This nonnative trematode is found in marshes along the Pacific coast of North America where it infects only a single species of first intermediate host—the nonnative mudsnail, Batillaria attramentaria. The cercariae are released from infected snails and penetrate into native estuarine fish as a second intermediate host. Length of cercariae is about 0.5 millimeter. Photo courtesy of Todd Huspeni.

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multiple native trematode species will ultimately result in local eliminations of trematode infections in mollusks, crustaceans, and potentially several fishes that serve as second intermediate hosts for C. californica’s parasites (Martin 1972). Although the exact manifestations are still unclear, the reduction in infections will likely alter host population dynamics (Lafferty 1992) and potentially the metabolism, foraging, and abundance of shorebirds as well (Lafferty and Morris 1996). Thus, the most farreaching effects of Cerithidea’s loss on the community structure of the native marsh system seems very likely to stem from the concomitant loss of its dependent, trophically transmitted trematode species. VENERUPIS PHILIPPINARUM

Even in typically closed marsh populations, not every species exerts competitive effects (Byers 2009). Filter (suspension) feeding species in particular have been suggested to experience reduced competition because their planktonic food source is often not a limiting factor (Levinton 1972; Peterson 1979; Byers 2009). A nonindigenous suspension-feeding species, Venerupis philippinarum (⫽ Ruditapes philippinarum, ⫽ Tapes japonica), the Japanese littleneck clam, was accidentally introduced into the eastern Pacific in the 1930s with imported oyster seed from Japan and is now found in low-energy embayments and marshassociated mudflats from British Columbia to southern California (Quayle 1941; Haderlie and Abbott 1980). It is the most prolific introduced clam species in the San Juan Islands, Washington, and accounts for 50 percent of the annual commercial landings of hard-shell clams in the state (Washington Department of Fish and Wildlife 2000). Byers (2005) experimentally examined the effects of high Venerupis densities on mortality, growth, and fecundity of the confamilial clam, Protothaca staminea, and whether differences in predator abundance mitigate density dependent effects. Even at densities 50 percent higher than those found naturally in the field, Venerupis had 50

no direct effect on itself or Protothaca. Rather, the variable of overwhelming influence on the clams was crab biomass, which decreased growth of both species and increased mortality of Venerupis. The annualized loss rate of Venerupis was 50 percent when exposed to excavating crab predators due to a very shallow burial depth, nearly three centimeters shallower than Protothaca (Byers 2005). In fact, when exposed to predators, Venerupis was up to seven times more likely to be taken than Protothaca, whose mortality did not differ significantly with crab abundance. By taking the brunt of predator pressure, Venerupis seems to play a sacrificial role that at least partially protects Protothaca from predator mortality. However, high consumption of Venerupis likely has negative effects on other invertebrate species. When an exotic species is consumed by a native predator, losses to the exotic population are converted to additional predator biomass. Hence, predation on an exotic species can indirectly harm the predator’s native prey via apparent competition (Rand and Louda 2004). Noonburg and Byers (2005) demonstrate conditions under which resource competition from an invasive species is less detrimental to a native consumer than increased losses to a native predator population that is boosted by the invader. Thus, although no direct competitive effect of Venerupis on Protothaca was detected, by serving as an easy prey source for crabs, this nonindigenous prey species may be boosting regional crab abundance and productivity. Given the ubiquity of Venerupis throughout the West Coast of North America, this crab food subsidy could be substantial. In addition, the thin-shelled, nonindigenous Nuttallia obscurata, present in high abundance throughout the U.S. Northwest and British Columbia coast, is also an easy, novel prey item for crabs when it occurs in areas without appropriate physical refuges (Byers 2002b). Because Cancer crabs are omnivorous predators, their increase (particularly the less harvested species, C. gracilis and C. productus) potentially affects many other native prey

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species, including worms, fish, crustaceans, and other bivalves. Subsidies of native predators may be a largely underappreciated means by which nonindigenous species that are consumed heavily by native predators enhance apparent competition and thus escalate the impact of predators on native species (Courchamp, Langlais, and Sugihara 2000; Byers 2002b, 2005; Noonburg and Byers 2005). CARCINUS MAENAS

Due to inadvertent human transport, the European green crab, Carcinus maenas, is now a cosmopolitan species (Geller et al. 1997; Hidalgo. Baron, and Orensanz 2005). Its impacts have been well studied in marshes and embayments of the western coast of North America. Due to its omnivorous diet, Carcinus was found to significantly decrease the abundance of several invertebrate prey species in Bodega Bay, California (Grosholz et al. 2000). Most notably, two native clams, Nutricola tantilla and N. confusa, and a native shorecrab, Hemigrapsus oregonensis, decreased five- to tenfold within three years of the introduction of Carcinus. Carcinus indirectly promoted several polychaete and crustacean species, which apparently were not major prey items but rather benefited from Carcinus’s removal of competing species (Grosholz et al. 2000). In addition, Carcinus predation on Nutricola clam species facilitated spread of an introduced competitor clam, Gemma gemma throughout Bodega Harbor (Grosholz 2005). Carcinus thus has both direct and indirect predatory effects. Similarly, on the East Coast of North America, where it has been established for over one hundred years, Carcinus has been shown to have equally profound predatory impacts, particularly on clams. In Nova Scotia, Floyd and Williams (2004) measured 80 percent reductions of small Mya arenaria within less than four months, with consumption rates of three to twenty-two clams per crab per day. Whitlow, Rice, and Sweeney (2003) also measured high predation rates of Carcinus on M. arenaria. Although these authors demonstrated that Mya

was able to partially mitigate the high predation rate from the excavating crabs by burrowing more deeply in the sediment, deeper burial typically decreases feeding efficiency and thus growth rates of clams (Zaklan and Ydenberg 1997). Thus, predatory effects of Carcinus can be both density and trait mediated.

RECOMMENDATIONS AND CONCLUSIONS With increasing globalization, the influx of nonindigenous species to salt marshes is not likely to abate soon. The relatively retentive nature of most embayments, estuaries, and associated marshes seems to contribute both to higher invasive establishment as well as to strong competitive effects. Because nonindigenous species often originate from estuaries and bays, the similarity of source and recipient habitats also likely contributes to a high rate of successful establishment. Finally, the high rate of disturbance to marsh environments suggests that selection regimes may be sufficiently altered to promote high establishment rates and subsequent impact of introduced species. Policies to reduce the supply rate of exotic propagules, such as new mandatory ballast water exchange during transoceanic crossings (Federal Register 2004) are extremely commendable, since preventing invasions in the first place is the most proactive policy. But because this policy alone will not eliminate future invasions or mitigate effects of invaders already present, it is important to maintain vigilance in marsh habitats. In the United States, the Environmental Protection Agency’s Environmental Monitoring and Assessment Program and several National Oceanographic and Atmospheric Administration subagencies such as Sea Grant have conducted periodic regional inventories of marshes and bays for exotic species (Cohen et al. 2001; Lee et al. 2004; Pederson et al. 2005). Taking these efforts one step further to implement a standardized, national or international protocol to detect incipient invasions would strengthen our approach

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even more (e.g., Wasson et al. 2002). Ad hoc assessments have certainly led to some invaders being overlooked, presenting enormous problems in establishing appropriate baselines of impact assessment (Dayton et al. 1998) and slowing potential mitigation by resource managers. Protocols and policies need to be in place for eradication or containment of incipient invasions, which often require fast action to be effective (e.g., Anderson 2005). In addition to detecting introduced species and following the dynamics of established ones, quantitative monitoring is important because it aids restoration. Specifically, it complements adaptive management, which through frequent data-driven assessments can refine intervention techniques to restore native marsh populations, habitat, and ecosystem function. Clearly restoration efforts should be prioritized based on an understanding of invaders’ impact, be they engineering or trophic effects. Because impact assessment can be slow, if forthcoming at all, on first approximation such prioritization may do best to focus on invasive habitat engineers (e.g., oysters, reef-building polychaetes) because of their often far-reaching and lasting legacy effects and their ability to transform abiotic properties of a marsh and thus alter ecosystem services (Byers et al. 2006). Because not all invasive species are detrimental to native species or systems (e.g., Bruno et al. 2005; Byers 2009), monitoring marshes for early impact detection may be crucial for proper prioritization of intervention efforts. However, it is very important to avoid false confidence in the promptness of impact detection in such monitoring programs, because exotic impact is sometimes virtually undetectable until after the exotic species is extremely abundant (Byers and Goldwasser 2001). This lag in impact may be especially pronounced if the mechanism of impact by the exotic species is to decrease births of a long-lived native species—a less conspicuous impact than increasing death of adults. In such cases, demographic lag times necessitate alternative, faster-responding metrics for impact detection than adult population densities. 52

As an illustration, Byers and Goldwasser (2001) sought to identify empirically measurable quantities that provide the earliest warning of impact by the invasive snail Batillaria on the native Cerithidea. Through an individual-based model parameterized with empirical data, they tracked many population- and individual-level responses of Cerithidea to Batillaria’s invasion, including population density, biomass, egg production, mean size, proportion of parasitized individuals, and individual growth rate, as well as availability of shared food resources. In model simulations, the initial number of invading Batillaria was set to guarantee extinction of Cerithidea within ninety years. Despite a rapid initial increase in the invader population, all metrics for Cerithidea were slow to exhibit signs of impact. Most took at least twenty-five years from invasion to exhibit detectable changes, by which time the exotic snail was established at extremely high densities. Cerithidea egg production was the fastest, most consistent response metric, exhibiting declines within twenty to twenty-five years after invasion in about 90 percent of simulations. Monitoring programs and risk assessment analyses must identify and concentrate on reliable, early-warning metrics. Habitat alteration and biotic effects by exotic species now join anthropogenic physical alterations and chemical inputs as dominant impacts on estuarine and marsh environments. Dramatic modification of selection regimes in marshes suggests that synergism between these disturbance agents enhances invasion rates and impacts. Therefore, policies to protect marshes and estuaries from physical and chemical anthropogenic disturbance should simultaneously improve their resistance and resilience to biological invasions. Although the closed nature of most estuarine and marsh habitats and their populations increases the propensity for impacts by nonindigenous species, it also makes the species easier to document, track, control, and potentially eradicate. For example, the tractable, discrete boundaries of marsh systems certainly make intervention more successful than in other marine systems such as the

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open coast with its porous boundaries, complicating currents, and an influx of propagules from elsewhere. Marshes are a jeopardized, increasingly compromised habitat to which we must pay strong attention; otherwise, we risk losing many of the native species and valuable ecosystems services they provide. Acknowledgments. I thank Irit Altman, Tyra Byers, John Meyer, Whitman Miller, and Ted Grosholz for helpful edits and discussions on this chapter. REFERENCES Anderson, L. W. J. 2005. California’s reaction to Caulerpa taxifolia: A model for invasive species rapid response. Biological Invasions 7: 1003–1016. Bruno, J. E., J. D. Fridley, K. D. Bromberg, and M. D. Bertness. 2005. Insights into biotic interactions from studies of species invasions. Pages 13–40 in D. F. Sax, J. J. Stachowicz, and S. D. Gaines (eds.), Species Invasions: Insights into Ecology, Evolution, and Biogeography. Sunderland, MA: Sinauer. Byers, J. E. 1999. The distribution of an introduced mollusc and its role in the long-term demise of a native confamilial species. Biological Invasions 1: 339–353. ———. 2000a. Competition between two estuarine snails: Implications for invasions of exotic species. Ecology 81: 1225–1239. ———. 2000b. Differential susceptibility to hypoxia aids estuarine invasion. Marine Ecology Progress Series 203: 123–132. ———. 2002a. Impact of nonindigenous species on natives enhanced by anthropogenic alteration of selection regimes. Oikos 97: 449–458. ———. 2002b. Physical habitat attribute mediates biotic resistance to nonindigenous species invasion. Oecologia 130: 146–156. ———. 2005. Marine reserves enhance abundance but not competitive impacts of a harvested nonindigenous species. Ecology 86: 487–500. ———. 2009. Competition in Marine Invasions. In: Biological invasions in marine ecosystems: Ecological, management, and geographic perspectives. Eds.: Gil Rilov & Jeff Crooks. SpringerVerlag. pp. 245–260. Byers, J. E., K. Cuddington, C. G. Jones, T. S. Talley, A. Hastings, J. Lambrinos, J. A. Crooks, and W. G. Wilson. 2006. Using ecosystem engineers to restore ecological systems. Trends in Ecology and Evolution 21, no. 9: 493–500.

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Phragmites australis in Eastern North America: A Historical and Ecological Perspective Laura A. Meyerson, Kristin Saltonstall, and Randolph M. Chambers Phragmites australis is arguably one of the most successful plant invaders in coastal marsh systems of the United States. An aggressive Phragmites lineage—likely introduced to the northeastern United States in the nineteenth century—presently is sweeping through coastal wetlands of Atlantic states and can also be found in parts of the Gulf and Pacific coasts as well as inland habitats. While native Phragmites was previously found in many of these habitats, observation and experimental research have made it clear that the spread of introduced Phragmites is closely coupled with anthropogenic disturbance of the physical and chemical environment. The primary factors favoring establishment and spread of the introduced type are physical disturbance of wetland margins that opens up new sites for initial colonization and eutrophication, which relaxes interspecific competition for nutrients belowground and promotes the rhizomatous expansion of tall, dense monocultures. Overall, the ecological impacts of Phragmites invasion and expansion are viewed negatively—that is, reduction/loss of native plant coverage and diversity, alteration of wetland hydrology, changes in ecosystem functions, and perceived failures of wetland creation and restoration projects owing to Phragmites invasion. However, Phragmites wetlands must be considered effective sinks for nutrients, and their rapid growth and organic sediment accumulation rates may be sufficient to match projected rises in sea level, facilitating landward migration of salt marsh systems. In this chapter, we review recent progress in the science and management of Phragmites including comparisons between the native and introduced lineages, characteristics that enhance the invasiveness of Phragmites, the consequences and benefits of Phragmites invasions, and the role of Phragmites in marsh restoration.

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Biological invasions pose one of the greatest threats to native species and the ecosystems they inhabit (Mooney and Hobbs 2000; Sala et al. 2000). Invasive species are organisms introduced from outside their historical range to novel ecosystems and cause harm to the environment, the economy, or human health (National Invasive Species Council [NISC] 2001). Invasive species can alter the structure and function of environmental systems at all levels of organization, from genes to ecosystems, and are thought to have impacted at least half of the species listed as threatened or endangered in the United States (Wilcove et al. 1999). Virtually every state and ecosystem in the United States has been affected by invasive species. Estimates of the number of nonnative species (plants, animals, insects, microorganisms) in the United States vary by orders of magnitude, but the numbers of invasive introductions have clearly increased over time. Invasive plants alone are estimated to infest more than a hundred million acres of U.S. land at a cost of more than $13 billion per year (NISC 2001). Biological invasions impose significant economic costs and management challenges for every major ecosystem type in the United States. Of these, wetland ecosystems have been affected perhaps most dramatically. Disturbances such as alteration of hydrologic cycles and nutrient regimes, habitat degradation, and loss via land conversion have impaired wetland structure and function but, more significantly, have facilitated invasive species establishment. Anthropogenic factors often account for the arrival and establishment of plant invaders in novel terrestrial and aquatic environments; wetlands may be particularly vulnerable to plant invasions due to their downstream position relative to other ecosystem types. This positioning makes wetlands “landscape sinks” for nutrients, debris, and other materials that facilitate plant invasions (Zedler and Kercher 2004). Phragmites australis (Cav.) Trin. ex. Steud. (common reed) is one of the most prominent recent invaders of coastal marsh systems in the United States, particularly along the Atlantic 58

coast. Phragmites is a robust, perennial emergent grass found on every continent with the exception of Antarctica (Tucker 1990). It has a wide range of tolerance for environmental conditions and can grow in fresh, brackish, and salt marsh systems (Hocking, Finalyson, and Chick 1983; Marks, Lapin, and Randall 1994). In North America, introduced Phragmites establishes new stands by both seed and dispersal of rhizome fragments, while expansion of existing stands is primarily clonal (Meyerson, Saltonstall, et al. 2000; Bart and Hartman 2002, 2003; Saltonstall 2003b; Farnsworth and Meyerson 2003). Phragmites can produce large quantities of seeds, but germination rates are variable and generally thought to be low (Harris and Marshall 1960; Galinato and van der Valk 1986; Tucker 1990; Marks et al. 1994), probably due to self-incompatibility, the tendency of the species to grow in large genetically identical homogeneous clones (Tucker 1990), and abiotic factors such as hydroperiod, salinity, and temperature that regulate seed germination rates (Galinato and van der Valk 1986). Most Phragmites invasions into tidal wetlands are presumed to be initiated by the deposition and establishment of rhizome fragments (Bart et al. 2006); once established, expansion occurs primarily through clonal growth. Growth of Phragmites populations can be rapid—a single clone can cover an eighth of a hectare in two years (Hocking et al. 1983)—and the slow decomposition of Phragmites detritus can significantly reduce the availability of nutrients, light, and space, making the survival or establishment of other species unlikely (Meyerson 2000). Phragmites is a very successful colonizer of North American coastal marshes and has produced a suite of ecological changes—some of which are considered beneficial and others that are not. Here we review the documented biotic and abiotic alterations that have resulted from Phragmites colonization; progress in the science and management of Phragmites, including comparisons between the native and introduced lineages; characteristics that enhance

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the invasiveness of Phragmites; consequences and benefits of Phragmites invasions; and the potential role of Phragmites in marsh restoration. Finally, we identify knowledge gaps and research needs for Phragmites.

PHRAGMITES IN NORTH AMERICA HISTORICAL DISTRIBUTION

Phragmites has been present in North America for thousands of years and therefore should be considered a native plant. The oldest evidence comes from the southwestern United States where Phragmites remnants have been found in preserved Shasta ground sloth dung dating back forty thousand years. From dietary analysis, Phragmites composed up to 63 percent of the sloth’s diet at some sites, indicating that during the Pleistocene, Phragmites might have been quite common in this part of the United States (Hansen 1978). Findings from archeological sites throughout the Southwest dating from 600 to 1400 AD have also found a number of artifacts made of Phragmites, indicating that it was widespread during this time period and used by indigenous peoples in numerous ways (Adams 1990; Kiviat and Hamilton 2001). Peat core profiles provide an accurate assessment of the plant communities present in a given horizon, particularly in salt marsh communities. In coastal areas, Phragmites rhizomes preserved in peat have been found in a number of sites (California: Goman and Wells 2000; Delaware: Kraft 1971; New Jersey: Waksman et al. 1943; Connecticut: Niering, Warren, and Weymouth 1977; Orson 1999; New York: Clark 1986). Based on these studies, it appears that historical populations of Phragmites were often limited to the high marsh and typically grew in mixed communities with sedges and forbs and not as monocultures that are more common today (Waksman et al. 1943; Orson 1999). Clearly, Phragmites has been present in both coastal and inland marshes for thousands of years and certainly was present before the arrival of Europeans to North America.

Herbarium records and floras from the nineteenth and early twentieth centuries also indicate that Phragmites was found across the continent, although it was likely not present in some southeastern states, such as South Carolina, Georgia, Arkansas, and Kentucky (Saltonstall 2002). In coastal marshes of New England and the mid-Atlantic region, Phragmites was considered rare or not common during the nineteenth century (Torrey 1843; Willis 1874; Dame and Collins 1888), but by the early twentieth century, its distribution and abundance were beginning to increase (Graves et al. 1910). Throughout the twentieth century, the distribution of the species continued to expand and by the mid-1970s, and Phragmites was reported in all of the lower forty-eight United States and across southern Canada (Chambers, Meyerson, and Saltonstall 1999; Saltonstall 2002). GENETIC DIVERSITY OF PHRAGMITES IN NORTH AMERICA

As early as the 1870s, a nonnative form of Phragmites was thought to have been introduced and established in North America (PH-22894; Burk 1880). In the 1980s and 1990s, a number of researchers also speculated an introduction had occurred based on the relatively recent and aggressive spread of the species (Metzler and Rosza 1987; Tucker 1990; Mikkola and Lafontaine 1994) and morphological evidence that suggested differences between historical and modern populations (Besitka 1996). Genetic evidence confirms that both native and introduced Phragmites lineages are found today in North America (Saltonstall 2002; Saltonstall 2003a, 2003b; fig. 4.1a, 4.1b). Using chloroplast DNA (cpDNA), Saltonstall (2002) showed that two different forms of Phragmites occur across much of the continent. One type, hereafter referred to as introduced Phragmites, belongs to a single lineage, or cpDNA haplotype M. The lineage is common throughout Europe and Asia and closely related to other haplotypes found there and clearly not native to North America. In contrast, thirteen native North American Phragmites cpDNA haplotypes have

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(A)

(B)

(C)

FIGURE 4.1 Distribution of (A) native (subsp. americanus), (B) introduced, and (C) Gulf Coast (ssp. berlandieri) Phragmites australis in North America. Although not documented across the Gulf Coast except for in the Mississippi Delta (Saltonstall 2002), introduced Phragmites may already have invaded these regions and certainly has the potential to spread into them. The distribution of introduced Phragmites is not known south of the U.S. border and thus is not included in this figure. Reprinted with permission from Saltonstall et al. 2004.

been identified (hereafter referred to in a group as native Phragmites; Saltonstall 2002; Meadows and Saltonstall 2007), all of which share several mutations not found in Phragmites populations elsewhere in the world. These thirteen native haplotypes are distributed around the continent, except along the Gulf coast, and genetic structuring can be seen among Atlantic Coast, midwestern, and southwestern populations (Saltonstall 2003a). These results are confirmed by nuclear microsatellite analyses that show different patterns of allele sizes distinguishing native and introduced Phragmites as well as regional structuring in native populations from different parts of the continent (Saltonstall 2003b). Morphological differences in ligule height, glume lengths, guard cell size, and stomatal density also support the genetic distinction 60

between native and introduced Phragmites lineages (Saltonstall, Peterson, and Soreng 2004; Saltonstall et al., 2007). Stem color, stem texture, adherence of leaves to the stem, and clonal growth patterns may also be used to distinguish native and introduced Phragmites, although these characters are more subject to observer bias and are not as reliable when identifying the different lineages (Blossey 2003; Saltonstall et al. 2004). Based on these genetic and morphological differences, native North American Phragmites has been named a distinct subspecies, P. australis subsp. americanus (Saltonstall et al. 2004). A third subspecies, commonly called the Gulf Coast lineage (Pellegrin and Hauber 1999; cpDNA haplotype I, P.a. subsp. berlandieri), is found across the southernmost states from Florida to the Gulf of California and south into

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Mexico (Saltonstall and Hauber 2007). Although genetically distinct from native and introduced Phragmites, at this time a native/introduced status cannot be assigned to this subspecies as it is also found in South America and Asia. How long the Gulf Coast lineage has been present in North America is unknown (Saltonstall 2002, 2003a, 2003b). Gulf Coast Phragmites is not considered to be an aggressive invader, although it is common throughout much of its range and can be the dominant species in the marshes it occupies. Introduced Phragmites has also been identified in the Mississippi River Delta where its range overlaps with the Gulf Coast lineage (Saltonstall 2002; White, Hauber, and Hood 2004).

CURRENT DISTRIBUTIONS OF PHRAGMITES Phragmites australis is one of the most widely disseminated plant species worldwide (Tucker 1990), with a geographic distribution extending from the tropics to cold temperate regions in both hemispheres. Phragmites is common throughout Europe—some of the largest stands occur on the Danube River delta—and reed beds are valuable ecosystems often protected for their important ecological functions. Over 150 species of insects, herbivores in particular, have been reported in European reed beds, and these form the base of a multilayered food web (Tscharntke 1999; Tewksbury et al. 2002). Phragmites stands provide habitat for wild fowl and fauna (Brix 1999) and are used by many bird species during their annual migrations between Europe and Africa. A group of European warblers uses Phragmites exclusively for breeding habitat (Berthold et al. 1993). Phragmites is also used by humans for a variety of purposes, including cellulose production, matting, thatching of roofs (Haslam 1972), land reclamation and erosion prevention (Bakker 1960; Coops, Geilen, and van der Velde 1994), and wastewater treatment (Cooper and Green 1995). Recent declines of Phragmites in parts of Europe have prompted much research (i.e., the

EUREED project—for a summary, see van der Putten 1997). Factors that have been identified as key contributors to these declines include increased disturbance, habitat destruction, manipulation of hydrologic regimes, eutrophication, and pollution (Ostendorp 1989; van der Putten 1997). Ironically, many of these factors are also considered causal in the invasive growth of Phragmites seen in North America in recent decades. In North America, Phragmites typically has been considered a signature of tidal wetland alteration (Niering and Warren 1980), but over the past fifty years, introduced Phragmites has established and invaded wetlands even where apparent anthropogenic impacts are minor or absent. Furthermore, rates and extent of spread into different wetland types and geographic regions have been variable. For example, regularly flooded tidal wetlands with salinity above eighteen parts per thousand are for the most part protected from Phragmites expansion. These polyhaline wetlands (Cowardin et al. 1979) typically experience Phragmites establishment at the upland margin (Marks et al. 1994), but growth out from the margins does not occur (but see Amsberry et al. 2000; Burdick and Konisky 2003). In oligohaline and mesohaline portions of estuaries, however, Phragmites expansion is a real threat. Phragmites has the potential to establish and spread throughout tidal wetlands occurring in a broad salinity range from 0.5 to 18 parts per thousand. Introduced Phragmites is one of the few salt-tolerant invasive wetland plants (Chambers et al. 2003), and its distribution in North American estuaries—especially in the Northeast—highlights this point. The vast majority of Phragmites populations found along the Atlantic coast are introduced, particularly in salt and brackish marsh systems and inland habitats. In New England and in the mid-Atlantic, introduced Phragmites is expanding in mesohaline marshes, particularly in those wetlands that have been diked and/or drained (R. Rosza, personal communication). An invasive wave of Phragmites has crested throughout much of the Northeast and appears

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to be pushing south into the mid-Atlantic states (Saltonstall 2002, 2003a). For example, Phragmites colonization is extensive in the tidal wetlands of New Jersey and Delaware, and the front appears to be migrating south through Maryland and Virginia. Using the National Wetland Inventory’s (NWI) GIS database from 2002, we estimated a total of 480 square kilometers of Phragmites in tidal wetlands of lower Delaware Bay and the eastern shore of Chesapeake Bay. By comparison, Phragmites covers only sixty square kilometers in tidal wetlands on the western shore of Chesapeake Bay (R. M. Chambers, unpublished data). Introduced Phragmites has moved rapidly down the Delmarva Peninsula but is moving more slowly across the major tidal rivers down the western shore of the Chesapeake Bay. Based on a visual survey of almost 8,400 kilometers of shoreline, Chambers et al. (2008) show that roughly 14.6 percent of the more northern Maryland shoreline is occupied by Phragmites. Farther south of this current front, Phragmites occupies only 2 percent of the Virginia shoreline of Chesapeake Bay. Wetland managers in North Carolina and South Carolina report that introduced Phragmites is becoming established primarily in

wetlands managed for waterfowl perhaps due to physical disturbance and impoundments, water drawdowns that allow germination of Phragmites seeds, and nutrient additions from waterfowl guano (J. Stanton, U.S. Fish and Wildlife Service, NC, personal communication). These observations are consistent with human facilitation of Phragmites expansion. Following the pattern of spread seen farther north, managed wetlands may act as a source of rhizomes and seeds to facilitate the introduction of invasive Phragmites into other tidal wetlands in the region. Native populations that historically were more abundant now are rare along the Atlantic coast, except in the mid-Atlantic region where native Phragmites can be found in low-salinity tidal wetlands (Packett and Chambers 2006; Meadows and Saltonstall 2007). These areas are perhaps under the greatest threat from the continued expansion of introduced Phragmites. Native Phragmites typically is smaller in stature, grows in mixed plant communities, and has a lower culm density than introduced Phragmites, although populations with high culm densities can occur (Meadows 2006) (fig. 4.2). Common associates of the native in the mid-Atlantic include big cordgrass (Spartina cynosuroides),

FIGURE 4.2 Native marsh community including P. australis subsp. americanus along Wicomico Creek, Allen, Maryland. Photo by K. Saltonstall.

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smooth cordgrass (S. alterniflora), narrowleaved cattail (Typha angustifolia), rice cutgrass (Leersia oryzoides), wild rice (Zizania aquatica), Halberd-leaved tearthumb (Polygonum arifolium), river bulrush (Scirpus fluviatilis), and marsh mallow (Althaea officinalis; R. E. Meadows and K. Saltonstall, unpublished data). In the low-salinity wetlands of some river systems on the eastern shore of Maryland, more acreage is covered by native than introduced Phragmites. However, introduced Phragmites is also found throughout these watersheds and may be in relatively early stages of colonization (Meadows and Saltonstall 2007). Assessments of the relative amounts of native and introduced Phragmites are underway and provide a preliminary indication of distribution patterns. For example, Packett and Chambers (2006) provide data on the number, size, and location of native and introduced Phragmites stands in the Rappahannock River. Meadows and Saltonstall (2007) have done similar work on the eastern shore of Maryland and on the lower Delaware Bay. Similarly, Lambert and Casagrande (2007) have documented native Phragmites populations in Rhode Island, Payne and Blossey (2007) have done so in Massachusetts, and additional assessments are underway in other locations as well. Despite the lack of publications on the relative patterns of native and introduced Phragmites, researchers have consistently described native stands as smaller and fewer. In the mid-Atlantic region, native stands are distributed mostly in oligohaline to tidal fresh marshes, while in New England they are found in mesohaline marshes as well. Introduced Phragmites stands, however, are larger, more numerous, and distributed for the most part in oligohaline to mesohaline marshes. Both native and introduced Phragmites populations are readily found throughout much of the northern half and southwestern parts of the continental United States and much of Canada (S. De Blois, personal communication; Saltonstall 2002). Native populations are typically found in undisturbed, natural areas, and

evidence for recent expansion of some native populations exists (Lynch and Saltonstall 2002). Around the Great Lakes, introduced Phragmites is common, particularly along roadsides and lakeshores. These habitats may provide corridors for dispersal of nonnative populations and appear to be facilitating spread in the Midwest. Introduced populations are rare in western North America but are more typically found in urban areas. Few roadside populations are seen (Saltonstall 2003a). Along the West Coast of the United States, the number and area of tidal wetlands are very small relative to the Gulf and Atlantic coasts of North America. To date, no native Phragmites populations have been identified in any tidal wetlands of Washington State, but introduced Phragmites has been found in small patches in Grays Harbor and Puget Sound (G. Haubrich, personal communication; Saltonstall 2002). In California, six stands of introduced Phragmites have been identified in the Humboldt Bay area (T. Gedik, personal communication, Saltonstall 2002) and in San Diego County, but introduced Phragmites is probably most extensive in the San Francisco Bay estuary, especially Suisun Marsh and the Sacramento–San Joaquin delta (J. G. Mensik, personal communication). In Florida and the Gulf Coast states, Gulf Coast Phragmites is far more common than the introduced lineage (Saltonstall 2002). In most tidal wetlands of this region, Gulf Coast Phragmites populations are stable. However, they appear to be expanding in some places, such as the marshes surrounding Mobile, Alabama (W. Finch, personal communication). In the Mississippi Delta, introduced Phragmites established in the first half of the twentieth century and now is widespread, estimated to cover up to 50 percent of the colonizable areas in the Delta (D. Hauber, personal communication; Saltonstall 2002). These nonnative populations could have established from propagules washed downriver from upstream (Saltonstall 2003a), or they could represent a novel introduction that has occurred along the Gulf coast

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(Saltonstall 2003b). Although production of viable seed is thought to be low in Phragmites along the Gulf coast (Fournier, Hauber, and White 1995), the spread of introduced populations to other areas of the Gulf Coast is clearly possible via dispersal of seeds or rhizome fragments (Saltonstall 2003a). Despite the overlap in the range of the two Phragmites subspecies, Saltonstall (2003b) detected no evidence for interbreeding between the native and introduced lineages. This is curious, given that they are considered to be the same species. However, microsatellite techniques that can detect interbreeding within this species have only recently been developed (Saltonstall 2003b) and have not yet been widely utilized. Hybrid seedlings of the north American native P. a. subsp. americanus and introduced P. australis native have recently been produced by L. A. Meyerson (unpublished data) in a controlled pollination, common garden greenhouse study in Rhode Island in the United States. Multiple native and introduced populations of Phragmites that grow together in the wild are documented to have overlapping flowering periods both in the field and in the common garden setting (L. A. Meyerson, unpublished data). Further studies are currently underway by Meyerson to conduct additional crosses between native and introduced Phragmites populations in the common garden and in the field, to backcross hybrid offspring to parent populations, and to compare fitness and vigor between the parent populations and hybrid offspring. Outbreeding depression that results in reduced fitness of the F1 generation may explain why intraspecific hybrids of native and introduced Phragmites have not been found in the wild (Hufford and Mazer 2003), but these individuals may also have simply been overlooked. Surprisingly, the reproductive biology of Phragmites is not fully understood even after decades of study and hundreds of publications. While it is generally thought that selfcompatibility of Phragmites is low (e.g., Gustaffson and Simak 1963; McKee and Richards 1996; Ishii and Kadono 2002) and 64

therefore seed set is low (Gervais et al. 1993; Pellegrin and Hauber 1999), work in Europe and Japan has clearly demonstrated the ability of Phragmites to spread by means of sexual reproduction (Koppitz et al. 1997; Alvarez, Taon, and Mauchamp 2005). In North America, K. Saltonstall (unpublished data) observed that introduced Phragmites populations along several rivers in Connecticut are not genetically identical and therefore are likely spreading via seed dispersal rather than by rhizome fragments. Furthermore, selfincompatibility in Phragmites australis has not been conclusively proven since recent work in Japan did not emasculate Phragmites inflorescences and therefore could not determine whether seed set was due to selfing or asexual seed formation (Ishii and Kadono 2002).

WHAT MAKES INTRODUCED PHRAGMITES SO INVASIVE? While a number of competing hypotheses have been proposed regarding the rapid spread of Phragmites in North America (Meyerson, Vogt, and Chambers 2000; Windham 2001; Silliman and Bertness 2004), the presence of a nonnative strain is likely responsible for the rapid increase in the distribution and abundance of this species across the continent (Saltonstall 2002). Research on the Atlantic coast has shown that this introduced lineage is a robust competitor and typically has high biomass, produces large numbers of seeds, can readily establish itself from rhizome fragments, and rapidly excludes other plant species upon invading a marsh. It is also well adapted to a range of salinity, nutrient, and hydrological conditions, and it is able to take advantage of conditions created by human modifications of marsh ecosystems (e.g., Meyerson, Saltonstall, et al. 2000; Meyerson, Vogt, et al. 2000; Warren et al. 2001; Bart and Hartman 2003; Burdick and Konisky 2003; Chambers et al. 2003; Minchinton and Bertness 2003; Silliman and Bertness 2004). Introduced Phragmites populations often are found at higher elevations within a marsh

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complex than other species (Windham and Lathrop 1999; Warren et al. 2001; Rooth, Stevenson, and Cornwell 2003). These populations may preferentially establish at higher sites that are well drained and where salinity levels and wave action may be lower. Although salinity between zero and ten parts per thousand has been shown to stimulate germination of Phragmites seeds, salinity greater than twentyfive parts per thousand inhibits germination of seeds and seedling growth (Wijte and Gallagher 1996). Similarly, growth of rhizome cuttings is negatively affected at levels of twenty parts per thousand (Adams and Bate 1999). In a study where rhizomes were buried in laboratory-simulated salt marsh habitats, Bart and Hartman (2002) found that growth depended on welldrained soils, regardless of salinity or sulfide concentrations, but salinity of eighteen parts per thousand decreased survival, growth, and biomass storage both above- and belowground. However, observations indicate that poorly drained high-salinity areas may still be invaded likely due to very high levels of plasticity in the species, and tolerance for high-salinity and anoxic conditions may vary among ecotypes (Vasquez et al. 2005). Furthermore, once established, Phragmites ventilates its rhizosphere, thus increasing soil oxygen levels and ameliorating anoxic stress (Armstrong and Armstrong 1988). To the extent that Phragmites has been used as an indicator of wetland disturbance, Bertness. Ewanchuk, and Silliman (2002) and Silliman and Bertness (2004) have argued that Phragmites occurrence in New England marshes is closely tied to upland development that decreases soil salinity because of higher freshwater inputs and increases the availability of nitrogen to the system. Similarly, Minchinton and Bertness (2003) demonstrated that disturbances that remove competing vegetation or increase nutrient availability promote the spread of Phragmites over other marsh species. Brooks et al. (2006) suggest that, more broadly, upland development in the watershed is positively associated with more extensive Phragmites occurrence in wetlands and that Phragmites has

higher nitrogen content in leaf tissue, particularly in developed areas (Meyerson 2000; Brooks et al. 2006). Although agricultural runoff has been suspected of stimulating Phragmites expansion, growth in agricultural watersheds has not been shown to be different from undeveloped watersheds (Brooks et al. 2006; King et al. 2007). An additional mechanism for successful Phragmites invasions in coastal marsh systems is associated with Spartina alterniflora aboveground biomass. Minchinton (2002a) manipulated and monitored the presence of S. alterniflora wrack and found that although the presence of wrack initially suppresses Phragmites growth, decomposition, or wrack removal of Spartina by tides releases Phragmites. In fact, Spartina wrack may facilitate spread of Phragmites by killing the underlying marsh turf (Minchinton 2002a). BIOMASS/RAPID GROWTH

Introduced Phragmites can have extremely high aboveground biomass (typical range: one thousand to four thousand grams per square meter; Meyerson, Chambers, and Vogt 1999; Meyerson, Saltonstall, et al. 2000; Talley 2001; Warren et al. 2001; Farnsworth and Meyerson 2003; Windham and Meyerson 2003), with live culm density often reaching over one hundred culms per square meter (Osgood et al. 2003). Recent work has demonstrated increasing Phragmites biomass as invasions progress and mature over time (Hunter et al. 2006). As dead culms and litter often persist for at least one additional growing season, aboveground biomass can be even higher (Meyerson 2000; Rooth et al. 2003). Belowground biomass is difficult to quantify in field populations (but see Farnsworth and Meyerson 2003) because roots and rhizomes can extend down beyond one meter in depth. Phragmites roots and rhizomes form a densely packed matrix that inhibits growth of other species and accesses additional resources. Lateral belowground growth can also be extensive as the plant sends out runners. Phragmites clones are often capable of

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expanding their cover over 10 percent per year (Haslam 1969; Rice, Rooth, and Stevenson 2000; Burdick, Buchsbaum, and Holt 2001; Havens, Berquist, and Priest 2003; Farnsworth and Meyerson 1999), allowing them to rapidly take over a site. Clonal integration, where plants establish in higher-quality habitat and then expand into less favorable ones, has also been shown to occur in this species (Amsberry et al. 2000; Bart and Hartmann 2002). Slow litter decomposition and buildup provides an additional mechanism by which Phragmites biomass affects other species in both freshwater and brackish systems (Meyerson 2000; Windham 2001; Minchinton, Simpson, and Bertness 2006).

COMPARING NATIVE AND INTRODUCED PHRAGMITES IN NORTH AMERICA A second area of research is the comparison of the autecology of native and nonnative Phragmites. Native Phragmites is typically less aggressive and appears to have less tolerance for salt water and flooding (Vasquez et al. 2005). Perhaps as a consequence, Packett and Chambers (2006) found that native Phragmites tends to occur in the tidal freshwater portions of estuaries, and Meadows and Saltonstall (2007) have documented numerous native populations in oligohaline and freshwater marshes of the mid-Atlantic coast but none in higher-salinity marshes. However, a few remnant native Phragmites populations are known to persist in New England salt marsh systems (D. Burdick, personal communication; Lambert and Casagrande 2007; Payne and Blossey 2007). While direct comparisons with native Phragmites are few, introduced Phragmites typically has higher culm density and aboveground biomass than the native under field conditions (League et al. 2006; Meadows and Saltonstall 2007). However, culm density of the native can be higher than introduced under some conditions (Meadows 2006). A recent study compared rhizome and shoot growth of native and introduced Phragmites and found that shoot emergence in the 66

introduced lineage occurred significantly earlier in the growing season, shoots were significantly more abundant, and shoots have twice the biomass relative to the native lineage (League et al. 2006). Clearly this growth strategy gives introduced Phragmites an advantage over the native. Greenhouse studies by Vasquez et al. (2005) using plants collected in Delaware have shown that introduced Phragmites has higher growth rates across a range of salinity and can survive and grow at salinity up to 0.40 M NaCl compared to just 0.13 M NaCl for two native haplotypes. Furthermore, in a one-year common garden experiment, Saltonstall and Stevenson (2007) found that while both subspecies showed a significant growth response to nutrient additions, introduced Phragmites was by far the superior performer. Introduced Phragmites grew taller, had higher culm density, and had higher above- and belowground biomass than the native under varying nutrient and salinity conditions. Under low-nutrient conditions, introduced Phragmites performed as well as the native in the high-nutrient treatment in both above- and belowground biomass, clearly supporting the hypothesis that the introduced subspecies is a superior competitor under eutrophic conditions. COMPARISONS WITH OTHER SPECIES

From the few available, direct comparisons of growth with other native wetland species (including native Phragmites), introduced Phragmites typically has higher aboveground biomass (Templer, Findlay, and Wigand 1998; Meyerson 2000; Warren et al. 2001; Windham 2001; Farnsworth and Meyerson 2003; Rooth et al. 2003), a greater capacity for rapid spread once established (Bart and Hartman 2002; Burdick and Konisky 2003), and apparently greater tolerance for fluctuating environmental conditions on an interannual basis. Introduced Phragmites also emerges earlier than many other marsh plants, and its leaves persist longer, giving it a competitive edge (Farnsworth and Meyerson 2003). When grown in field

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experiments with other common salt and brackish marsh species, Phragmites was shown to be a robust competitor across a mesohaline– polyhaline salinity gradient (Burdick and Konisky 2003). Minchinton (2002b) demonstrated that the 1997–1998 El Niño event facilitated the spread of Phragmites in two southern New England marshes, presumably because of increased freshwater inputs that ameliorated salinity stress. Interactions among factors such as disturbance and pollution, and a plastic response to changing environmental conditions and resource availability, may contribute to the success of this introduced lineage (D. Burdick, personal communication), but further study is needed to understand the characteristics that make Phragmites so successful.

IMPACTS OF INTRODUCED PHRAGMITES A number of studies have shown that once established, Phragmites will increase marsh elevation to a greater extent than other marsh species through higher accumulation of organic and mineral matter. This is largely a result of its high biomass production and high rates of litter accumulation (Harrison and Bloom 1977; Windham and Lathrop 1999; Meyerson 2000; Rooth et al. 2003). In a study comparing Phragmites with Typha and Panicum virgatum communities in a freshwater tidal marsh of Chesapeake Bay, Rooth and Stevenson (2003) found a twofold increase in litter accumulation in twenty-year-old Phragmites stands that enhanced combined mineral and organic sediment accretion rate by three to four millimeters per year over the other communities. Because Leonard et al. (2002) found that mineral sedimentation was not different among Phragmites and non-Phragmites in salt marshes of the Chesapeake Bay, organic accumulation could account for the majority of accelerated accretion. By elevating the marsh surface, hydrological flow within a marsh is modified, and loss of first-order streams may occur (Lathrop, Windham, and Montesano 2003). Decreased water flows and loss of microtopography have

detrimental effects fish and other estuarine species that typically use the marsh surface during periods of tidal flooding (Able, Hagan, and Brown 2003). BIOTA

Native Phragmites typically grows in mixed communities and rarely forms dense stands, whereas introduced Phragmites often forms a monoculture that rapidly outcompetes other species, leading to local extinction of susceptible populations. Other plant species are occasionally observed within introduced Phragmites stands (particularly near the edges), but they are likely nonreproductive. Establishment of introduced Phragmites within a marsh environment typically entails rapid change in the community. In salt marsh systems, transformation of short grass (e.g., Spartina spp.) communities to tall grass habitat is often rapid, and complete replacement of the former community can occur. In freshwater systems that are more diverse both floristically and structurally, conversion from a mixed plant community to Phragmites monoculture can also occur quickly and result in hydrological and sedimentary changes that affect animal communities using the marsh. Approximately fifty species of birds have been reported to breed in P. australis in the United States, though no Phragmites specialists have been identified in North America (Berthold et al. 1993). Phragmites has been described as an impediment to animal movement (Benoit and Askins 1999; Meyerson, Saltonstall, et al. 2000), but Phragmites may also provide protective cover to species such as muskrat (Ondatra zibethicus), wading birds (Parsons 2003), and some flightless ducks during summer molt (Ward 1942; Lynch, O’Neill, and Lay 1947). The outcomes of several studies suggest that detrimental effects of Phragmites on fish communities are ubiquitous among young-of-theyear resident nekton, with potentially important implications for long-term population sustainability and secondary production. For example,

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Hunter et al. (forthcoming) found that in the mid-Atlantic, the stage of Phragmites invasion (i.e., early, middle, or late) influences habitat quality for Fundulus spp. As an invasion progresses, habitat quality for F. heteroclitus and F. luciae appears to decline and may even result in the extirpation of the less common F. luciae in mid-Atlantic coastal marshes. At the same time, adult resident nekton have been documented with the same density among Phragmites and non-Phragmites stands unless impact on hydrology and microtopography is demonstrable (Able and Hagan 2000, 2003; Able et al. 2003; Fell et al. 2003; Meyer, Johnson, and Gill 2001; Osgood et al. 2003). For coastal marsh restoration, this result implies that the physical setting can be restored and food web function can be maintained without needing to completely eradicate Phragmites stands. Other studies have shown little or no effect of Phragmites on animal communities, and some even suggest benefits. For example, Mclary (2004) found that the abundance of the Gukensia demissa was greater in Phragmites than in S. alterniflora stands in an urban habitat. Fell et al. (1998) and Rilling, Fell, and Warren (1999) found that Phragmites and nonPhragmites areas of the lower Connecticut River provide similar food resources for typical tidal marsh invertebrates and mummichogs (F. heteroclitus). A literature survey by Weis and Weis (2003) showed that benthic biota and nekton use of S. alterniflora and Phragmites marshes were comparable. However, larval mummichogs in Phragmites stands were reduced when compared to S. alterniflora. Hunter et al. (2006) suggest that at least for Fundulus spp., the inability to detect an effect on abundance could have been due to natural variability between systems and because researchers had not considered fish life stages and invasion stages. Clearly, much of the evidence remains open to debate and suggests the need for further study. For example, relative to Phragmites, Robertson and Weis (2005) found that S. alterniflora supported higher density of stemdwelling epifaunal communities and higher 68

taxon-specific abundance and community diversity. Their data show that P. australis and S. alterniflora are not functionally equivalent for stem epifauna, at least at the two sites where the work was conducted. On the other hand, a study by Yuhas (2001) showed no differences in meiobenthos sampled from the same P. australis– and S. alterniflora–dominant marshes of the Robertson and Weis (2005) study, suggesting that some species are more likely to use emergent vegetation than others. Other studies indicate that Phragmites detritus provides valuable food resources to various species and is a component of estuarine food webs, particularly for fish (e.g., Weis and Weis 2003; Currin et al. 2003). However, work by Gratton and Denno (2006) suggests that utilization of Phragmites relative to Spartina may vary by trophic group. For example, they found evidence that Phragmites invasions may cause arthropod food webs to become detritus based instead of plant based because the herbivore assemblages the arthropods depend on are largely absent. This is reversed, however, once salt marsh vegetation is restored (Gratton and Denno 2005, 2006). More studies such as these are needed to clarify the impacts of this invasive grass on marsh systems across trophic and taxonomic groups and to assess the success of restoration for taxa other than plants. CHANGES IN NUTRIENT DYNAMICS

Phragmites has the ability to indirectly influence several aspects of nutrient cycling. Oxidation of the rhizosphere is likely to be greater under Phragmites due to convective through-flow of gases (Armstrong and Armstrong 1988) and higher transpiration rates (Windham and Lathrop 1999), which may cause phosphorus and other limiting nutrients to become bound and thus less available. Other biogeochemical processes may also influence Phragmites invasions. In brackish tidal marshes, Chambers (1997) and Chambers et al. (1999) found positive correlations between sulfide concentrations and both NH4⫹ concentrations and species occurrence in brackish marshes. These studies

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noted the exclusive presence of Spartina alterniflora in areas of high sulfide/high NH4⫹/high salt concentrations and a decrease in Phragmites height with increasing sulfide concentrations. Due to sulfide inhibition of NH4⫹ uptake, Phragmites may be restricted from high-sulfide areas in brackish and salt marshes (Chambers et al. 1999, 2003). INORGANIC NITROGEN

Windham and Meyerson (2003) reviewed the literature on the effects of Phragmites on nitrogen dynamics in the tidal marshes of the northeastern United States. In general, Phragmites increases the standing stocks of nitrogen in the systems that it colonizes—in large part because of the significantly higher biomass of Phragmites relative to other species. Nitrogen concentrations in Phragmites leaves are significantly higher than Phragmites stems (Meyerson 2000), and decomposition rates of whole plants are significantly slower than Spartina spp. or Typha spp.—typical competitors of Phragmites (Windham 1999; Meyerson 2000). Interestingly, significant differences between extractable nitrogen in soils, nitrogen in porewater, and mineralization rates of nitrogen were found in stands of Phragmites in different marsh types (i.e., freshwater, brackish, and salt). Brackish and salt marsh systems show higher concentrations of nitrogen in porewater beneath Spartina and Typha relative to Phragmites, but this does not occur in freshwater systems when Phragmites is compared to Typha or mixed species assemblages (Templer et al. 1998; Meyerson, Saltonstall, et al. 2000; Meyerson, Vogt, et al. 2000; Windham and Meyerson 2003). ORGANIC NITROGEN

While most studies on nitrogen uptake and assimilation in Phragmites stands have focused on inorganic forms of nitrogen (e.g., ammonium and nitrate), dissolved organic nitrogen (DON) has been hypothesized to be an important source of nitrogen for Phragmites and a mechanism by which Phragmites is able to retain site dominance in nitrogen-limited systems

(Meyerson, Vogt, et al. 2000). Mozder (2005) found that introduced Phragmites is capable of taking up and assimilating organic forms of nitrogen, including amino acids and urea. Other wetland species, including native Phragmites, however, also assimilate organic nitrogen, under some circumstances more efficiently than introduced Phragmites. Phragmites has a large nitrogen requirement that cannot be met by novel uptake of organic forms of nitrogen alone (T. Mozdzer, personal communication). Therefore, assimilation of DON may be one of several mechanisms working in concert with other factors (e.g., biomass accumulation, emergence before other marsh species, late-season nutrient uptake) that provides Phragmites with a competitive advantage over other plants in coastal systems (Meyerson, Vogt, et al. 2000). LIGHT

Speculation persists as to whether competition between Phragmites and other species is for light or space. Differences in light and temperature regimes in invasive Phragmites-dominated areas relative to other species have been documented by Meyerson (2000). In a comparison between invasive Phragmites and Typha spp., significant differences in available light were found throughout the canopy and above and below the detrital layer. Available light level in Typha stands was 100 percent at 1.5 meters above the marsh surface and approached between 5 and 10 percent at the marsh surface. On the other hand, available light in the Phragmites stands was less than 10 percent at 1.5 meters and dropped to zero at the marsh surface (Meyerson 2000). Clearly, available light at different strata within the stands of these two species has important implications for competition at the beginning of and throughout the growing season. The differences are due in part to plant morphology and postsenescence strategies (e.g., Phragmites stands can retain dead standing stems for two or more years), as well as biomass accumulation associated with slow rates of decomposition. Reduced light penetration into

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Phragmites stands resulted in lower air and soil temperatures relative to other parts of the same system—in some cases delaying “spring melt” at the marsh surface of the Phragmites stand by several weeks relative to other marsh vegetation (L. A. Meyerson, unpublished data). Similarly, Meadows (2006) compared light levels between native and introduced Phragmites and found that while light levels were comparable at 0.5 meters above the marsh surface, introduced Phragmites at 2.0 meters had significantly lower levels of available light despite having lower culm density and no significant differences in height from the native. However, total leaf area was higher in the introduced population, as was aboveground biomass, suggesting that the higher leaf area seen in introduced Phragmites over the native provides it with a competitive advantage.

RESTORING MARSH SYSTEMS INVADED BY PHRAGMITES? Restoration of degraded coastal systems has become increasingly important for habitat protection as pressures mount from development, population growth, and global change (Zedler 2000, D’Antonio and Meyerson 2002). The U.S. Atlantic Coast has experienced widespread changes, including the disturbance of hydrologic cycles and nutrient regimes, habitat degradation and loss, biological invasions, and reduced ecosystem services (Meyerson et al. 2005). For example, in Narragansett Bay, 65 percent of remaining coastal wetlands have been identified as candidates for restoration because of ditching and tidal restrictions (Tiner et al. 2003). Efforts to restore degraded coastal systems and to create mitigation wetlands for those lost have had mixed results. Some restoration efforts have successfully reached plant community goals or have restored underlying physical marsh processes, while others have failed to prevent Phragmites reinvasion or have not increased productivity (e.g., Farnsworth and Meyerson 1999; Warren et al. 2001).

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Many land managers and restoration practitioners want to create invasion-resistant, native wetland communities, but reaching the goal has proven elusive. The problem is often considered to be challenging because recent studies suggest native species diversity and exotic species diversity are positively correlated (Planty-Tabacchi et al. 1996; Wiser et al. 1998; Levine and D’Antonio 1999; Stohlgren et al. 1999; Symstad 2000). This is in contrast to earlier theory that argued that species-poor communities are more invasible because they have less “biotic resistance” (see Levine and D’Antonio 1999). Small-scale experimental studies suggest that high native diversity can decrease vulnerability to invasion, but largerscale investigations suggest these local effects are overwhelmed by regional factors influencing diversity (Levine and D’Antonio 1999). Clearly, large-scale monitoring and experiments are needed to determine the role of diversity and other factors in mitigating biological invasions. However, only recently has the role of propagule pressure in the success of biological invasions received significant consideration from researchers. Evidence is building that in many cases, propagule pressure can overwhelm other factors in a system that resist species invasions; therefore, this must also be seriously considered in ecological restorations (Lockwood, Cassey, and Blackburn 2005; Von Holle and Simberloff 2005). Long-term monitoring results of restoration programs to eradicate Phragmites are in notoriously short supply. Many restoration projects claim success after a short period of two or three years and then fail to report the reinvasion by Phragmites. Farnsworth and Meyerson (1999) monitored such a site for three years after two different treatments were applied to eradicate Phragmites. Over that time Phragmites clearly was reinvading the site and would again become dominant. However, the most interesting result was that the model produced from the monitoring data showed that Phragmites was systematically occupying the marsh space

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FIGURE 4.3 Eroded shoreline exposing extensive belowground growth of introduced P. australis at Reedy Island, Delaware. Photo by R. E. Meadows.

through a localized proliferation of dense rhizomes rather than sending out long tillers in advance of complete occupation (Farnsworth and Meyerson 1999) (fig. 4.3). Recent work by League, Seliskar, and Gallagher (2007) indicated that rapid bioassays of in situ Phragmites rhizomes can help managers to prioritize their efforts. They showed that rhizome color could be effectively used to predict Phragmites vitality, particularly in stands that had been previously treated with herbicides (League et al. 2007). Such studies are urgently needed to assist practitioners to better target Phragmites management and control efforts, particularly where follow-up treatments are needed to assure success. Mitigated and created wetlands frequently serve as unintentional nurseries for introduced Phragmites. Constructed tidal wetlands are engineered to encourage growth of native species, but Phragmites often establishes and spreads to the exclusion of these other species (Havens et al. 2003). As a consequence, wetlands lost to development are replaced by created wetlands dominated by Phragmites. The current U.S. policy of “no net loss” of tidal wetlands appears to favor Phragmites expansion over sustainability of wetland habitats with native plant communi-

ties. This outcome is clearly not in line with the intent of the no net loss policy. PROTECTING NATIVE PHRAGMITES?

Ironically, after extensive financial and human resources have been devoted to controlling and eradicating Phragmites, there is a groundswell to protect the remaining stands of native Phragmites, particularly in areas like the northeastern United States. A reasoned, science-based debate is urgently needed on this issue so that better management can be undertaken. Is it possible to control or eliminate populations of introduced Phragmites while protecting native populations? In the mid-Atlantic region, the U.S. Fish and Wildlife Service is using herbicides to control introduced Phragmites on the Rappahannock River, but no efforts have specifically been made to promote growth of native Phragmites, and it does not appear to colonize any space made available by the control efforts on the introduced (R. Chambers, personal communication). Other efforts in Maryland and Delaware are attempting to control introduced Phragmites populations on rivers that still have an abundance of native Phragmites populations—a promising effort still in its early stages (R. Meadows, personal communication). However, in Rhode

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Island, native stands of Phragmites are being “loved to death” by researchers who are studying them and may decline because human disturbance is creating windows of opportunity of colonization by other invasives (Meyerson 2007a, 2007b). Interspecific hybridization can lead to extinction of native flora and faunal gene pools and also has been shown to produce increased aggressiveness in some hybrid forms relative to their parental lineages. Hybridization between native and introduced species can potentially swamp out the native gene pool (Rhymer and Simberloff 1996; Antilla et al. 2000; Vila, Weber, and D’Antonio 2000; Pooler, Dix, and Feely 2002), but interbreeding among subspecies has been virtually ignored as a serious threat to native species (D’Antonio, Meyerson, and Denslow 2001). Because native and introduced Phragmites australis are subspecies, it seems possible that cross-mating could occur in the wild, but no evidence existed for this outcome (Saltonstall 2003b). However, the production of hybrids by Meyerson et al. (in press) in a greenhouse study makes clear that interbreeding is indeed possible. Conservation of native biological diversity is a widely held societal value, as evidenced by dedication of resources and various forms of legislation that seek to protect biodiversity. However, there is also wide recognition that resources are limited, and therefore priorities for conservation and management of species must be identified. Setting priorities for management of introduced Phragmites should include an assessment of potential loss of native Phragmites populations through interbreeding with introduced populations. Where this potential for hybridization exists, the introduced stands should receive priority for management and control, and the native Phragmites populations should receive priority for conservation by conferring protected status on some of the remnant native Phragmites populations (Meyerson 2007a, 2007b). Because current knowledge on the ecology of native Phragmites is limited, management strategies that would promote the growth of na72

tive Phragmites over the introduced form cannot be implemented at this time. The rhizomes of native Phragmites tend to be small relative to the introduced type, are more sparsely distributed, and can undergo intense competition from the high diversity of wetland plants in oligohaline and tidal freshwater marsh systems—all factors that are likely to inhibit the natural spread of native Phragmites. Like introduced Phragmites, production of viable seed in native populations is highly variable and varies interannually (K. Saltonstall, unpublished data). To date, native Phragmites has not been used in marsh restoration efforts, so its ability to survive and prosper in restored systems is unknown. More basically, we do not yet understand which native genotypes should be used in marsh restoration, which habitats are most suited for native Phragmites, and what other native plants would best suit a marsh system that was intended to encourage the growth of native Phragmites. In the absence of growth information on native Phragmites, the precautionary principle should be applied to prioritize preservation of remaining stands of native Phragmites.

FUTURE OUTLOOK AND NEEDED RESEARCH IS BIOCONTROL FOR INVASIVE PHRAGMITES POSSIBLE?

The majority of Phragmites management programs use conventional herbicide treatments applied using either helicopter or truck sprayers, usually in the autumn after plants have flowered. This is sometimes combined with burning of remaining aboveground biomass in the winter months. Because of the high volume of underground reserves, a large monoculture is difficult to kill with a single application of herbicide, and follow-up treatments in subsequent years are often necessary (Norris, Perry, and Havens 2002). Another alternative control method currently under study is the use of insects as a biocontrol agent for managing Phragmites

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(Blossey 2003). Surveys of insect communities found on introduced Phragmites populations have revealed that while twenty-six arthropod species are found feeding on the plant, only five species are native to North America (Casagrande, Balem, and Blossey 2003). Of these, several are under consideration as possible biocontrol agents. In a study of the lepidopteran Rhizedra lutosa, which is native to Europe, Casagrande et al. (2003) found this species can be found in low density throughout the Atlantic Coast region and thus does not have a significant impact on the growth of Phragmites. Tewksbury et al. (2002) developed a list of other insect species under consideration for biological control, all of which are native to Europe. Preliminary testing of several of these species is currently underway in Europe to determine their impacts on both native and introduced Phragmites from North America (B. Blossey, personal communication). However, the difficulties associated with identifying lineage-specific control organisms that will selectively target introduced over native populations may hinder the development of an effective biocontrol program. GLOBAL CHANGE

Current levels of atmospheric CO2 are less than optimal for C3 plant growth; therefore, recent and projected increases in CO2 are expected to stimulate the growth of many plant species (Ziska and George 2004), including Phragmites (Farnsworth and Meyerson 2003). On the whole, invasive plants are expected to respond with greater growth rates than noninvasive plants to rising levels of atmospheric CO2, with the potential for preferential selection of invasive plants within plant communities (Ziska and George 2004). However, the number of experiments that have specifically addressed the effects of rising CO2 on invasive plants is relatively small, and therefore evidence remains anecdotal. Pattison, Goldstein, and Ares (1998) found that, as with Phragmites, several invasive plant species in Hawaii shared higher relative growth rates and maximal photosynthetic rates than

their less invasive counterparts. Earlier leaf emergence and greater leaf longevity relative to native vegetation, observed for introduced Phragmites (Farnsworth and Meyerson 2003), characterize several species of exotic shrubs studied around the world (Milton 1981; Harrington et al. 1989). These traits also have important implications for the relative ability of these species to respond to facets of global change, including increased carbon availability, rising temperatures, and sea-level rise. Phragmites appears to efficiently convert CO2 and available nutrient resources to aboveground biomass in both freshwater and brackish marsh types (Farnsworth and Meyerson 2003). Phragmites and other species may thrive under rising CO2 scenarios and alter the carbon and nutrient cycling of wetlands (Findlay, Dye, and Kuehn 2002). Little is known of how species composition affects carbon dynamics in midlatitude wetlands (Gross et al. 1993; Brix 1999; Matamala and Drake 1999), yet changes in species composition of coastal marshes will determine the capacity of wetlands to act as carbon sources and sinks. As Phragmites comes to dominate coastal marshes, a long growing season, high photosynthetic rate, and long-term storage in standing dead stems may cause these marshes to act increasingly as carbon sinks. Similarly, as atmospheric CO2 increases, Phragmites may not be limited in its long-term CO2 responsiveness by nitrogen availability, due to its ability to sequester high concentrations of nitrogen in leaf tissues, litter, and soil (Mason and Bryant 1975; Meyerson 2000; Farnsworth and Meyerson 2003).

PHRAGMITES AS A BIOENGINEER Rapid sea-level rise resulting from global change may make preservation of coastal marsh systems (and the ecosystem services that they provide) a significant challenge. Phragmites has been used as a bioengineer in Europe to prevent soil erosion and to protect shorelines (Bakker 1960), and similar benefits of Phragmites colonization have been suggested in the United

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States (Rooth and Stevenson 2000; Rooth et al. 2003). As described earlier, Phragmites facilitates marsh accretion via biomass accumulation that traps marsh sediments, altering the marsh hydrology and decreasing the frequency of tidal flooding (Windham and Lathrop 1999; Meyerson, Vogt, et al. 2000; Rooth et al. 2003), and Phragmites has been shown to oxidize the rhizosphere via the convective through flow of gases (Armstrong and Armstrong 1988). These characteristics, in combination with its tolerance for a wide range of environmental conditions and ability to reproduce vegetatively, suggest that Phragmites may keep pace in marsh systems where sea-level rise has the greatest impact (Rooth and Stevenson 2000; Rooth et al. 2003; Farnsworth and Meyerson 2003). The irony of this potential outcome highlights that it is both values and circumstances that determine whether an introduced species is considered to have beneficial or harmful effects on an ecosystem. Each site must be specifically considered for species, community, and habitat management; for ecosystem services rendered; and for trade-offs associated with management and control of Phragmites. Agriculture provides another example. A relatively large amount of research has been conducted on nitrogen concentrations and Phragmites. However, less work has been dedicated to understanding the role of Phragmites in removing phosphorus and other elements that result from agricultural runoff before they are released to unmanaged areas, particularly in freshwater systems (Meyerson, Saltonstall, et al. 2000). Studies conducted in terrestrial systems have shown the potential importance of certain plant species in modulating and controlling ecosystem nutrient cycles when they contribute disproportionately to the cycling of an element compared to other plants growing in the same environment (Dahlgren, Vogt, and Ugolini 1991; Vitousek 1990). Plants may accumulate and increase the cycling of trace metals (e.g., iron, manganese) that can be directly toxic to other species or indirectly result in soil nutrients becoming unavailable because of the recal74

citrant complexes that they form with these nutrients (Dahlgren et al. 1991; Meyerson, Saltonstall, et al. 2000). Studies in brackish and salt marsh systems have noted the enhanced capacity of Phragmites to concentrate metals (e.g., nickel, cadmium, lead, copper, and iron) in both above- and belowground tissues relative to other plants and have noted their potential for phytoremediation in these systems (Kraus 1987; Weis and Weis 2003). Others have addressed the use of Phragmites for phytoremediation in constructed wetlands receiving landfill leachate and have found that most metals accumulate in the roots (e.g., Peverly, Surface, and Wang 1995), creating a stable sink for these elements. However, research that investigates the sequestration of nutrients and metals released from agricultural lands by Phragmites has not been carried out. If Phragmites is performing this ecosystem service, the removal of Phragmites stands that abut agricultural land may in fact result in unintended further degradation of the ecosystem.

CONCLUSION Unlike most other invasive species in the United States, Phragmites has been carefully studied in terms of its origins, distribution across the landscape, ecology and ecophysiology, and effects on nutrient cycles, habitat, and other resources. Significant and important advances in Phragmites research have been made in the last decade, particularly the identification of the introduced lineage. While much work remains to be accomplished to complete our understanding of the causes and consequences of introduced Phragmites invasions, the body of knowledge about Phragmites could serve as a template for more systematic and complete investigations of other invasive plant species. For example, Phragmites has been examined at the population, community, and ecosystem levels; and biogeographic comparisons of different populations, both native and exotic, have been conducted. Such a generalized road map for invasive plant research could help to guide future

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research and to fill knowledge gaps that are species-specific and that also apply more generally to invasive plants overall. Despite the voluminous work that has been conducted on Phragmites, further research is needed to pro-

vide better and more complete information to coastal managers who conduct the on-theground work with Phragmites on a daily basis. Some of the information gaps and research needs are presented in table 4.1.

TABLE 4.1 Examples of information gaps on Phragmites and associated research that could begin to fill these gaps

Information Gap

Research Need

Mechanism of invasion success

• Clonal resource allocation and integration • Pathways and scale at which Phragmites disperses and spreads • Role of propagule pressure

Causes of extensive dieback in Europe and expansion in North America

• Role of abiotic factors • Role of pathogens and pests

Reproductive biology and propagule survival

• Frequency that new stands are established through sexual reproduction • Survival and spread of Phragmites propagules across a range of environmental conditions

Potential for intraspecific hybridization

• Determination of cross-breeding between the subspecies • Methods for field identification of hybrid populations

Landscape-level analysis of distribution and impacts

• Data integration and meta-analysis of existing data • Greater use of aerial and remotely sensed data to cover larger areas • Linking of land use change and water quality change with increased or reduced colonization of introduced Phragmites

Ecosystem impacts on energy flows and nutrient cycles (microbial dynamics)

• Effects of Phragmites on food webs

Effects on other species (community composition and structure)

• More quantification of the changes Phragmites colonization has on native communities (e.g., plants, infauna, benthos, fish, birds, etc.) • Quantitative comparisons of the effects of native versus introduced Phragmites on community structure and species richness

Appropriate scale of control and effective nonchemical forms of control

• How does the scale of an infestation affect management and control efforts? • Further research and monitoring on nonchemical methods to control Phragmites such as the use of rye grass mats

Interaction of introduced Phragmites with other invasive species

• Positive or negative interactions of introduced Phragmites with other invasive plants and animals

Beneficial uses for Phragmites

• Wastewater treatment using native Phragmites • Thatch and reed mats

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University of Massachusetts–Amherst, the University of Rhode Island, and the Narragansett Bay Estuary Program. National Wetlands Inventory Cooperative Interagency Report. Torrey, J. 1843. Flora of the State of New York. Albany: Carroll & Cook. Tscharntke, T. 1999. Insects on common reed (Phragmites australis): Community structure and the impact of herbivory on shoot growth. Aquatic Botany 64: 399–410. Tucker, G. C. 1990. The genera of Arundinoideae (Graminae) in the southeastern United States. Journal of the Arnold Arboretum 71: 145–177. van der Putten, W. H. 1997. Die-back of Phragmites australis in European wetlands: An overview of the European Research Programme on Reed Dieback and Progression (1993–1994). Aquatic Botany 59: 263–275. Vasquez, E. A., E. P. Glenn, J. J. Brown, G. R. Guntenspergen, and S. G. Nelson. 2005. Salt tolerance underlies the cryptic invasion of North American salt marshes by an introduced haplotype of the common reed Phragmites australis (Poaceae). Marine Ecology Progress Series 298: 1–8. Vila, M., E. Weber, and C. M. D’Antonio. 2000. Conservation implications of invasion by plant hybridization. Biological Invasions 2: 207–217. Vitousek, P. M. 1990. Biological invasions and ecosystem processes: Towards an integration of population biology and ecosystem studies. Oikos 57: 7–13. Von Holle, B., and D. Simberloff. 2005. Ecological resistance to biological invasion overwhelmed by propagule pressure. Ecology 86: 3212–3218. Waksman, S. A., H. Schulhoff, et al. 1943. The Peats of New Jersey and Their Utilization. Trenton, NJ: Department of Conservation and Development. Ward, E. 1942. Phragmites management. Transactions of the North American Wildlife Confederation 7: 294–298. Warren, R. S., P. E. Fell, J. L. Grimsby, E. L. Buck, G. C. Rilling, and R. A. Fertik. 2001. Rates, patterns, and impacts of Phragmites australis expansion and effects of experimental Phragmites control on vegetation, macroinvertebrates, and fish within tidelands of the Lower Connecticut River. Estuaries 24: 90–107. Weis, J. S., and P. Weis. 2003. Is the invasion of the common reed Phragmites australis, into tidal marshes of the eastern U.S. an ecological disaster? Marine Pollution Bulletin 46: 816–820. White, D. A., D. P. Hauber, and C. S. Hood. 2004. Clonal differences in Phragmites australis from the Mississippi River Delta. Southeastern Naturalist 3: 531–544.

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Wijte, A. H. B. M., and J. L. Gallagher. 1996. Effect of oxygen availability and salinity on early life history stages of salt marsh plants. I. Different germination strategies of Spartina alterniflora and Phragmites australis (Poaceae). American Journal of Botany 83: 1337–1342. Wilcove, D. S., D. Rothstein, J. Dubow, A. Phillips, and E. Losos. 1998. Quantifying threats to imperiled species in the United States. BioScience 48:607–617. Willis, O. R. 1874. Catalogue of Plants Growing without Cultivation in the State of New Jersey. New York: Schermerhorn. Windham, L. 2001. Comparison of biomass production and decomposition between Phragmites australis (common reed) and Spartina patens (salt hay grass) in brackish tidal marshes of New Jersey, USA. Wetlands 21: 179–188. Windham, L., and R. G. Lathrop. 1999. Effects of Phragmites australis (common reed) invasion on aboveground biomass and soil properties in a brackish tidal marsh of the Mullica River. Estuaries 22: 927–935.

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PART II

Human Inputs and Consumer Effects

5

Opportunistic Herbivores, Migratory Connectivity, and Catastrophic Shifts in Arctic Coastal Systems Hugh A. L. Henry and Robert L. Jefferies Overgrazing has contributed to the degradation of numerous terrestrial systems. Here, we summarize how feedbacks among grazers, plants, and soil have led to major losses of salt marshes that frequently occur along a coastline of two thousand kilometers in the Hudson Bay Lowlands. Agricultural food subsidies to overwintering and migrating lesser snow geese in the United States and Canada have led to high densities of geese on their Arctic breeding grounds. Breeding and staging geese feed destructively in early spring, leading to a widespread loss of Arctic salt- and freshwater coastal vegetation. The resulting exposed intertidal mudflats represent an alternative stable state where an entire successional stage has been lost. We examine the mechanisms whereby large numbers of geese impact soil processes and plant growth in these systems and compare them to vertebrate grazing systems in other polar regions. Vegetation loss caused by goose grubbing and intense grazing result in soil physical changes, such as erosion and compaction of sediment, that lead to desertification of the system. The increased evaporation from exposed sediment produces widespread hypersalinity and decreases in soil moisture that restrict plant establishment and growth. These soil changes also disrupt soil nitrogen fixation and nitrogen mineralization, further restricting plant growth. The coalescence of patches of exposed sediment at increasingly larger spatial scales maintains the alternative stable state. Spring hunts and modified hunting season regulations have been implemented to reduce goose population size and to limit the degradation of remaining Arctic coastal marshes. Vegetation surveys are essential to monitor ongoing habitat change in response to any reduction in foraging. Alternative management strategies for reducing goose population size are under active discussion, but each of these measures may be difficult to implement or poses risks to other wildlife.

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Grazers reduce the standing crop of vegetation and alter plant community composition directly by feeding selectively on palatable species. However, interactions between plants and herbivores can also lead to changes in the physical and chemical properties of soil that feed back on plant and animal growth (Rietkerk and Van de Koppel 1997). Herbivores indirectly affect soil microclimate and nutrient cycling rates by modifying vegetation structure and rates of plant litter input (Pastor and Naiman 1992; Augustine and McNaughton 1998; Stark et al. 2000), but they also impact soils directly by excreting waste products (Bazely and Jefferies 1985; Hamilton et al. 1998) and by trampling (Hirneaux et al. 1999; Cumming and Cumming 2003). At low to moderate herbivore densities, the nutrients in fecal inputs can stimulate compensatory plant growth, resulting in increased primary production, despite a decreased standing crop (Cargill and Jefferies 1984a, 1984b; Frank and Evans 1997). Grazing systems can remain in a productive state provided that herbivore densities are restricted to these moderate levels. However, at high herbivore densities, these systems can exhibit sudden instability, leading to the development of an alternate stable state (Noy-Meir 1975; Van de Koppel, Rietkerk, and Weissing 1997). Although the demonstration of the latter in ecological systems has received considerable debate (e.g., Connell and Sousa 1983; Sutherland 1990; Petraitis and Latham 1999), here the alternate states are defined as stable, because they persist as exposed, degraded soil. Degradation leading to desertification (sensu Graetz 1991) is driven and maintained by positive feedbacks between declines in plant cover associated with heavy grazing and decreased soil quality (Graetz 1991; Rietkerk and Van de Koppel 1997). Many examples of grazing-induced degradation can be attributed to overgrazing by domestic livestock, as indicated by the desertification of semiarid grasslands in Africa (Milton and Dean 1995; Kerley, Knight, and De Kock 1995), the southwestern United States (Hess and Holechek 1995), Russia (Zonn 1995), and Australia (Ludwig and Tongway 1995). In these systems, 86

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water and food subsidies and protection from disease and predation have permitted artificially high stocking densities. However, in years of low rainfall when productivity declines, high foraging intensity results in exposure of soil and large reductions in standing crop, initiating positive feedbacks that result in soil degradation (Sinclair and Fryxell 1985; Schlesinger 1990). Whereas the effects of domesticated livestock grazing are mostly localized, migratory animals can use agricultural crops as a nutrient subsidy and subsequently exert high grazing pressure in distant, nonagricultural systems, as a result of “migratory connectivity” (see Webster et al. 2002). In recent decades, agricultural food subsidies to lesser snow geese on their wintering grounds and along their flyways have led to an increase in population size and to high densities of geese on their Arctic breeding grounds (Jefferies, Henry, and Abraham 2004; Abraham, Jefferies, and Alisauskas 2005). Increased goose foraging has promulgated the widespread degradation of Arctic salt marshes that represents a scale difference in the use of space by the geese from the regional level on the wintering grounds to the local level at the breeding grounds (see de Roos, McCauley, and Wilson 1998). The loss of vegetation in coastal marshes can be detected with LANDSAT imagery (Jano, Jefferies, and Rockwell 1998), and a recent survey indicates that at nine coastal sites, where geese breed, over thirty-five thousand hectares of vegetation have been lost (Jefferies, Jano, and Abraham 2005). The total area of vegetation loss for the entire southern and western coastlines of Hudson Bay and the coastline of James Bay is well in excess of this estimate. Is this large-scale conversion of vegetated marsh to mudflats reversible, or does it represent an alternate stable state? Schröder, Persson, and de Roos (2005) have concluded that experimental evidence of the existence of alternative stable states at La Pérouse Bay, Manitoba, meets the criteria they have established for a switch in states. In this chapter, we examine the mechanisms whereby migratory connectivity has led to soil degradation of the

marshes. We focus on plant–soil feedbacks initiated by foraging and compare similar feedback processes in different polar grazing systems in order to assess the resilience of these systems to herbivory.

THE IMPACT OF MIGRATORY CONNECTIVITY, AGRICULTURAL NUTRIENT SUBSIDIES, AND EXPLODING GOOSE POPULATIONS ON THE DESTRUCTION OF ARCTIC SALT MARSHES The midcontinent population of the lesser snow goose (Anser caerulescens caerulescens A.O.U.; fig. 5.1) migrates from wintering grounds in the southern United States and Mexico to the coastal breeding grounds in the eastern and central Canadian Arctic. The population has increased geometrically in recent decades at approximately 5 to 7 percent per annum (fig. 5.2), and estimates of the current population size in late fall are in excess of five million (Cooke, Rockwell, and Lank 1995; Abraham et al. 1996). This rapid population increase has been attributed to increased winter survival (Francis 1999), explained by the birds’ ability to benefit from changes in agricultural practices (Boyd, Smith, and Cooch 1982; Jefferies, Klein, and Shaver 2004; Abraham et al. 2005). Before the 1940s, the wintering population was restricted

FIGURE 5.1 The lesser snow goose (Anser caerulescens caerulescens A.O.U.). Photo courtesy of Hudson Bay Project.

to coastal marshes of Louisiana and Texas (Bent 1925; McIlhenny 1932; Lynch 1975), but it has since expanded into rice prairies and corn (maize) fields in these states, as well as in Arkansas, Missouri, Iowa, Kansas, and Nebraska (Stutzenbaker and Buller 1974;

FIGURE 5.2 Midwinter index of lesser snow geese and Ross’s geese in the midcontinent population. Data from the Mississippi and Central Flyways courtesy of K. Gamble, U.S. Fish and Wildlife Service, Columbia, Missouri, and D. Sharp, U.S. Fish and Wildlife Service Denver, Colorado.

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FIGURE 5.3 Evidence of goose grubbing for grass roots and rhizomes. Photo courtesy of Hudson Bay Project.

Hobaugh, Stutzenbaker, and Kaminski 1989; Robertson and Slack 1995). The geese also stop over in agricultural fields along the flyways during the fall and spring migrations (Burgess 1980), and as a result, arrival of the birds at their wintering grounds has been delayed from late October to early December each year. Flyway counts of lesser snow geese from 1950 to 1990 are highly correlated with increases in nitrogen fertilizer use and yields of rice, corn, wheat, and soybean (Jefferies et al. 2004; Abraham et al. 2005), although the latter crop is not widely eaten by geese (Krapu, Brandt, and Cox 2004). Nutrient subsidies have improved both the quantity and quality of food available in croplands, and the geese benefit from “on demand,” readily available sources of nitrogen and carbohydrates. There is an extensive network of U.S. National Wildlife Refuges, state wildlife management areas, and some private sanctuaries in this agricultural landscape, which are used by the geese (Lynch 1975; Bateman, Joanen, and Stutzenbaker 1988; Abraham et al. 2005), both on the wintering grounds and along the migration routes. The refuges, which are frequently large wetland complexes adjacent to agricultural fields, provide food and protection for the

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birds. Increased flexibility in movements of birds, particularly in response to land use changes (refugia and agricultural fields) since the 1950s has led to shifts in migratory pathways (Alisauskas, Ankney, and Klass 1988; Alisauskas 1988; Widner and Yaich 1990). These changes have resulted in more frequent stopovers and reduced long-haul migration between the wintering and breeding grounds (Johnsgard 1974; Dzubin, Boyd, and Stephen 1975; Francis and Cooke 1992). The midcontinent population of lesser snow geese are now less vulnerable to unpredictable climatic events than when they migrated over longer distances to fewer staging or winter sites (Alisauskas 2002; Krapu and Schultze, forthcoming). The rise in numbers of lesser snow geese does not appear to have been caused by a reduction in harvesting. Between 1970 and 1998, the total number of birds from the midcontinent population harvested annually fluctuated between three hundred thousand and seven hundred thousand (Cooke et al. 1999; Kruse and Sharp 2002). However, over the same period, the proportion of the postbreeding population that was harvested failed to rise at the same rate as the overall growth of the midcontinent population; thus, the impact of hunting on overall

mortality has likely declined. Although the establishment of refugia along flyways, reduced hunting pressure and climate amelioration (MacInnes et al. 1990) may have contributed to increased survival and reproductive success over this time, agricultural nutrient subsidies appear to have been the primary factor driving the rapid population increase (Jefferies et al. 2004; Abraham et al. 2005). Female lesser snow geese are philopatric, and population increases have resulted in high nesting densities and geographic expansion of breeding colonies in coastal areas of Hudson Bay (Abraham and Jefferies 1997). Intense foraging on the wintering grounds and during migration takes place in order for the birds to meet the energy and nutritional demands of migration, egg laying, and incubation. In particular, geese require protein for egg laying, yet protein is heavy (it contains 77 percent water by weight). Hence, the protein demand for reproduction cannot necessarily be carried by migrating birds over long distances. Arctic breeding geese forage intensively on arrival in northern latitudes, often within the vicinity of the breeding grounds, in order to meet the protein requirements for reproduction. Geese arriving early in spring prior to the onset of plant growth are restricted to grubbing for roots and rhizomes in thawed intertidal marshes (fig. 5.3). In coastal freshwater marshes, shoot pulling of sedges is common where melt has occurred. The shoot is discarded after the nutrient-rich basal portion is consumed. Both grubbing and shoot pulling are highly invasive and result in exposed sediment and peat. At low to moderate goose densities, exposed sediment patches created by grubbing are small (less than twenty centimeters) and can be readily recolonized by adjacent plants (McLaren and Jefferies 2004). However, in years with cold, late springs, the combination of the added pressure of staging birds waiting to migrate farther north and delays in aboveground plant growth increase grubbing intensity, resulting in the coalescence of small patches into increasingly larger areas

of exposed sediment (McLaren and Jefferies 2004; figs. 5.4 and 5.5). Thus, the creation of large patches of exposed sediment may be triggered, at least in part, by climatic events (see Skinner et al. 1998).

FIGURE 5.4 An exclosure established in 1981 by D. R. Bazely and R. L. Jefferies. Loss of vegetation has occurred outside the exclosure since 1981 as a result of the grubbing activities of the geese and the associated edaphic changes. The exclosure plot from which the geese are excluded has remained vegetated.

FIGURE 5.5 The coalescence of grubbed patches at the margins of intertidal marsh ponds. Photo courtesy of Hudson Bay Project.

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SOIL PHYSICAL CHANGES IN DEGRADED SALT MARSH Following the creation of large patches of exposed sediment by goose grubbing, soil degradation is driven and maintained by feedbacks between declines in plant cover and decreased soil quality (Srivastava and Jefferies 1996 fig. 5.6). As observed for high densities of livestock in pastures (Bromley et al. 1997; Rietkerk and Van de Koppel 1997), exposed soil becomes compacted, and decreased water infiltration promotes surface runoff (McLaren and Jefferies 2004). While trampling by livestock can contribute significantly to soil compaction in heavily grazed pasture (Hirneaux et al. 1999; Cumming and Cumming 2003), vegetation removal leads to soil compaction in this and other grazed Arctic salt marshes (McLaren and Jefferies 2004), probably as a result of collapsed root channels, hypersaline conditions, loss of organic matter, and trampling by geese. Most salt marsh sites are flooded early in the growing season. However, low soil moisture, caused by poor water infiltration and increased evaporation, restricts plant regrowth in patches of exposed soil later in the season when soils dry out (Wilson and Jefferies 1996).

FIGURE 5.6 Positive feedback leading to reduced graminoid cover in a goose-grazed Arctic salt marsh. Following the creation of large patches of exposed sediment by goose grubbing, soil degradation is driven and maintained by feedbacks between declines in plant cover and decreased soil quality (solid lines). Heavy grazing can open up remaining swards, leading to increased evaporation, high salinity, and the coalescence of bare patches. Thus, overgrazing in response to a reduced sward area initiates a positive feedback between the remaining plant cover and hypersalinity (dotted line).

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A further consequence of increased evaporation from exposed sediment is the development of hypersaline soils (Iacobelli and Jefferies 1991; Srivastava and Jefferies 1995a, 1995b). Salt marshes along the Hudson Bay coast overlie marine sediments, and high rates of evaporation from patches of exposed soil draw marine salts to the soil surface, which interfere with plant growth and reestablishment (Srivastava and Jefferies 1995a, 1995b, 1996; Handa and Jefferies 2000). Both the loss of soil moisture and the high salinities are sensitive to patch size. Below a threshold size of approximately twenty centimeters in diameter, evaporation is reduced by surrounding vegetation, and recolonization of the patch readily occurs, in contrast to the absence of growth in patches greater than twenty centimeters in diameter (McLaren and Jefferies 2004). With low-intensity grubbing, patches of exposed soil are scattered and transient, whereas high-intensity grubbing leads to the development of large persistent patches of exposed sediment. Although patches do not expand into intact swards of vegetation in the absence of further grubbing, heavy grazing can open up remaining swards, leading to increased evaporation, high salinity, and the coalescence of bare patches (Bazely and Jefferies 1997).

Decreased plant re-establishment on bare patches

Intense grubbing/ overgrazing

High goose densities

Decreased plant N availability

Reduced graminoid cover

Increased soil compaction

Increased evaporation from bare soil

Increased water runoff

Increased soil salinity Decreased N mineralization

Increased erosion of organic layer

Decreased soluble organic N

Decreased N Fixation inputs

human inputs and consumer effects

Decreased soil moisture Drying and erosion of cyanobacterial mats

Wintering grounds south-east U.S.A.

Summer nesting grounds Hudson Bay Lowlands, Canada Original site

Increased goose population

New site TOP-DOWN

TOP-DOWN Decreased vegetation cover

Decreased vegetation cover

Soil degradation

Soil degradation

BOTTOM-UP

BOTTOM-UP

Agricultural food subsidy BOTTOM-UP

Thus, overgrazing in response to a reduced sward area initiates a positive feedback between the remaining plant cover and hypersalinity (fig. 5.6, dotted line). Overall, these bottom-up effects of soil degradation are an indirect consequence of a topdown grazing effect, itself driven by bottom-up processes in the goose wintering grounds and along their flyways (fig. 5.7). Although female geese are phylopatric, they shift from degraded areas to new, less degraded nesting and broodrearing sites. Relocation of the nesting colonies allows the geese to escape density dependence and maintain high population densities, but it has resulted in an increasing expanse of degraded salt marsh along the western and southern Hudson Bay coasts (Jefferies et al. 2005).

SOIL NUTRIENT AVAILABILITY IN DEGRADED AND INTACT SALT MARSHES As in most terrestrial systems, plant growth in the salt marshes of Hudson Bay is limited by nitrogen availability (Cargill and Jefferies 1984b). However, addition of nitrogen quickly leads to phosphorus limitation, and if these two nutrients are added together to experimental plots, aboveground biomass production is high. Low inorganic nitrogen and phosphorus availability are the result of low mineralization rates, which

FIGURE 5.7 Integration of bottom-up and top-down processes linked to the degradation of Arctic salt marsh. Dotted lines indicate expansion into new nesting and brood-rearing sites in subsequent years, following the degradation of the original site. This ongoing site switching allows the goose to avoid density-dependent effects.

result from cold soil temperatures, and the rates of nitrogen mineralization, at least, are further depressed in exposed sediments by the high salinities present in summer (Wilson et al. 1999). When 15NH4⫹ is added to sediments, the largest fraction after twenty-four hours of in situ incubation is located in the microbial pool, rather than the plant pool. However, in both the degraded intertidal and supratidal soils, the proportion of isotope in the microbial fraction falls compared with that in sediments beneath intact swards, and the amount of 15N remaining in the soil solution is high (Buckeridge 2004). Hence, microbial activity is reduced in these exposed sediments. Soluble organic nitrogen can be an important alternative source of nitrogen for plants in cold climates (Chapin, Moilanen, and Kielland 1993; Lipson and Näsholm 2001), and salt marsh plants readily take up soluble amino acids (Henry and Jefferies 2003). However, erosion of the surface organic layer in exposed soils reduces the pool of total nitrogen (Wilson and Jefferies 1996; McLaren and Jefferies 2004), and amino acid concentrations generally are low, consistent with the reduced microbial activity already mentioned (Henry and Jefferies 2002). While goose fecal inputs contribute substantially to the soil soluble pool of organic and inorganic nitrogen (Bazely and Jefferies 1985;

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Ruess, Hik, and Jefferies 1989; Henry and Jefferies 2002), the deposition of fresh feces tends to be concentrated on vegetated patches rather than exposed sediment. Where feces are present on sediment, higher surface soil temperatures are known to increase the rate of nitrogen volatilization from droppings (Schlesinger 1990).The primary net input of nitrogen into the system occurs via cyanobacterial nitrogen fixation on the soil surface beneath graminoid swards (Bazely and Jefferies 1989; Walker et al. 2003). Goose grazing can promote increased nitrogen fixation in intact swards by alleviating shading of the sediment surface and the buildup of plant litter (Bazely and Jefferies 1989). Cyanobacteria also grow on exposed sediments devoid of vegetation, particularly where the thin veneer of surface organic matter persists. However, when surface sediments dry out later in the season, high rates of evaporation and the absence of vegetation accelerate the drying and wind erosion of cyanobacterial mats, and little of the fixed nitrogen is likely incorporated into the remaining surface organic layer of exposed soils (Walker et al. 2003). As mentioned, the salt marsh is primarily nitrogen limited (closer to phosphorus saturation than nitrogen saturation), but the adjacent freshwater marshes are phosphorus limited (Ngai and Jefferies 2004). As a result, shoots of the freshwater sedge, Carex aquatilis, have significantly higher nitrogen: phosphorus ratios than those of salt marsh grass, Puccinellia phryganodes. Shoots of the grass also have higher concentrations of calcium, potassium, and manganese, as well as lower carbon: nitrogen and carbon: phosphorus ratios than those of the sedge (Ngai and Jefferies 2004). It is possible that the decline in gosling weight during recent decades (Cooch et al. 1993) is not just a reflection of a deteriorating resource base linked to destructive foraging, but also a consequence of a smaller structural size. Increasingly inadequate intakes of phosphorus and calcium as the birds are forced to forage more in freshwater sedge meadows may contribute to a decline in skeletal mass. 92

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Microbial carbon limitation also may play an important role in regulating nitrogen mineralization in Arctic soils. Recent studies in Arctic and alpine sites have revealed that more than half of the annual soil respiration can take place under the snow pack (Grogan and Chapin 1999). This high soil respiration over winter corresponds with a high microbial biomass, a large proportion of which is fungal (Lipson, Schadt, and Schmidt 2002), including several newly characterized clades (Schadt et al. 2003; Lipson and Schmidt 2004). Similar high levels of microbial biomass have been recorded beneath the snow pack in intertidal and freshwater marshes on the Cape Churchill Peninsula (Buckeridge 2004; Hargreaves 2005; Edwards, personal communication). The winter communities of soil microorganisms differ functionally from summer communities, and they decompose unique profiles of plant litter and soil organic matter compounds (Lipson et al. 2000; Mikan, Schimel, and Doyle 2002). Their activity may influence the quantity and quality of nitrogen compounds available during the plant growing season (Grogan et al. 2004; Schmidt and Lipson 2004). The winter microbial community dies off in the early spring, and a large proportion of annual nitrogen turnover occurs at this transition period between winter and summer microbial communities. However, at moderate levels of grazing in summer, increased exudation of labile carbon compounds from roots into the rhizosphere can occur (Dyer et al. 1991; Bardgett, Wardle, and Yeates 1998) that can lead to enhanced microbial activity, including increased mineralization rates. The increased turnover of the microbial pool and fine root fraction in the summer at moderate levels of goose grazing, in part, may drive compensatory plant growth in grazed plots by providing a supply of nitrogen and phosphorus. Geese also may promote litter decomposition and nitrogen mineralization by accelerating the incorporation of the previous year’s litter into the soil, as a result of their trampling (Zacheis, Hupp, and Ruess 2001; Zacheis, Ruess, and Hupp 2002).

RECOLONIZATION OF DEGRADED SALT MARSH: IS THE DAMAGE REVERSIBLE? Soil physical and chemical changes limit the abilities of goose forage species to recolonize bare patches, even in the absence of further grazing. In assisted revegetation trials, transplants established in degraded sediments only when soils had adequate moisture and low salinities (Handa and Jefferies 2000). Growth of plants increased significantly when combined additions of peat mulch and inorganic nitrogen and phosphorus were added to the consolidated sediments. Vegetative fragments of the goose forage grass, Puccinellia phryganodes, can establish in fresh unconsolidated sediment (the grass is a sterile triploid in North America; hence, the species can only spread by clonal propagation). In less degraded sites, goose grazing at moderate intensity maintains Puccinellia–Carex grazing lawns on the salt marsh flats by retarding the development of dicotyledonous plants (Bazely and Jefferies 1986). Except for occasional grazing by small numbers of caribou, geese are the dominant herbivore in this system. In addition to lesser snow geese, Canada geese (Branta canadensis) and Ross’s geese (Chen rossii) feed in small numbers on intertidal vegetation. Eventually, the effects of regional isostatic uplift modify the physical environment and result in the replacement of these Puccinellia–Carex lawns by Calamagrostis deschampsiodes and Festuca rubra, which are not as heavily grazed (Jefferies et al. 1994). However, Puccinellia and other graminoids have failed to establish naturally in degraded areas of the supratidal marsh that have been exclosed for twenty-four years. Some natural revegetation can occur in moist intertidal sediments within five years in exclosed plots, but there is significant spatial and temporal variability in establishment success related to the moisture and salinity contents of the sediments (Handa and Jefferies 2002). In addition, tillers of P. phryganodes rarely reestablish in sediments where the redox potentials are negative, and impeded drainage within the

intertidal marsh is common, resulting in anaerobic conditions. However, annual halophytic plants (e.g., Salicornia borealis), which are not eaten by the geese, colonize exposed sediment, particularly where a veneer of soil organic matter is still present, and the dominance of these species is reflected in the seed bank composition of exposed sediment (Chang, Jefferies, and Carleton 2001). When these annuals die, the mat of vegetation is removed by the wind, and little input of organic matter to the sediments occurs.

OVERGRAZING IN OTHER POLAR SYSTEMS: WHEN ARE VEGETATION SHIFTS REVERSIBLE? Although heavy foraging by vertebrate herbivores leads to habitat deterioration in most other polar systems, an alternate stable state fails to establish if foraging ceases, in contrast to the degraded salt marshes in the Hudson Bay region (i.e., the vegetation states are reversible on a decadal time scale). Nevertheless, largescale changes in vegetation in northern latitudes have been recorded in response to increases in vertebrate herbivore populations during recent decades (table 5.1). These changes provide a continuum of responses of vegetation to herbivory, and they are similar to those shown by coastal marshes of the Hudson Bay Lowlands discussed earlier. The George River caribou herd (Rangifer tarandus), which occupies much of the UngavaLabrador peninsula in Canada, has grown from 5,000 in 1955 to 725,000 in the early 1990s, resulting in severe habitat degradation of their summer foraging grounds (Messier 1995). As with snow geese, the high carrying capacity of their wintering grounds is responsible for the high population densities in their summer foraging and calving areas (Messier et al. 1988; Crête and Huot 1993; Messier 1995; Manseau, Huot, and Crête. 1996). Under the dry conditions of summer, the lichens of this shrub tundra are vulnerable to caribou disturbance (Boudreau and Payette 2004a, 2004b), and the

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Hudson Bay Lowlands, Canada

Salt marsh

Agricultural nutrient subsidies

Grubbing, grazing

Bare soil

Erosion, hypersalinity, N loss

Stable degraded state

Vegetation

Herbivore population driver

Activity

Alternate state

Contributing edaphic factors

Anticipated recolonization in absence of herbivore

Geese

Location

Herbivore

Moderate

Decreased soil fertility

Maritime tundra

Decline in fecal inputs

Introduced predator

Grassland

Aleutian Islands, Alaska

Seabirds

Moderate

None

Bare soil

Grazing, trampling

High winter carrying capacity

Shrub tundra

Quebec-Labrador, Canada

Stable degraded state

Erosion

Bare soil

Grazing, trampling

Domesticated

Grassland

Yamal and Gydansky peninsulas, Russia

Caribou (Reindeer)

Moderate

Increased soil temperature, fecal inputs

Grassland

Grazing, trampling

Semidomesticated

Tundra

Fennoscandia

TABLE 5.1 Mechanisms of herbivore-induced soil degradation or vegetation shifts in polar systems

Slow (interaction with reindeer grazing)

None

Dead stems

Bud mining, defoliation

Condition/availability of host trees, warm winters

Birch forest

Fennoscandia and Karelia, Russia

Moths

severe trampling and grazing by large numbers of animals degrade superficial organic horizons, exposing mineral soil at heavily used sites (Boudreau and Payette 2004a, 2004b). The negative impact of caribou on plant production in the summer range has been attributed to the low resilience of the vegetation to regrow (Manseau et al. 1996), as opposed to the effects of adverse changes in soil conditions on vegetation. With reduced caribou activity in the 1990s (Boudreau et al. 2003), exposed ground was recolonized, and rates of lichen regrowth were inversely correlated with the intensity of grazing at that time. A more extreme example is shown by the high densities of domesticated reindeer that have led to overgrazing and soil erosion on the Yamal and Gydansky peninsulas in Russia, where the lichen tundras were replaced by plantless “badlands” characterized by wind and thermal erosion (Vilchek 1997). Analogous to the exposed sediment of intertidal flats of Hudson Bay, this is effectively an alternate stable state where the prospects for the long-term recovery of these grasslands are very poor. Vertebrate herbivore-induced changes in edaphic conditions also drive vegetation changes in Scandinavia, where semidomesticated reindeer are responsible for the rapid conversion of heathland to grassland. Heavy grazing and trampling reduce the depth of the moss layer, and the resultant increase in soil temperature facilitates grass establishment (Van der Wal, Van Lieshout, and Loonen 2001; Olofsson, Stark, and Oksanen 2004). The high relative growth rates of grasses are maintained because of fecal and urine inputs from the reindeer (Van der Wal and Brooker 2004). Although healthland plants can establish readily in the grassland soil, the system is slow to return to heathland in the absence of grazing, probably because of seed bank depletion and limitations to seed dispersal. In a related study on the Aleutian Islands, the introduction of Arctic foxes induced strong shifts in plant productivity and community structure, resulting in a trophic cascade (Croll et al. 2005). By preying on seabirds, the foxes reduced nutrient transport

from the ocean to the land, which affected soil fertility and led to dwarf shrub communities replacing the former grasslands that had depended on the nutrient input. Collectively, the results of these studies indicate that top-down processes promote vegetation change in polar systems. In extreme cases, there is the development of an alternate stable state, where the time scale of replacement of the vegetation is the order of decades, and the original vegetation may be replaced by a different assemblage of species because of changes in edaphic conditions. Alternatively, the effects of herbivory can lead to a change in vegetation state, but the changes are readily reversible.

DO TOP-DOWN EFFECTS POSE SIMILAR THREATS TO TEMPERATE SALT MARSHES? The results of earlier studies of North American salt marshes indicated that plant–herbivore interactions appeared to be of little importance in the structuring of plant communities (Odum and Smalley 1959; Teal 1962). However, in recent years sufficient evidence to the contrary has been presented that indicates a reevaluation of the original paradigm is overdue. Geese can affect adversely the vegetation of temperate marshes where birds stage or winter. For example, the entire population of the greater snow goose (Chen caerulescens atlantica) stages in spring and autumn in the tidal brackish marshes of the estuary of the St. Lawrence River dominated by Scirpus pungens. Formerly, staging was restricted to these estuarine marshes (Giroux and Bédard 1987; Reed 1989), and even today the rhizomes of this species are still an important component of the diet. As the population has grown, the birds have exploited additional food resources from farmland. Giroux and Bédard (1987) found a lower Scirpus rhizome biomass in heavily used marshes compared with corresponding values for lightly used marshes. However, in exclosed plots after two years, production reached levels similar to those in lightly used areas. In spite of the

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continuous population increase, Gauthier et al. (2005) did not observe significant changes in Scirpus aboveground biomass between 1983 and 1999 in marshes on the south shore of the estuary. In contrast, a 47 percent decline in Scirpus stem density has been recorded on the north shore in the vicinity of Cap Tourmente since 1971 (Reed 1989). A similar situation exists in the Fraser Delta in British Columbia, where lesser snow geese eat a substantial proportion of the total rhizome of each plant of Scirpus (Burton 1977). At both these sites, there is no evidence of long-term decline in the production of these sedges and increased exposure of sediments, although primary production may be reduced in the short term. On the wintering grounds of the greater snow geese in Delaware and adjacent states, there also are reports of top-down effects on vegetation, although again the effects appear to be transient (Smith and Odum 1981; Gauthier et al. 2005). Before the 1970s when the population was much smaller than at present, feeding in marshes of Spartina alterniflora was common. The geese foraged during the period when the marsh plants were dormant; hence, the effects of grazing and rhizome grubbing were not as severe as when active growth was occurring. Nevertheless, foraging by the geese led to a decrease in net belowground production and a change in species composition, but not to an alternate stable state. In some marshes on the eastern seaboard of North America, plant biomass and production are largely controlled by invertebrate grazers and their predators (Silliman and Zieman 2001; Silliman and Bertness 2002), and these grazing effects can act synergistically with drought stress (Silliman et al. 2005). Frequently, the changes in trophic interactions leading to top-down effects are indirectly or directly the result of anthropogenic effects at the landscape level. In addition, the fall in primary production may be the outcome of bottom-up effects triggered by the foraging activities of the herbivore, as in the case of grubbing by lesser snow geese. 96

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MANAGEMENT ACTIONS AND RECOMMENDATIONS The present hunting regulations in the United States and Canada are controlled by the U.S. Fish and Wildlife Service and the Canadian Wildlife Service, respectively, but individual states and provinces can modify the regulations within prescribed limits set by the agencies, with respect to both the length (up to 107 days) and timing of the hunting season (between September 1 and March 10), and the bag limit for each species (see Kruse, Sharp, and Gamble 2005). However, in an attempt to control the increasing numbers of lesser snow geese, a 1999 act of the U.S. Congress established a special wetlands conservation order that allows spring hunting of lesser snow geese until late April or very early May when the birds are migrating northward. During this open season, electronic calls are permissible, shotguns can be unplugged (more than three cartridges allowed), and there is no bag limit. A spring hunt also takes place in Manitoba, Quebec (greater snow goose), and Saskatchewan with similar regulations set by the Canadian Wildlife Service. In Nunavut and the Northwest Territories, Canada, hunting of geese and egg collections are encouraged at any time, as they provide a muchneeded food source for the indigenous population. In practice, however, little or no hunting occurs during the breeding and posthatch seasons. The long-term effects on the midcontinent lesser snow goose population of changes in the hunting regulations leading to the introduction of a spring hunt in 1999 are inconclusive at this stage. The changes include not just the introduction of the special spring hunt, but also modifications to bag limits in the regular hunting season that increase or eliminate daily bag limits. The objective of these changes in regulations is to reduce the annual growth rate of the goose population below the replacement rate of 1 and to achieve a reduction in population size of approximately 50 percent over ten years. The midwinter index of the midcontinent population indicates a fall in

numbers between 1999 and 2003, but recently numbers have stabilized for reasons that are unclear. A number of alternative management strategies designed to further contain the midcontinent population of lesser snow geese are under active discussion at this time, but no decision has been made to implement additional control measures until there has been a thorough review of recent changes and research findings. The measures include egg destruction, the use of baits laced with contraceptive agents, and the hiring of professional hunters. Each of these measures poses difficulties of implementation or risks to wildlife, and it is by no means certain that the measures will meet the desired objective of reducing the midcontinent population. Although detailed surveys of coastal vegetation have been made over a number of years in coastal marshes of Hudson Bay, it is important that these surveys are continued and expanded geographically to assess the ability of the vegetation to recover as goose numbers decline and to determine the success, or otherwise, of the decision to reduce the population. It is far too expensive to undertake restoration measures, particularly as goose numbers are still increasing. As all of these biotic interactions involve transfers across ecosystems, it is difficult to predict the final outcome of changes in management policies. “The devil is very much in the details” at this stage. At present, the top-down effects of the geese on these Arctic wetlands continue unabated (Jefferies et al. 2006).

CONCLUSION These examples illustrate that the biotic and abiotic processes driving the degradation of Arctic salt marshes, where lesser snow geese forage, exhibit many common features characteristic of other grazing systems in northern latitudes and elsewhere. Increases in herbivore densities are often driven directly or indirectly by anthropogenic influences ranging from changes in land use to the introduction of domesticated animals or the control of preda-

tors. Migratory connectivity frequently leads to high grazing pressure at the local scale in otherwise unconnected systems, and extreme climatic conditions often play a role in triggering episodic intense grazing events. In all these cases, changes to the systems are driven either by an increase in consumers brought about by reductions in the sizes of predator populations, or by subsidies to consumers, which do not trigger a corresponding increase in numbers of predators. Many of the processes involve cross-ecosystem interactions, as in the case of the movement of increased numbers of snow geese in the North American Arctic and reindeer in the Russian Arctic, which result in runaway consumption of primary production. Elsewhere, top-down processes in temperate and subtropical coastal marshes have gone unrecognized or have been largely ignored, but recent studies indicate that often they can be important in determining the pattern of vegetation change at the landscape level. Plant recolonization of these modified landscapes in the absence of further foraging appears to be determined to a large degree by the extent of soil degradation, although depleted soil seed banks and poor dispersal also contribute to the lack of establishment. Recovery from the effects of foraging is inherently slow in many northern grazing systems, and unrelated processes, such as isostatic uplift, may dictate both the rate and trajectory of change. Acknowledgments. Both authors gratefully acknowledge financial support from NSERC, the Arctic Goose Joint Venture, and the Department of Northern Affairs and Development of the Government of Canada. Logistic support was provided by the Hudson Bay Project, Parks Canada, Hudson Bay Helicopters, and the Churchill Northern Studies Centre. REFERENCES Abraham, K. F., and R. L. Jefferies. 1997. High goose populations: Causes, impacts and implications. Pages 7–72 in D. J. Batt (ed.), Arctic Ecosystems in Peril: Report of the Arctic Goose Habitat Working Group. Memphis, TN: Ducks.

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Odum, E. P., and A. E. Smalley. 1959. Comparison of population energy flow of a herbivorous and deposit invertebrate in a salt marsh ecosystem. Proceedings of the National Academy of Sciences of the USA 45: 617–622. Olofsson, J., S. Stark, and L. Oksanen. 2004. Reindeer influence on ecosystem processes in the tundra. Oikos 105: 386–396. Pastor, J., and R. J. Naiman. 1992. Selective foraging and ecosystem processes in boreal forests. American Naturalist 139: 690–705. Petraitis, P. S., and R. E. Latham. 1999. The importance of scale in testing the origins of alternative community states. Ecology 80: 429–442. Reed, A. 1989. Use of a freshwater tidal marsh in the St. Lawrence Estuary by greater snow geese. Pages 605–616 in R. R. Sharitz and J. W Gibbons (eds.), Freshwater Wetlands and Wildlife Conference. DOE Symposium Series No. 61. Oak Ridge, TN: U.S. Department of the Environment, Office of Scientific and Technical Information. Rietkerk, M., and J. van de Koppel. 1997. Alternative stable states and threshold effects in semi-arid grazing systems. Oikos 79: 69–76. Robertson, D. G., and R. D. Slack. 1995. Landscape change and its effects on the wintering range of a lesser snow goose Chen caerulescens caerulescens population: A review. Biological Conservation 71: 179–185. Ruess, R. W., D. S. Hik, and R. L. Jefferies. 1989. The role of lesser snow geese as nitrogen processors in a sub-arctic salt marsh. Oecologia 79: 23–29. Schadt, C. W., A. P. Martin, D. A. Lipso, and S. K. Schmidt. 2003. Seasonal dynamics of previously unknown fungal lineages in tundra soils. Science 301: 1359–1361. Schlesinger, W. H. 1990. Biological feedbacks in global desertification. Science 247: 1043–1048. Schmidt, S. K., and D. A. Lipson. 2004. Microbial growth under the snow: Implications for nutrient and allelochemical availability in temperate soils. Plant and Soil 259: 1–7. Schröder, A., L. Persson, and A. M. de Roos. 2005. Direct experimental evidence for alternative stable states: A review. Oikos 110: 3–19. Silliman, B. R., and M. D. Bertness. 2002. A trophic cascade regulates salt marsh primary production. Proceedings of the National Academy of Sciences of the USA 99: 10500–10505. Silliman, B. R., J. van de Koppel, M. D. Bertness, L. E. Stanton, and I. A. Mendelssohn. 2005. Drought, snails, and large-scale die-off of southern US salt marshes. Science 310: 1803–1806. Silliman, B. R., and J. C. Zieman. 2001. Top-down control of Spartina alterniflora production by

periwinkle grazing in a Virginia salt marsh. Ecology 82: 2830–2845. Sinclair, A. R. E., and J. M. Fryxell. 1985. The Sahel of Africa: Ecology of a disaster. Canadian Journal of Zoology 63: 987–994. Skinner, W. R., R. L. Jefferies, T. J. Carleton, R. F. Rockwell, and K. F. Abraham. 1998. Prediction of reproductive success and failure in lesser snow geese based on early season climatic variables. Global Change Biology 4: 3–16. Smith, T. J., III, and W. E. Odum. 1981. The effects of grazing by snow geese on coastal marshes. Ecology 62: 98–106. Srivastava, D. S., and R. L. Jefferies. 1995a. The effects of salinity on the leaf and shoot demography of two arctic forage species. Journal of Ecology 83: 421–430. ———. 1995b. Mosaics of vegetation and soil salinity: A consequence of goose foraging in an arctic salt marsh. Canadian Journal of Botany 73: 75–85. ———. 1996. A positive feedback: Herbivory, plant growth, salinity and the desertification of an arctic salt marsh. Journal of Ecology 84: 31–42. Stark, S., D. A. Wardle, R. Ohtonen, T. Helle, and G. W. Yeates. 2000. The effect of reindeer grazing on decomposition, mineralization and soil biota in a dry oligotrophic Scots pine forest. Oikos 90: 301–310. Stutzenbaker, C. D., and R. J. Buller. 1974. Goose depredation on ryegrass pastures along the Texas Gulf coast. Special Report, Federal Aid Project 106R, Texas Parks and Wildlife Department, Austin. Sutherland, J. P. 1990. Perturbations, resistance, and alternative views of the existence of multiple stable points in nature. American Naturalist 136: 270–275. Teal, J. M. 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43: 614–624. Van de Koppel, J., J. M. Rietkerk, and F. J. Weissing. 1997. Catastrophic vegetation shifts and soil degradation in terrestrial grazing systems. Trends in Ecology and Evolution 12: 352–356. Van der Wal, R., and R. W. Brooker. 2004. Mosses mediate grazer impacts on grass abundance in Arctic ecosystems. Functional Ecology 18: 77–86. Van der Wal, R., S. Van Lieshout, and M. Loonen. 2001. Herbivore impact on moss depth, soil temperature and Arctic plant growth. Polar Biology 24: 29–32. Vilchek, G. E. 1997. Arctic ecosystem stability and disturbance: A West-Siberian case history. Pages 179–190 in R. M. M. Crawford (ed.), Disturbance and Recovery in Arctic Lands: An Ecological Perspective. NATO ASI Series, vol. 25. Dordrecht: Kluwer Academic.

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6

Top-Down Control and Human Intensification of Consumer Pressure in Southern U.S. Salt Marshes Brian R. Silliman, Mark D. Bertness, and Mads S. Thomsen Salt marsh ecosystems in the western Atlantic are widely considered to be controlled exclusively by bottom-up forces (e.g., salinity, redox, nutrients). However, mounting experimental evidence challenges this established theory and shows that top-down control of marsh grass growth by consumers (i.e., plant-grazing snails and their predators) has been greatly underestimated and that human disturbance is currently triggering intensified consumer control. In the southeastern United States, unambiguous experimental evidence has revealed that marsh snail populations, likely released from predator regulation and stimulated to overgraze marsh grass by drought-induced soil stress, have formed massive grazer fronts and are contributing to large-scale salt marsh die-offs by denuding marsh substrate and leaving exposed mudflats in their wake. Unfortunately, such runaway consumption is likely to increase in the near-future concomitant with predicted increases in drought stress and human harvest of key marine predators (e.g., blue crabs) that facilitate marsh grass via a trophic cascade. Combined, these experimental results call for a complete realignment of the current marsh paradigm and for marsh scientists to incorporate experiments into future investigations of salt marsh productivity and die-off to properly account for consumer impacts. Reliance on field and lab correlation studies, which has led to dangerous theory dependency in salt marsh conservation and ecology, can no longer be the default scientific methods of choice. The lack of recognition of the importance of consumers in this system may have grave repercussions for coastal ecosystem management and conservation, since government agencies and nongovernmental organizations are managing marsh die-offs only considering bottom-up forces as causal agents.

TOP-DOWN AND BOTTOM-UP CONTROL OF PLANT COMMUNITY STRUCTURE Prevailing ecological theory has long held that climate ultimately controls the distribution and primary productivity of ecosystems. Of these

ecosystems, those dominated by plants (e.g., tundra, deciduous and coniferous forest, rain forest, grasslands, kelp forest, seagrass beds, mangroves, salt marshes) are typically green in appearance, and consumer control in these communities was considered insignificant 103

or subtle—for example, potentially affecting species composition, productivity, and some ecosystem properties, but not capable of regulating the long-term persistence and distribution of these systems. In 1960, Hairston, Smith, and Slobodkin proposed an opposing view that suggested the consumers play a much more important role in structuring plant-dominated ecosystems. They argued that the world is green because higher trophic levels control herbivore abundance. In essence, they suggested that grazers must not be food limited given the abundance of green material on Earth and that, therefore, what keeps herbivore populations and their negative impacts on plants in check is top-down control by predators. Thus, the world is kept green by a tritrophic interaction, where predators control grazers that would otherwise overgraze and denude substrate. In response, other scientists pointed out that what is green is not necessarily edible or of sufficient quality to allow increases in herbivore populations. This chemically mediated, bottom-up view purports that most plants have “won” the predator–prey arms race and are heavily defended and free from significant enemy attack. This debate is ongoing, but the dominant view remains that consumers can impact many aspects of plant ecology, but that they are not key drivers of plant productivity over entire ecosystems. In recent decades, however, examples of conspicuous grazer control of entire plant ecosystems have emerged. These studies show that runaway herbivory can replace more subtle community effects with obvious overgrazing, converting potentially green ecosystems to barrens. In several cases, foundation plant species (sensu Dayton 1972) are grazed and replaced with unvegetated substratum or small inconspicuous species. In terrestrial systems, both invertebrates (e.g., native or introduced beetles and moths) and vertebrates (e.g., introduced possums) can denude entire forest canopies (Hard, Holsten, and Werner 1983; McManus et al. 1992; Holsten, Werner, and DeVelice 1995; Norton 2000), insects can defoliate mangrove 104

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stands (Anderson and Lee 1995; Feller 2002), and ungulates and elephants can convert savannas to sandy deserts. Similar examples exist from marine systems, where urchins can convert kelp forests and seagrass beds to barren rock bed and sandflats, respectively (Camp, Cobb, and van Breedveld 1973; Ogden, Brown, and Selesky 1973; Estes and Palmisano 1974; Lawrence 1975; Estes et al. 1998; Rose et al. 1999). These examples of runaway consumption are considered undesirable from a conservation and management perspective, because overgrazed systems tend to have lowered biodiversity and productivity, and less economic and aesthetic human value. From an ecological perspective, these examples warn that whole-ecosystem regulation by consumers may be much more prevalent than currently recognized.

TROPHIC CASCADES: CONSUMER CONTROL TRICKLES UP THE LADDER Since the world is generally green, runaway grazer effects have been considered exceptions and viewed as relatively unimportant. Over the past fifty years, however, it has become clear that consumers can be important drivers of ecosystems even when they are overwhelmingly green. This has happened by focusing on the predators of grazers. When ecosystems are green, predators are often holding grazers in check, while when they are overgrazed, predator loss or removal is often responsible for elevated grazer densities and plant loss. This tritrophic interaction, where predators facilitate plants by controlling grazer populations, is known as a trophic cascade. Hairston et al. (1960) first hypothesized that the world is green because predators control grazers, and Carpenter and colleagues further developed the trophic cascade concept with experimental studies in lakes, demonstrating that fish can control zooplankton that, in turn, control phytoplankton (Carpenter, Kitchell, and Hodgson 1985; Carpenter and Kitchell 1988). Since these seminal studies, manipulative experiments have demonstrated trophic cascades in many other communities,

and rules of thumb have emerged predicting where and when trophic cascades occur. In general, trophic cascades tend to be more important in aquatic versus terrestrial systems, in simple versus complex food webs, in homogenous versus heterogeneous systems, in communities dominated by nonvascular plants (i.e., algae), and in systems where impalatable plants don’t replace those that have been overgrazed (Strong 1992; Pace 1999; Shurin et al. 2005). With the widespread occurrence of ecosystem shaping trophic cascades, ecologists and conservation biologists must now consider the relative importance of top-down effects in controlling plant ecosystems and must not end their scientific inquiry at the second trophic level, as they have in the past. Predators and their indirect effects need to be incorporated into ecological models and conservation strategies of plant communities. Importantly, ecologists differentiate between population- and community-level trophic cascades. In population-level cascades, predator removal leads to overgrazing and local extinction of a plant species. However, the overgrazed plant is then replaced with an impalatable plant species so that the ecosystem remains green and intact. In contrast, in a community-level cascade, predator removal leads to overgrazing of the entire plant community with the concomitant loss of associated ecosystem services. Both types of cascades occur in coastal salt marshes, with community-level cascades having the potential to destroy marshes and the services they provide (see chaps. 5 and 7, this volume). To identify management strategies needed to help preserve coastal plant communities, conservation practitioners and researchers need to use field experiments to test for the presence and impacts of both types of trophic cascades.

THE BOTTOM-UP-ONLY MARSH PARADIGM AND ITS REACH For nearly fifty years, ecologists have recognized and promoted salt marshes as the quintessential model ecosystem controlled by physical

forces, where primary production was controlled by bottom-up factors, such as soil nutrient concentrations, pH, salinity, redox, air temperature, sea level, and precipitation (Teal 1962; Odum and De La Cruz 1967; Adam 1990; Mendelssohn and Morris 2001). This paradigm grew from classic work by Eugene Odum, John Teal, and others on Sapelo Island, Georgia, in the 1950s and 1960s stressing the dominant role of physical factors in regulating ecosystem productivity and structure. The importance of consumers, while not rigorously tested with experiments that removed grazers, was largely disregarded, and the dogma that herbivores were unimportant became deeply entrenched in coastal wetland ecology. Since marshes provide crucial ecological and societal services, this paradigm also became the bedrock of coastal conservation. The Odum model of physically controlled ecosystems gained such wide acceptance that it was exported to other ecosystems dominated by lush vascular plant production, including mangrove forests, seagrass meadows, and temperate and tropical forests. This conceptual exportation to other systems, however, did not question its basic, untested assumption that consumers were irrelevant. Thus, theory dependency and demonstration, rather than falsifying science, led to the widespread application of the salt marsh bottom-up theory throughout coastal conservation and ecology. THE OVERTURNING OF THE BOTTOM-UP ONLY PARADIGM

Salt marshes are relatively homogeneous, lowdiversity ecosystems characterized by simple food webs. Despite the early work of Odum and colleagues suggesting that consumers were irrelevant to marsh plant community dynamics, these community and food web characteristics of salt marshes theoretically makes them ideal candidates for top-down control via trophic cascades (see earlier discussion; Strong 1992). Indeed, in clear contrast to the still-persisting and widely invoked bottom-up paradigm of salt marsh ecology, large-scale, top-down control of plant growth was demonstrated experimentally

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in arctic salt marshes by Robert Jefferies and coworkers over two decades ago (e.g., Bazely and Jefferies 1986; Jefferies 1988; Kotanen and Jefferies 1997; Jefferies and Rockwell 2002; Jefferies, Jano, and Abraham 2006). Jefferies studied these systems in the early 1970s, initially focusing on the positive effects of geese grazing on primary production through soil disturbance and nitrogen cycling. But by the 1980s, the snow geese that annually migrated to Hudson Bay switched from feeding in temperate zone wetlands to agricultural fields and golf courses, which were receiving nitrogen fertilizer subsides. Consequently, snow geese populations nearly tripled during the 1980s, leading to runaway consumption and the denuding of extensive areas of Arctic marshes (currently more than thirty-seven thousand hectares in southern Hudson Bay alone). This collapse was driven by birds grubbing roots and rhizomes that then led to low plant cover and hypersaline and anoxic soil conditions. This grazer-generated soil stress created a negative feedback loop where the remaining vegetation died, soil salinity increased even further, and plants that recruited into the newly denuded areas died rapidly from osmotic stress, preventing ecosystem recolonization. Essentially, at high densities, geese foraging turned off habitat-ameliorating, positive feedbacks that had historically allowed plants to establish and support arctic marsh ecosystems. Jefferies and colleagues attributed this ecosystem collapse to a trophic cascade, where geese populations increased due to human activities of declining hunting pressure and expansion of agricultural fields and golf courses that subsidized geese with plant nitrogen. Most recently, the marsh bottom-up paradigm has been challenged on Sapelo Island, Georgia, where salt marsh ecosystem ecology started. On this island and in marshes in Virginia, South Carolina, and Louisiana, Silliman and colleagues employed consumer removal and addition experiments to demonstrate that salt marsh primary production

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throughout the region is largely regulated and suppressed by the most common and widespread (from Maryland to Texas) grazer in the system, the marsh periwinkle, Littoraria irrorata (Silliman and Zieman 2001; Silliman and Bertness 2002; Silliman and Bortolus 2003; Silliman and Newell 2003; Silliman et al. 2005). Removal of these snails at commonly occurring field densities (fifty to five hundred snails per square meter) increased production of the dominant salt marsh plant in the southeastern United States, the smooth cordgrass Spartina alterniflora, by 30 to 80 percent (fig. 6.1). These results clearly challenged the bottom-up paradigm, but they were widely disregarded by established marsh ecologists who claimed results were largely generated from experimental artifacts of caging. These scientists also dismissed the work because in numerous past studies, Littoraria had been shown to eat only dead marshgrass and associated fungi (Marples 1966; Alexander 1979; Kemp, Newell, and Hopkinson 1990; Currin, Newell, and Paerl 1995). Unfortunately, as was the case in past studies that investigated the importance of marsh grazers (Teal 1962; Odum and De La Cruz 1967), studies of Littoraria feeding ecology did not use experiments. Instead, they relied solely on correlation studies (i.e., gut content and isotopic analyses) and concluded that snails preferred fungi and dead grass, did not graze live grass, and were the most important detritivores in the marsh system. However, when Silliman employed manipulative experiments to examine snail feeding in the field, he noticed periwinkles grazed not only on dead but also live marshgrass (Silliman and Zieman 2001). When grazing on live Spartina, Littoraria used its radula to create longitudinal wounds one to twenty centimeters long and one to three millimeters wide (see fig. 6.2; Silliman and Zieman 2001; Silliman and Bortolus 2003). Those grazergenerated wounds subsequently became infected with fungi (fig. 6.2), and snails then concentrated their grazing activity on those

(A)

(B)

FIGURE 6.1 Effects of periwinkle grazing on Spartina standing crop and canopy structure in the tall zone after eight months: (A) low-density plot and (B) high-density plot. After twenty months, cordgrass in all medium-density plots was reduced wholesale and the marsh substrate completely denuded (B). From Silliman and Bertness 2002.

wounds to consume the facilitated fungal crop (Silliman and Newell 2003). Thus, even though Silliman’s results concurred with past marsh researchers’, that snails preferred fungi for food, the data disagreed with how periwinkles procured fungus. Instead of waiting for grass to die and be colonized by fungi, snails grazed live grass and exhibited lowlevel, fungal-farming behavior that utilized live leaves as a substrate for fungal crop growth (Silliman and Newell 2003). Further experiments in the same marshes that crossed fungicide application with snail presence showed that

FIGURE 6.2 Periwinkle snails and the radulations (i.e., grazing scars) they generate.

the end result of this farming activity for the plant host, Spartina, was growth suppression or death due to grazer facilitation of fungal infection (Silliman and Newell 2003). Thus, even though snails did not consume much live plant tissue, they severely suppressed grass growth through facilitation of microbial invasion, giving these grazers the ability to exert strong, topdown control that is greatly disproportional to the amount of live grass consumed. Caging experiments further showed that at naturally occurring high densities (1,200 snails per square meter) the top-down effect of fungalfarming snails is even more dramatic, and Littoraria can completely denude marsh substrate and reduce standing crop by as much as 3,500 g dry wt C m⫺2 to 0 g dry wt C m⫺2 (Silliman and Bertness 2002). When snail presence was crossed with nutrient additions (a bottom-up force) and across naturally occurring nutrient gradients (from the short to the tall Spartina zone), snail impacts were greatly amplified. Increased bottom-up resources for plants and resultant higher nitrogen content in their tissue tripled grazing intensity and resulted in snails denuding the substrate in less than eight

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months (Silliman and Zieman 2001; Silliman and Bertness 2002). Combined, these caging studies demonstrate that if snail populations are left unchecked by regulatory forces, these systems have the potential to be overwhelmed and destroyed by runaway grazing. Simultaneous food web experiments examining the role of predators in controlling snail distribution and abundance across the marsh surface in Georgia salt marshes showed that marine predators such as blue crabs and terrapins largely control snail distribution (Silliman and Bertness 2002; fig. 6.3). Tethering and caging experiments demonstrated that predators completely exclude snails from the highly productive tall S. alterniflora zone, where they both recruit and grow better, and strongly suppress their densities in the short S. alterniflora zone. These results, combined with caging experiments showing snails, at naturally occurring densities, can completely kill a tall Spartina zone in eight months, reveal that the growth and success of southern U.S. marshes is controlled by a trophic cascade, where crabs, fish, and terrapins suppress densities of potent plant-grazing snails that would otherwise increase in density and destroy salt marsh plants (fig. 6.3). These experimental findings clearly overturn the long-held paradigm that grazers and their predators are unimportant players in salt marsh ecosystems. Thus, the new marsh paradigm must now reflect the fact that both bottom-up and top-down forces are important in controlling salt marsh structure and function. In addition, this paradigm realignment compels marsh scientists and conservation biologists to use field experiments to test for the relative importance of top-down impacts when investigating physical controls on marsh grass growth and success. Given that blue crabs have declined dramatically in recent decades in many southeastern estuaries (Jordan 1998 and references therein), partly due to overharvesting and partly to disease and drought, it is possible that the newly

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− + −

FIGURE 6.3 The marsh trophic cascade. From Silliman and Bertness 2002.

revealed marsh trophic cascade has been partially unleashed, thereby increasing top-down control by snails and the likelihood of runaway consumption events. Because the new interaction network of controls on cordgrass growth now includes top-down forces (fig. 6.3) in the form of snails and their blue crab predators, humans must also be included as apex predators. Humans, by suppressing and stressing blue crab populations via fishery take, could be indirectly intensifying top-down control on Spartina grass by releasing fungal-farming snails from predator regulation. As in the case

of Jefferies’s research in subarctic marshes, humans thus have the potential to unleash a powerful and devastating trophic cascade in intertidal grasslands.

MASSIVE DIE-OFF OF SOUTHEASTERN SALT MARSHES: COULD TOP-DOWN FORCES BE INVOLVED? Over the past decade, massive marsh die-offs have occurred on the southeastern and Gulf coasts of the United States. These die offs have been extensive, impacting over 250,000 hectares, and have received considerable media and scientific attention in South Carolina, North Carolina, Georgia, Florida, Texas, and Louisiana (McKee, Mendelssohn, and Materne 2004). Until recently, however, these die-offs had been attributed solely to harsh physical conditions killing marsh plants, an interpretation consistent with the Odum model and interannual correlations between drought stress and the occurrence of die-offs. Experimental work over the past five years, however, has shown that increasing consumer control of marsh plants is an important contributing factor driving these die-offs. As was the case in past studies challenging the bottom-up marsh paradigm, these experimental results are in direct contrast to those found by marsh ecologists employing correlation studies. Observational studies have concluded that top-down effects are not playing a direct role in the recent dieback events (Carlson et al. 2001; McKee et al. 2004; L. Blum, personal communication). Upon examination of die-back plants at particular Louisiana sites, McKee et al. (2004) reported no visible evidence of grazing on live S. alterniflora tissue. At other sites where snail grazing was observed, grazing activity appeared to occur post–plant mortality (McKee et al. 2004). These assertions, however, were based on observations of plant condition and not on experimental study or data gathering on snailgrazing intensity (McKee et al. 2004). When

experiments were used at many of these same sites in Louisiana to test for snail-grazing effects, extensive snail fronts (twenty to one hundred meters long, one to two meters wide, with five hundred to two thousand snails per square meter) were discovered on the edges of dieback areas (fig. 6.4). After six to twelve months of exclusion, marsh plants in control plots exposed to snail grazing were completely killed and exposed mudflats remained, while those in caged areas protected from snails were robust and growing healthily (Silliman et al. 2005). Importantly, reconstruction of snail densities prior to Louisiana die-off events clearly showed that snails could not have initiated these die-off events. Instead, Silliman et al. (2005) concurred with McKee et al. (2004) that climate-induced stress(es) (e.g., soil acidity or salinity) likely caused initial grass mortality and/or extreme sublethal stress. This physical stress then likely intensified snail grazing, as stressed plants are more susceptible to snail grazing (Silliman et al. 2005), resulting in increased grazer impacts and potential to kill off marsh grasses. These experimental results demonstrate that grazers played a significant contributory role in Louisiana marsh die-off events (i.e., they increased original die-off area by at least 15 percent) and challenge the observational conclusion of McKee et al. (2004) that grazers were unimportant in Louisiana die-off. Carlson et al. (2001) also observed heavy grazing on dead S. alterniflora shoots at dieback sites along the Florida panhandle. No data were taken on this observation or whether snails were actively grazing remaining live marsh grass. The authors concluded that grazing was a secondary event and not the primary cause of that dieback event. Again, in the absence of direct experimental evidence or even data collection, it is scientifically difficult, if not impossible, to accurately assess the relative effects of grazing or its role in plant mortality. Manipulative studies should therefore be conducted wherever possible and will be necessary

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FIGURE 6.4 Snail front exclosures (wrapped in hardware cloth) installed on die-off border at one site in Louisiana.

before the relative role of consumers (top-down forces) and bottom-up forces in driving marsh dieback can be fully elucidated. Experimental and survey evidence from several other dieback sites in Georgia and South Carolina also support the finding that snail grazing played a key but overlooked, contributory role in dieback of southern salt marshes (fig. 6.5). Silliman et al. (2005) reported elevated

density of the marsh periwinkles on the edges of dieback areas (Littorina irrorata) at eleven of twelve sites surveyed across 1,200 kilometers of shoreline in Louisiana, Georgia, and South Carolina. The highest densities of snails were located at the boundary between dieback areas and healthy marsh where snail density ranged from ~400 to 2,200 individuals per square meter (fig. 6.6). Exclusion experiments in Georgia, as

FIGURE 6.5 Effect of snail exclusion cage on Spartina alterniflora biomass on die-off border at the Light House marsh on Sapelo Island, Georgia. Plants protected from snails in cages are healthy and robust, while those exposed to snails in uncaged control areas, and those in uncaged, wrack removal areas (with many stakes) where completely killed and exposed mudflat was left in their wake.

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FIGURE 6.6 Representative extreme snail densities found in snail fronts in Georgia.

in Louisiana (discussed earlier), resulted in a more than 400 percent increase in S. alterniflora biomass over three months; after six months, plants in control areas still exposed to snail grazing were completely grazed down, and exposed mudflats remained. These snail fronts maintained their integrity for six to twelve months and continued moving through healthy areas of S. alterniflora, resulting in expansion of original dieback areas by as much as 185 percent in what was described as a runaway grazing event (Silliman et al. 2005). As mentioned, coincident with initial marsh die-off events in the southern United States were record droughts throughout the Gulf and southeastern Atlantic coasts from 1999 to 2001 (McKee et al. 2004; Silliman et al. 2005). Researchers investigating the dieback throughout the region suspected that the drought had played an important role as an antecedent condition (trigger event) for marsh dieback. One set of proposed mechanisms operating at the Louisiana dieback sites (episodic acidification and elevated metal toxicity) was investigated in the lab experimentally (McKee et al. 2004). To test for the potential interactive effects of grazing and sublethal, drought-induced physical stress (in this case, elevated soil salinities, which were observed in both Louisiana and Georgia), Silliman et al. (2005) experimentally elevated salts in the presence (controls) and absence (exclusion cages) of snails in a healthy Georgia marsh. Experimental results revealed that exposure to sublethal salinities, combined

with grazing, resulted in nearly doubling the reduction in S. alterniflora biomass, as compared to single-stressor treatments, and caused severe plant mortality and small-scale, localized marsh die-off. Combined, the outcome of this stressenhanced field experiment, grazer exclusion experiments at the edges of die-off areas in both Louisiana and Georgia, and model analyses of snail front movement and formation (Siliman et al. 2005), suggests that drought-induced soil stress and grazers acted synergistically and, to varying degrees, to cause initial plant death (if snails were not present at sites, then there could not be a synergism, and just drought-induced effects are implicated). Following these localized disturbances, if snails were present, snail fronts formed on die-off edges and subsequently propagated through healthy marsh, leading to cascading vegetation loss. Whether or not declines in densities of one of the snails’ major predators, the blue crab, Callinectes sapidus, over the past ten years in southeast estuaries (Silliman et al. 2005) contributed indirectly to marsh die-off by leading to elevated snail salinities and thus greater potential for overgrazing, remains to be investigated. Importantly, though, the potential for the marsh trophic cascade to act as a significant contributing factor to marsh health deterioration can no longer be ignored or dismissed. The recent, catastrophic die-off of salt marshes that seemingly fit the Odum physical control model on the southeastern and Gulf coasts of North America is increasingly being controlled by consumers, whose influence is likely greatly exaggerated and intensified due to human-generated diffuse disturbance (i.e., intensification of drought events via climate change and overharvesting of predators—terrapins and blue crabs—of potent snail grazers). This potentially catastrophic development—one that is completely opposite of the current marsh paradigm and of what is currently being promoted in conservation circles as the detrimental effects of human impacts on marine food webs (i.e., humans dampen top-down control)—may be an early sign of global shifts in the processes controlling salt marshes and coastal ecosystem

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services. The indirect role of humans in generating and unleashing intensified top-down control in salt marshes that can reduce biomass of plant foundation species wholesale and leave exposed mudflats in its wake can clearly no longer be ignored or dismissed.

FUTURE RESEARCH AND RECOMMENDATIONS TO MARSH MANAGERS Future experimental research investigating the importance of top-down control in southeastern U.S. marsh systems should focus on (1) the species-specific role of marsh predators in regulating historical and present-day grazer densities and distributions, (2) the relative role of recruitment and predation in controlling grazer populations across marsh landscapes, (3) how variation in physical soil stressors and climate change impact grazer distribution and grazing intensity, and (4) the relative role of consumers in regulating marsh grass recovery from die-off events. Marsh ecologists and managers in systems throughout the world should also strive to use experiments to test for the presence of powerful, but cryptic, top-down impacts in their local systems (e.g., Chile, South Africa, China). To mitigate top-down impacts on marshes, managers should look to maintain healthy populations of predators that suppress densities of commonly occurring marsh grazers through regulation and marine reserves. Management must also begin to monitor both marsh grazer (e.g., snails in the Southeast and sesarmid crabs and insects in the Northeast; see chap. 8, this volume) and predator densities along with soil salinities, redox, and acidity levels in healthy marshes, so that the magnitude of marsh grass stressors can be measured before, during, and after dieback events and the relative importance of top-down and bottom-up forces properly assessed. In addition to reexamining current regulations in tidal marshes, managers should consider actions such as freshwater releases on managed rivers to coincide with drought and

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tighter regulation of populations of marsh predators (e.g., blue crabs). Acknowledgments. We thank two anonymous reviewers for their comments. We acknowledge funding from Georgia Sea Grant, the National Science Foundation, The Nature Conservancy, the National Atmospheric and Oceanic Association National Estuarine Research Reserve System, the Georgia Coastal Ecosystems-LTER, the University of Georgia Marine Institute, and an EPA-STAR and David H. Smith Conservation Research Fellowship to B. R. Silliman. This work could not have been completed without the help of Tracy Buck, Sarah Lee, Andrew Altieri, Irv Mendelssohn, Lee Stanton, and Jane Garbisch. REFERENCES Adam, P. 1990. Saltmarsh Ecology. Cambridge University Press, Cambridge Alexander, S. K. 1979. Diet of the periwinkle Littorina irrorata in a Louisiana salt marsh. Gulf Research Reports 93: 295. Anderson, C., and S. Y. Lee. 1995. Defoliation of the mangrove Avicennia marina in Hong Kong: Cause and consequences. Biotropica 27: 218–226. Bazely, D. R. 1987. Snow buntings feeding on leaves of salt-marsh grass during spring migration. The Condor 89: 190–192. Camp, D. K., S. P. Cobb, and J. F. van Breedveld. 1973. Overgrazing of seagrasses by a regular urchin, Lytechinus variegatus. BioScience 23: 37–38. Carlson, P. R., Jr., L. A. Yarbro, F. X. Courtney, T. Leary, H. Arnold, D. Leslie, J. Hughes, and N. Craft. 2001. Panhandle salt marsh mortality: A prelude to Louisiana brown marsh? In R. E. Stewart Jr., C. E. Proffitt, and T. M. Charron (eds.), Abstracts from “Coastal Marsh Dieback in the Northern Gulf of Mexico: Extent, Causes, Consequences, and Remedies.” Information and Technology Report, USGS/BRD/ITR—2001– 0003. Washington, DC: U.S. Geological Survey, Biological Resources Division. Carpenter, S. R., and J. F. Kitchell. 1988. Consumer control of lake productivity. BioScience 38: 764–769. Carpenter, S. R., J. F. Kitchell, and J. R. Hodgson. 1985. Cascading trophic interactions and lake productivity. BioScience 35: 634–639. Currin, C. A., S. Y. Newell, and H. W. Paerl. 1995. The role of standing dead Spartina alterniflora and

benthic microalgae in salt marsh food webs: Considerations based on multiple stable isotope analysis. Marine Ecology Progress Series 121: 99–116. Dayton, P. K. 1972. Towards an understanding of community resilience and the potential effects of enrichment to the benthos of McMurdo Sound, Antarctica. Proceedings of the Colloquium on Conservation Problems in Antarctica: 81–96. Estes, J. A., and J. F. Palmisano. 1974. Sea otters: Their role in structuring nearshore communities. Science 185: 1058–1060. Estes, J. A., M. T. Tinker, T. M. Williams, and D. F. Doak. 1998. Killer whale predation on sea otters linking oceanic ecosystems. Science 282: 473–476. Feller, I. C. 2002. The role of herbivory by wood-boring insects in mangrove ecosystems in Belize. Oikos 97: 167–176. Hairston, N. G., F. E. Smith, and L. S. Slobodkin. 1960. Community structure, population control, and competition. American Naturalist 94: 421–425. Hard, J. S., E. H. Holsten, and R. A. Werner. 1983. Susceptibility of white spruce to attack by spruce beetles during the early years of an outbreak in Alaska. Canadian Journal of Forest Research 13: 678–684. Holsten, E. H., R. A. Werner, and R. L. DeVelice. 1995. Effects of a spruce beetle (Coleoptera: Scolytidae) outbreak and fire on lutz spruce in Alaska. Environmental Entomology 88: 1539–1547. Jefferies, R. L. 1988. Pattern and process in arctic coastal vegetation in response to foraging by lesser snow geese. Pages 341–369 in L. D. Gottlieb and S. K. Jain (eds.), Plant Evolutionary Biology. London: Chapman & Hall. Jefferies, R. L., A. P. Jano, and K. F. Abraham. 2006. A biotic agent promotes large-scale catastrophic change in the coastal marshes of Hudson Bay. Journal of Ecology 94: 234–242. Jefferies, R. L., and R. Rockwell. 2002. Foraging geese, vegetation loss and soil degradation in an Arctic salt marsh. Applied Vegetation Science 5: 7–16. Jordan, S. J. 1998. National, international, and regional perspectives on blue crab fisheries. Journal of Shellfish Research 17: 367–587. Kemp, P. F., S. Y. Newell, and C. S. Hopkinson. 1990. Importance of grazing on the salt-marsh grass Spartina alterniflora to nitrogen turnover in a macrofaunal consumer, Littorina irrorata, and to decomposition of standing-dead Spartina. Marine Biology 104: 311–319. Kotanen, P. M., and R. L. Jefferies. 1997. Long-term destruction of sub-arctic wetland vegetation by lesser snow geese. Ecoscience 4: 179–182.

Lawrence, J. M. 1975. On the relationships between marine plants and sea urchins. Oceanography and Marine Biology Annual Review 13: 213–286. Marples, T. G. 1966. A radionuclide tracer study of arthropod food chains in a Spartina salt marsh ecosystem. Ecology 47: 270–277. McKee, K. L., I. A. Mendelssohn, and M. D. Materne. 2004. Acute salt marsh dieback in the Mississippi River deltaic plain: A drought-induced phenomenon? Global Ecology and Biogeography 13: 65–73. McManus, M., N. Schneeberger, R. Reardon, and G. Mason. 1992. Gypsy moth. Forest Insect and Disease Leaflet. Washington, DC: U.S. Department of Agriculture, Forest Service. Mendelssohn, I. A., and J. T. Morris. 2001. Eco-physiological constraints on the primary productivity of Spartina alterniflora. In M. P. Weinstein and D. A. Kreeger (eds.), Concepts and Controversies in Tidal Marsh Ecology. Dordrecht: Kluwer Academic. Norton, D. A. 2000. Benefits of possum control for indigenous vegetation. Pages 232–240 in T. Montague (ed.), Possums in New Zealand: The Biology, Impact and Management of an Introduced Marsupial. Lincoln: Mannaki Whenua. Odum, E. P., and A. A. De La Cruz. 1967. Particulate organic detritus in a Georgia salt-marsh estuarine system. Pages 383–385 in G. H. Lauff (ed.), Estuaries. AAAS Publication 83. Washington, DC: American Academy of Arts and Sciences. Ogden, J. C., R. A. Brown, and N. Salesky. 1973. Grazing by the echinoid Diadema antillarum Philippi: Formation of halos around West Indian patch reefs. Science 182: 715–717. Pace, M. L., J. J. Cole, S. R. Carpenter, and J. F. Kitchell. 2004. Trophic cascades revealed in diverse ecosystems. Trends in Ecology and Evolution 14: 483–488. Rose, C. D., W. C. Sharp, W. J. Kenworthy, J. H. Hunt, W. G. Lyons, E. J. Prager, J. F. Valentine, M. O. Hall, P. E. Whitfield, and J. W. Fourqurean. 1999. Overgrazing of a large seagrass bed by the sea urchin Lytechinus variegatus in outer Florida bay. Marine Ecology Progress Series 190: 211–222. Shurin, J. B., E. T. Borer, E. W. Seabloom, K. Anderson, C. A. Blanchette, B. Broitman, S. D. Cooper, and B. S. Halpern. 2005. A cross-ecosystem comparison of the strength of trophic cascades. Ecology Letters 5: 785–791. Silliman, B. R., and M. D. Bertness. 2002. A trophic cascade regulates salt marsh primary production. Proceedings of the National Academy of Sciences of the USA 99: 10500–10505. Silliman, B. R., and A. Bortolus. 2003. Underestimation of Spartina alterniflora production

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in Western Atlantic salt marshes. Oikos 143: 549–555. Silliman, B. R., and S. Y. Newell. 2003. Fungal farming in a snail. Proceedings of the National Academy of Sciences of the USA 100: 15643–15648. Silliman, B. R., J. van de Koppel, M. D. Bertness, L. Stanton, and I. Mendelsohn. 2005. Drought, snails, and large-scale die-off of southern U.S. salt marshes. Science 310: 1803–1806.

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Silliman, B. R., and J. C. Zieman. 2001. Top-down control of Spartina alterniflora production by periwinkle grazing in a Virginia salt marsh. Ecology 82: 2830–2845. Strong, D. R. 1992. Are trophic cascades all wet? Differentiation and donor-control in speciose ecosystems. Ecology 73: 747–754. Teal, J. M. 1962. Energy flow in the salt marsh ecosystem of Georgia salt marshes. Ecology 43: 614–624.

7

Alligator Hunters, Pelt Traders, and Runaway Consumption of Gulf Coast Marshes a trophic cascade perspective on coastal wetland losses Paul A. Keddy, Laura Gough, J. Andy Nyman, Tiffany McFalls, Jacoby Carter, and Jack Siegrist The rate of loss of Gulf Coast marshes in general, and the Louisiana coastline in particular, is now a national issue, particularly following the 2005 hurricanes in the region. We suggest that current management paradigms for marsh restoration may focus too exclusively on plants and sediment, with a bottom-up view of coastal wetlands. Top-down processes also merit consideration and may expand the array of potential tools for coastal management and restoration. Here we propose an alligator trophic cascade hypothesis incorporating a top-down approach: that alligator hunting, by reducing the density and mean size of alligators, removes a natural control on the primary herbivores in wetlands, enabling the runaway consumption of coastal marshes. We present current evidence to support this hypothesis. Mammalian grazing can directly remove plant biomass and make plants less tolerant to flooding and salinity, therefore increasing erosion of sediments. Both muskrats and nutria have been implicated in this process, with the larger, nonnative nutria of greater current concern. Annual aerial surveys beginning in 1998 indicated that 321 to 415 square kilometers of Louisiana’s 14,164 square kilometers of coastal wetlands were severely damaged by nutria. Adult alligators eat muskrats and nutria, but the role of alligators as potential controllers of mammal populations, and thus as controllers of marsh damage, has received minimal consideration. Our hypothesis cannot be tested with existing data because almost no numbers exist for nutria populations, and data on marsh integrity, plants, nutria, and alligator densities are not being collected in a systematic way across multiple sites. However, we explore these relationships with a modeling exercise and propose several different ways to test these relationships empirically. If the hypothesis is supported, reducing the alligator harvest or closely controlling the size of the animals being harvested may prove a valuable management tool in conserving coastal wetlands. Adopting multiple working hypotheses, including a top-down approach, may be crucial to adequately managing and restoring coastal areas.

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Should I say, that the river (in this place) from shore to shore, and perhaps near half a mile above and below me, appeared to be one solid bank of fish, of various kinds, pushing through this narrow pass of St. Juans into the little lake, on their return down the river, and that the alligators were in such incredible numbers, and so close together from shore to shore, that it would have been easy to have walked across on their heads, had the animals been harmless. William Bartram (1791, 123)

The rapid loss of wetlands along the Gulf Coast of North America poses a serious threat to wildlife populations, human infrastructure, and local and regional economies (Boesch et al. 1994). Coastal wetland loss in Louisiana averaged one hectare a day from 1978 through 1990 (Barras, Bourgeois, and Handley 1994). Estimates of one-time conversion of marsh to open water caused by Hurricane Katrina in 2005 approach seventy-five square kilometers in one area of coastal Louisiana (20 to 26 percent of the study area; U.S. Geological Survey [USGS] 2005). Some marsh loss is natural in the Mississippi Deltaic Plain, where the river has changed drainage patterns over millennia, instigating natural cycles of construction and degradation (Teller and Thorleifson 1983; Boyd and Penland 1988). Humans have accelerated this loss in several ways, including building levees on the Mississippi River that prevent spring floods from delivering fresh water and sediments to the marshes, constructing canals and spoil banks that alter salinity and hydrology (Boesch et al. 1994; Turner 1997), controlling the amount of water (and associated sediment) in the Mississippi River that enters the Atchafalaya River basin, and introducing an exotic herbivore, nutria, Myocastor coypus (Keddy et al. 2007; figs. 7.1 and 7.2). These human alterations have reduced rates of vertical accretion that would otherwise offset the effects of global sea-level rise and local subsidence (Day and Templet 1989; Boesch et al. 1994). Nutria, and to a lesser degree muskrats (Ondatra zibethicus), are implicated in accelerat116

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FIGURE 7.1 Coastal marshes, such as these along the Mississippi River, have been lost and fragmented by the dredging of canals and creation of spoil banks. About 12 percent of Louisiana’s coastal wetland loss is attributable to this cause, while an additional 32 percent is due to factors such as hydrological changes and sedimentation reduction resulting from dams and levees.

FIGURE 7.2 A modern aerial view of the coastal marshes in the Manchac area, northwest of New Orleans, showing the scars left by pullboat logging of the former cypress swamp. Channels dredged for the boats and ditches gouged out by the logs have altered the marsh hydrology, facilitating saltwater intrusion and organic decomposition.

ing the loss of coastal wetlands in Louisiana (Coastal Wetlands Planning, Protection, and Restoration Act Task Force [CWPPRA] 2008) (and in Chesapeake Bay; Patuxent Wildlife Research Center [PWRC] 1999). Nutria were introduced into Louisiana in the 1940s and soon became abundant and widespread (Lowery 1974; Bernard 2002). Marsh damage by nutria accelerated after the mid-1980s when the fur industry declined and trapping efforts similarly lessened. Recently, Louisiana has attempted to increase nutria harvest to reduce marsh

damage. The state promotes human consumption of nutria in the United States and elsewhere (particularly in China) to try to dispel the local image of nutria as “swamp rats,” unfit to eat. Beginning in 2002, Louisiana implemented an incentive payment program of an additional $4 per nutria (increased to $5 in 2006) for registered trappers (since the 1990s a pelt has normally been worth $1) for up to four hundred thousand animals a year for five years (CWPPRA 2008; Louisiana Department of Wildlife and Fisheries [LDWF] 2008). The cost of this program was initially estimated at $69 million statewide (CWPPRA 2008). The extent of marsh area damaged has consistently declined since the program was introduced, with the lowest levels detected in 2008 (LDWF 2008). These results, along with documented recovery of marsh in areas where many animals have been removed, suggest the control efforts are working. This success may be attributable to harvest goals being set to levels similar to the late 1970s when pelt prices were higher and to procedures that focus on lands with the most intense nutria damage. Governments in other areas to which nutria have been introduced, such as Maryland and England, chose to completely eradicate their nutria populations through systematic hunting and trapping programs subsidized by the state and federal governments (Carter and Leonard 2002). In the Chesapeake Bay region of Maryland, 53 percent of the marsh remaining at Blackwater National Wildlife Refuge was damaged by nutria (PWRC 1999), and an $8.2 million study of the effects of trapping on nutria density was initiated. In late 2004, the refuge was declared “nutria-free” in the popular press after the state of Maryland paid trappers to kill every animal they encountered (Fahrenthold 2004). Nonetheless, nutria still persist in low numbers on the refuge and adjacent lands (D. Birch, Blackwater National Wildlife Refuge, personal communication). In all, approximately 8,300 nutria were killed by 2004, although the original population on the refuge was estimated to be closer to 50,000 animals. For perspective,

in just one of the heavily nutria-populated parishes in Louisiana, more than one hundred thousand nutria were killed in one season (CWPPRA 2008). Louisiana officials have not taken a similar eradication approach, partly because the more extensive marshes in the state make eradication logistically very difficult, and they instead encourage trapping as part of the state’s fur industry (Louisiana Fur and Alligator Advisory Council [LFAAC] 2004). Control of nutria through harvesting by humans in the absence of a sustainable commercial market, even if effective, will require continued government inputs, not only through incentive payments but also through the costs of management itself, such as the staff required to monitor nutria, regulate hunting activity, and pay incentives. Natural predation may be a more costeffective means to control nutria and ultimately protect marshes from damage if a strong trophic cascade exists among marsh vegetation, herbivores, and predators (figs. 7.3 and 7.4). The term trophic cascade refers to a situation where a predator controls the abundance of herbivores, thereby indirectly controlling the biomass and species composition of plant communities. When such predators are reduced in numbers, the herbivores may increase in population size, become limited by food availability rather than predation, and cause significant damage to vegetation. Recent reviews (e.g., Schmitz, Krivan, and Ovadia 2004; Borer et al. 2005) suggest that trophic cascades occur in a variety of ecosystems—aquatic, wetland, and terrestrial. They may be more important in food webs that resemble linear food chains with only one or two species of consumers in each trophic level (Borer et al. 2005), similar to what is found in Louisiana marshes. In southern U.S. salt marshes, blue crabs are important predators on snails that consume marsh vegetation, and crabs may thereby protect salt marshes from overgrazing (Silliman and Bertness 2002). In northern marshes in North America, exploding geese populations have damaged nearly twothirds of the approximately fifty-five thousand

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(A)

(B)

FIGURE 7.3 (A) The loss of coastal wetlands, such as this expanse of oligohaline coastal marsh at the west end of Lake Pontchartrain, is a serious issue in coastal Louisiana. Grazing by nutria (B) accelerates the rates of loss. We explore evidence that predation by alligators (C) may protect marshes by consuming nutria.

(C)

hectares of salt marsh along the coast of Hudson and James bays (Jefferies and Rockwell 2002; Abraham and Keddy 2005). To our knowledge, the potential of alligators to create a trophic cascade in the Gulf Coast region and thus protect marshes from herbivory has not been examined. Previous explanations of the ecological significance of alligators in wetlands have largely addressed their role in digging alligator holes

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that increase the variation in plant communities and provide deep-water refuges for many wetland species in the Florida Everglades (Loveless 1959; Craighead 1968; Gunderson and Loftus 1993; Palmer and Mazzotti 2004). Dundee and Rossman (1989) observed similar behavior in Louisiana, where holes one to three meters in diameter and one to two meters deep were connected to underground dens used as retreats during the winter.

researchers and managers the possibility that these relationships exist, and to advocate for properly designed field experiments to test the hypothesis.

EFFECTS OF NUTRIA AND MUSKRATS ON MARSH VEGETATION AND WETLAND LOSS

FIGURE 7.4 Alligators may have a positive indirect effect on marsh vegetation by their negative direct effect on nutria and muskrats. There is also a possible link to the blue crab–periwinkle–vegetation food web, depending on salinity.

Reliable data on the effects of alligator predation on herbivore abundance are not available. Here we review some of the evidence consistent with the alligator trophic cascade hypothesis. We propose that herbivore damage to Louisiana coastal wetlands results, at least in part, from the release of predation by alligators on nutria; we present the results of a model that supports the hypothesis; and we explore what is needed to empirically test this hypothesis. Our review is limited to the available data, including evidence from the effects of nutria and muskrats on marsh vegetation, evidence from the diet of alligators, and largely anecdotal evidence regarding the relative abundance of alligators and herbivores in Louisiana wetlands over the past one hundred years. Our goal is to suggest to wetland

Annual aerial surveys beginning in 1998 provide a conservative estimate that 321 to 415 square kilometers of Louisiana’s 14,164 square kilometers of coastal wetlands were severely damaged by nutria (LDWF 2008). This damage occurred almost exclusively in the Mississippi Deltaic Plain, rather than in the Chenier Plain (unpublished map, LDWF 2008; CWPPRA 2008). Marshes in the Mississippi Deltaic Plain probably are more sensitive to nutria damage because submergence (i.e., the combination of local subsidence and global sea-level rise) exceeds 1 centimeter a year in the Mississippi Deltaic Plain but averages only 0.57 centimeter a year in the Chenier Plain (Penland and Ramsey 1990), and because nutria increase the sensitivity of vegetation to flooding and salinity stress (Gough and Grace 1998a; Grace and Ford 1996). Also, nutria may prefer the vegetation and habitat found in the Mississippi Delta; fewer nutria are harvested in the western coastal areas, likely reflecting lower abundance. Nutria and muskrats affect marsh plants directly by reducing the biomass of vegetation (Evers et al. 1998; Fuller et al. 1985), sometimes creating “eat-outs” (fig. 7.5), areas of marsh denuded of vegetation (Lynch, O’Neil, and Lay 1947). Removal of plant biomass increases the sensitivity of marsh soils to erosion because of the loss of living roots that trap and hold sediment (McGinnis 1997). Indirectly, grazing may also increase the sensitivity of plant species to flooding or salinity stress (Gough and Grace 1998a, 1998b; Grace and Ford 1996) and reduce organic matter necessary for vertical accretion (McCaffrey and Thomson 1980; Bricker-Urso et al. 1989; Gosselink, Hatton,

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and Hopkinson 1984; Nyman et al. 1993; Callaway, DeLaune, and Patrick 1997). Ford and Grace (1998a) observed that soil elevation increased less in nutria-grazed than in ungrazed plots. Models for Louisiana wetlands (e.g., Reyes et al. 2000) include a term for the rate by which plants contribute to accretion. That model has a term for herbivory partly because, if such grazing is omitted, the model shows that rates of accretion exceed those observed in nature (G. P. Kemp, personal communication). Nutria grazing may also be a factor in the conversion of thick mat floating marshes to thin mat floating marshes and then to shallow open water (Visser et al. 1999). These effects may extend to forested landscapes. The regeneration of coastal swamps with bald cypress (Taxodium distichum) may have been slowed or even halted by herbivores (fig. 7.6). When bald cypress trees are planted in experiments or for restoration, they are frequently eaten by nutria. Saplings under a halfmeter tall may be cut off at ground level, while larger saplings may be killed when their bark is removed. Myers, Shaffer, and Llewellyn (1995) planted four hundred young bald cypress trees in the Manchac Wildlife Management Area in Louisiana; trees exposed to ambient herbivory suffered 100 percent mortality. Manipulative studies of the interactive effects of soil nutrients, fire, and herbivory illustrate

FIGURE 7.5 Nutria can strip marsh vegetation from large areas of coastal wetland, creating openings called eat-outs. These areas become more vulnerable to further disturbances. From U.S. Geological Survey 2000.

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FIGURE 7.6 Where a vast cypress swamp with trees a thousand years old once stood, coastal marsh dominates today. While succession might have restored such a sight, it is highly unlikely now because hydrological alteration and herbivory by nutria prevent cypress from establishing.

how nutria may affect plant species composition as well as biomass in Louisiana marshes. Ford and Grace (1998b) found that the abundance of Spartina patens, a frequently dominant perennial grass, was reduced by the combination of burning and herbivore exclosure in two marsh communities but not in a third, while other common species responded favorably to the same treatments. In particular, the sedge Schoenoplectus americanus, the preferred food species of muskrats and nutria, increased in relative abundance when mammalian herbivores were excluded. In a similar brackish marsh, Gough and Grace (1998b) documented an increase in biomass of S. americanus and a decrease in S. patens when protected from nutria, despite no change in community biomass after three years of treatment. This shift in dominance was exaggerated, and plant species richness declined when soil nutrients were amended, suggesting an important interaction between herbivory and soil nutrient availability. Ambient herbivory levels in these studies were not as high as those that generate eat-outs, but

these results suggest that moderate nutria activity can affect the structure of these plant communities, particularly when soil nutrient availability is increased, such as after sediment additions to help restore marshes. A more recent study site has been established in the Turtle Cove Experimental Marsh near Southeastern Louisiana University’s Turtle Cove Environmental Research Station to quantify the effects of multiple disturbance treatments, multiple fertility treatments, and the interactions between them, replicated within and outside of 40 ⫻ 60–meter mammalian exclosures (McFalls 2004; Geho, Campbell, and Keddy 2007). The oligohaline marsh was dominated by three species: S. americanus (39.0 percent), Polygonum punctatum (18.9 percent), and Sagittaria lancifolia (7.4 percent). Four ranked disturbance treatments were applied: no disturbance (control), prescribed fire, a single vegetation removal treatment, and a multiple vegetation removal treatment. The ranked fertility treatments were designed to simulate factors that affect production in Louisiana’s rapidly submerging coastal zone: no fertility enhancement (control), sediment addition, fertilizer addition, and a sediment ⫹ fertilizer addition.

Similar to the studies already described, areas protected from nutria had more (1.4 times) vegetation than areas open to herbivory as measured using biomass collected after two years of experimental treatments (McFalls 2004). Biomass steadily decreased with increased disturbance level when nutria were allowed to graze, while this effect was hard to detect in areas protected from nutria herbivory. Fertility and herbivory interacted so that biomass did not appear to increase with increasing fertility unless herbivores were excluded and significant disturbance such as fire or multiple herbicide applications occurred (fig. 7.7). Apparently, nutria had the greatest impact on biomass if another disturbance was already present; that is, they tended to amplify effects of disturbance. The likely mechanism is a preference for newly growing vegetation, a common phenomenon in herbivores (White 1993). Although fire occurs in Louisiana marshes and is a management tool applied to selected marshes (O’Neil 1949; Nyman and Chabreck 1995), the short-term results of this experiment indicate that plants regenerating after fire are particularly attractive to nutria, especially if there is enhanced fertility.

FIGURE 7.7 The effects of nutria on vegetation receiving different fertility and disturbance treatments (mean ⫹ 1 SE, n = 96, pooled into the error term). The bottom panel shows patterns in three fenced exclosures, while the top shows patterns in the three paired open areas. The fertility treatments are control, sediment, fertilizer, and sediment ⫹ fertilizer. The disturbance treatments are control, fire, and single and multiple herbicide application. Adapted from McFalls 2004.

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In summary, the short-term results of work at Turtle Cove in addition to the published studies reviewed earlier suggest that nutria have at least two additional indirect effects on marsh restoration. They may dampen treatments designed to increase production, while they may amplify treatments that increase disturbance. The damping effect may be caused by conversion of plant biomass to nutria biomass. Many biologists advocate increasing nutrient inputs to coastal wetlands (e.g., freshwater diversions, sewage effluent) without considering the possibility that this may simply trigger the growth of nutria. The accelerating effect may be caused by nutria preferring plants with higher nutrient levels produced by new shoots after local disturbances. In this case, initial disturbances from storms or fire may attract nutria, which will prevent regeneration and expand disturbed patches, thereby accelerating the loss of marshes. While these results are drawn from only the first two years of the experiment (McFalls 2004), they illustrate the potential for increased predation on nutria to control a wide array of marsh processes.

THE DIET OF ALLIGATORS Alligators are well-known generalist predators in wetlands. Determining diet from stomach content is always difficult, since the input of prey will vary with habitat, season, and predator size, while volume and digestibility differs among prey species (e.g., fish and turtles). In a recent review, Gabrey (2005) summarized the literature reporting alligator stomach contents. In the 1940s and 1950s, alligators in Louisiana were primarily consuming fish, crustaceans, and muskrats. Nutria were not detected in alligator stomachs until 1961 after their introduction to Louisiana in the 1940s (Valentine et al. 1972). In subsequent studies, nutria comprised a significant portion of the adult alligator diet throughout coastal Louisiana (fig. 7.8), while muskrats declined in importance (Wolfe, Bradshaw, and Chabreck 1987). 122

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FIGURE 7.8 Dead nutria recovered from an alligator’s stomach. Photo courtesy of Steven W. Gabrey, Northwestern State University, Natchitoches, Louisiana.

Juvenile alligators (less than 1.2 meters or 6 feet long) feed primarily on fish, insects, and crustaceans (Platt, Brantley, and Hastings 1990; Wolfe et al. 1987, reviewed in Gabrey 2005). Blue crabs (Callinectes sapidus) can account for 70 percent of prey biomass in brackish marshes (Elsey et al. 1992), but crawfish can dominate prey in fresher areas (Platt et al. 1990). As alligators grow larger, they include vertebrates in their diets, such as deer (Odocoileus virginianus) and other terrestrial mammals (e.g., Shoop and Ruckdeschel 1990). In coastal Louisiana, mammalian prey of adult alligators (greater than 1.2 meters) are dominated by nutria and muskrats (Wolfe et al. 1987). Differences in importance of various components of alligator diets have been correlated with habitat as well as predator size. For example, crustaceans including crabs are more important than nutria for large alligators in saline habitats (reviewed in Gabrey 2005, fig. 2). This suggests another potential trophic cascade in saline marshes: alligators may consume blue crabs that are important predators of snails, known to alter salt marsh vegetation (Silliman and Bertness 2002). Differences in prey composition for alligators likely reflect the habitat preference of nutria for brackish and fresh marshes, rather than a shift in alligator consumption patterns. Mammals have not been found to be important components of alligator diets in the Florida Everglades, probably because

mammal abundance is low (nutria and muskrats do not occur there). In a recent examination of 553 adult alligator stomachs collected from 2002 to 2004 in five parishes in coastal Louisiana, Gabrey (2005) found that approximately 31 percent of alligators had nutria present in their stomachs, while muskrats were found only in approximately 3 percent. Measured as frequency of alligators in which the prey item was found, crustaceans (about 64 percent) and fish (about 51 percent) were more frequently encountered than mammals (about 36 percent). When prey weight was examined, mean weight of mammals was the largest of the prey categories, but variation was too high for statistical analyses to be conducted (Gabrey 2005). Confirming earlier studies of the influence of habitat on alligator diet, alligators from an intermediate marsh had a much higher frequency of crab prey than those in fresh marshes. Also, alligators from western parishes tended to have consumed more turtles than those from eastern parishes where turtles are less common. Thus, this recent study supports earlier research suggesting the importance of nutria as prey for large alligators, but also the idea of alligators as opportunistic foragers, with their diet reflecting the prey available in a particular area. Other linkages undoubtedly occur in the diet of alligators—Gunderson and Loftus (1993) provide food web diagrams illustrating the breadth of freshwater prey consumed by alligators in the Everglades. Quantitative simulations of such food webs show that alligators can have major impacts through indirect linkages to lower trophic levels (Bondavalli and Ulanowicz 1999). McIlhenny (1935) noted that reduced alligator numbers coincided with reduced game fish abundance and attributed this to the release of garfish from alligator predation in freshwater systems in Louisiana. The stomach content data from Louisiana illustrate that the food webs may in fact be dominated by a few strong interactions (sensu Paine 1980)—invertebrates being favored by smaller alligators and nutria by larger alligators (Gabrey 2005).

HISTORICAL EVIDENCE OF CHANGES IN ALLIGATORS, MUSKRATS, AND NUTRIA There are no systematically collected data, but descriptions indicate that alligator populations in Louisiana declined precipitously between 1850 and 1960 (fig. 7.9). McIlhenny (1935), who was a keen naturalist, reported that alligators “fairly swarmed” prior to harvest that began in the 1880s; they remained common until 1900 but were exterminated from many areas of Louisiana by 1935. Initially, only alligators more than 2.4 meters were harvested; but by the 1930s, as the larger alligators disappeared, every alligator that could be captured was harvested (McIlhenny 1935). In the late 1950s, alligator populations in Louisiana remained low because of illegal overharvest (fueled by a demand for skins from small individuals), and alligators as small as 0.6 meter were illegally taken (Joanen and McNease 1987). All trapping was suspended in 1962 (Joanen and McNease 1987), and poaching was virtually eliminated in parts of southwest Louisiana by the early 1960s (Tarver, Linscombe, and Kinler 1987). Alligator populations recovered enough that by 1972 there was an experimental harvest of 1,337 animals in southwest Louisiana (Tarver et al. 1987). Alligator numbers continued to increase; by 1981, the harvest was statewide, and 15,534 hides were taken (Joanen et al. 1984). Data collected by LDWF beginning in 1970 document a relatively steady increase in nests counted throughout Louisiana’s coastal zone (fig. 7.10), with a dip in 2006 attributed to combined detrimental effects of drought and hurricanes in 2005. Nesting quickly recovered in 2007, however, to one of the highest counts on record. Clearly the suspension of hunting and subsequent hunting practices are not affecting recruitment of new individuals into the population. However, larger alligators still are preferentially harvested (Taylor and Neil 1984), making it highly probable that mean size is well below that of the 1850s, and estimates of adult population size are lacking.

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FIGURE 7.9 Overhunting between 1850 and 1960 caused a dramatic decline in alligator populations. It is likely that current population densities, and current alligator sizes, are still well below those encountered. “Alligator shooting in the swamps bordering on the Mississippi River, Louisiana.” Reproduced with permission, Corbis, New York.

Commercial trapping of muskrats began between 1900 and 1910 (O’Neil 1949). O’Neil (1949) concluded from descriptions by early Europeans and interviews with elderly trappers living in the 1940s that the muskrat had been rare in Louisiana coastal marshes prior to the late 1800s, and it had spread westward from the eastern part of the state in the late 1800s. Such increases coincided spatially and temporally with alligator declines reported by McIlhenny (1935). O’Neil reported that muskrat trapping increased as accessible alligators were eliminated and the coastal marshes were burned to make it easier to reach alligators in their holes. Arthur (1928) reported that many people believed that muskrat numbers increased as alligators were reduced but that author did not wholly subscribe to that theory alone. By the 1940s, “the general picture of the better 124

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marshes in every coastal parish is one of ‘eatenout’ marsh due to overpopulation of muskrats” (O’Neil 1949, 70), and in the 1950s, muskrat populations in coastal Louisiana were described as “fantastic” (St. Amant 1959). Nutria were introduced into Louisiana in the 1940s (Lowery 1974; Bernard 2002) and quickly established viable populations. It is thus difficult to separate effects of nutria from those of muskrats after that decade, although trapping data suggest nutria increased in numbers as muskrats declined, perhaps because of competitive exclusion in their overlapping preferred habitat. The increase in muskrats and nutria as alligators declined is well documented but circumstantial. For example, eat-outs became rarer in the 1970s and 1980s at the same time the alligator population increased statewide (Joannen et al. 1984). There are alternative or at least

FIGURE 7.10 Annual alligator nest counts throughout coastal Louisiana show a general increase with time. Data from Louisiana Department of Wildlife and Fisheries.

overlapping hypotheses that might explain such a pattern. For example, in addition to directly reducing alligator numbers, alligator hunters also increased the fire frequency in coastal marshes. Herbivore numbers may have subsequently increased in brackish marshes because the preferred food, S. americanus, is more abundant in the year following a fire than in subsequent years (Nyman and Chabreck 1995). Also, changes in habitat due to road building, oil and gas exploration, draining marshes for pasture, and management practices aimed at wintering waterfowl and muskrats all occurred at the time when alligators were declining and nutria and muskrats were increasing in the 1950s and 1960s.

BOTTOM-UP AND TOP-DOWN CONTROL The concept of the trophic cascade is part of a larger ecological issue—the recognition that there are two possible extremes in the way in which plants (and therefore coastal marshes in particular) are controlled in food webs: bottomup and top-down. Bottom-up hypotheses assume that systems are regulated by nutrient availability from below. They assume that all organisms live under harsh conditions where

there are shortages of resources (plant parts of sufficient quality, prey that are hard to catch), even if these resources seem superficially to be abundant (Sinclair et al. 2000). White (1993) presents an enormous number of examples where nutrients in general, and nitrogen in particular, appear to limit animal populations. In contrast, the top-down view consists of a set of hypotheses about how predators might control the abundance of species lower in the food web. The key issue, from the perspective of this chapter, is the possibility of top-down control, rarely considered in wetlands (Keddy 2000), which may occur along with some degree of bottom-up regulation. For example, Carpenter and Kitchell (1988) note that lake ecosystems can simultaneously exhibit elements of “bottom-up control” through physical factors such as nutrients as well as “top-down control” through biotic interactions such as competition and predation. The current paradigm in coastal management is almost entirely bottom-up: rivers deposit sediments, sediments allow plants to grow, and then the plants are converted into useful products that are harvested. There are several reasons why this perspective predominates. First, there is undoubtedly a geological component to the creation of wetlands

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(e.g., Boyd and Penland 1988), and a certain bottom-up logic is unavoidable: without sediments, there can be no plants. Second, the scientists who study coastal processes are inevitably compartmentalized into subdisciplines based on their training and expertise. For example, plant physiologists study the links between plant growth and sediment characteristics, and rarely interact with those who study predators, while those who study nutria and alligators often have little to do with the botanists. Hence, few look up or down more than one link in the food web. Third, those who look at the entire system have in many cases adopted Odum’s view of energy flow, a view that is inherently bottom-up. Fourth, alligators, like wolves, receive bad press and are seen often as merely an annoyance. Fifth, alligators are seen as an important economic resource and one that should not be restricted given current increasing population estimates. All of these reasons have contributed to the bottom-up focus: sediments make plants, and plants make wildlife. However, some land managers in Louisiana do not harvest all the alligators allowed them by the current system, because they suspect that maintaining many large alligators will reduce nutria and muskrat damage (D. Richard, Stream Wetland Services, LLC, and D. Nuth, National Parks Service, personal communication).

SIMULATION OF POTENTIAL EFFECTS OF ALLIGATOR PREDATION ON NUTRIA POPULATIONS To examine the alligator trophic cascade hypothesis using a modeling approach, we added alligator predation to a previously published nutria population dynamics model (Carter, Foote, and Johnson-Randall 1999). The nutria model was composed of three interacting submodels: a nutria population model, an annual model of aboveground plant biomass, and a marsh surface area model. Marsh loss was controlled by plant biomass: if biomass was above a critical threshold, no area loss occurred. Below this threshold, marsh area loss increased as plant 126

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biomass decreased. As nutria populations increased, herbivory reduced biomass, and marsh loss increased. The original model had a strong seasonal component: nutria populations that appeared sustainable in the summer on a given area of marsh caused marsh loss during the winter months, when plant biomass naturally decreased. The first step in adding a predation component to the model was to modify the published nutria model so that it used a specified numerical carrying capacity instead of a biomass-based carrying capacity. The function of alligator predation was then added to the model as a fixed number of alligators eating a fixed number of nutria per week (determined by sensitivity analysis based on published predation rates; see later discussion). Alligators were assumed to consume nutria for thirty-one weeks per year, thereby simulating seasonal predation patterns. This population model was then allowed to run until dynamic stability near the specified carrying capacity was achieved. In this investigation, “control” was achieved when a nutria population was eliminated within ten years after the start of predation. Alligator population sizes that did not control nutria populations generally reduced mean nutria population size, which would likely have positive consequences for marsh plants. Later we will focus on the more demanding criterion of alligator population sizes that eliminated nutria. Important model assumptions were that alligator predation was strictly additive to “natural” mortality, predation was independent of nutria population density, all alligators fed on all age classes of nutria, and alligator predation was not spatially explicit. To eliminate a nutria population over a ten-year period, the required ratio of alligators to nutria was 0.012 at carrying capacities of five hundred and above. That is, six alligators could eliminate a population consisting of five hundred nutria. As the carrying capacity decreased below five hundred, the ratio of alligator to nutria needed for control increased slightly. This increase was an artifact of alligator and nutria numbers approaching each other in magnitude,

FIGURE 7.11 Ratio of alligators to carrying capacity of nutria needed for control of nutria populations as a function of nutria carrying capacity. Above a nutria carrying capacity of five hundred, the ratio of alligators to nutria is approximately constant at 0.012. The ratio varies at lower carrying capacities because the nutria and alligator numbers are close in magnitude. Weekly alligator predation rate was constant at one for seven months of predation per year.

but the ratio was never higher than 0.067 (fig. 7.11). This result suggests that the larger nutria populations that may occur in especially fertile habitats should be no more difficult to control than those that may occur in less fertile habitats. According to an examination of alligator stomachs by Wolfe et al. (1987), weekly nutria predation rates range from around 0.08 for

smaller adult animals to 1.0 for alligators over three meters in total length. We therefore explored the possible effects of changing predation rates from the low of 0.08 per week to the high of 1.0 per week (fig. 7.12). In our model, the ratio of alligators to nutria needed to eliminate the nutria population increased by an order of magnitude as the predation rate declined by

FIGURE 7.12 Ratio of alligators to carrying capacity of nutria needed for control of nutria populations as a function of predation rate (number of nutria consumed per week). Nutria carrying capacity was constant at one thousand, and duration of predation was seven months per year.

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FIGURE 7.13 Ratio of alligators to carrying capacity of nutria needed for control of nutria populations as a function of the number of months of predation. Nutria carrying capacity was constant at one thousand, and weekly predation rate was one nutria per alligator.

an order of magnitude. For example, a nutria population stabilized around 1,000 could be eliminated by 12 larger alligators with a predation rate of 1 (each alligator eating one nutria per week), or by 120 smaller alligators with a predation rate of 0.1 (each alligator eating one nutria every ten weeks). Thus, one alligator over three meters in length has the effective kill rate of ten small alligators. The ratio of alligators to nutria needed for nutria elimination changed only slightly with the number of months per year that alligators fed (fig. 7.13). For a carrying capacity of one thousand and a predation rate of one per week, the ratio ranged from about 0.01 for eight months of predation to 0.02 at four months of predation. Thus, the number of alligators needed for effective control will likely decrease in climates with longer activity seasons and will vary locally with year to year variation in the weather. Because of the assumptions outlined here, the model may overestimate the impact of alligators on nutria populations, especially for the smaller alligators. Additionally, densitydependent predation rates in nature may result in the nutria population being reduced rather than eliminated. Therefore, this model should not be used to estimate the number of alligators needed to control a given nutria population.

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On the other hand, as a conceptual model, it does demonstrate that (1) a relatively small alligator: nutria ratio can control nutria populations and preserve plant biomass, (2) larger alligators (with higher predation rates) will have a disproportionate effect on nutria population dynamics than smaller animals, and (3) regions with longer alligator activity periods should have more effective control of nutria populations at a given alligator:nutria ratio. Finally, although our goal in this exercise was to predict conditions for the outright elimination of nutria populations, lower alligator population levels or lower predation rates could still effectively reduce the population size of nutria, which could still translate into significant benefits on marsh vegetation.

THE NEED FOR FIELD EXPERIMENTS TO TEST THE ALLIGATOR TROPHIC CASCADE HYPOTHESIS Although some historical trends and natural history observations are consistent with the alligator trophic cascade hypothesis, definitively making these correlations is currently impossible for several reasons. The biggest difficulty is estimating nutria population size and density. Wildlife managers and others have been

attempting to adequately sample the populations for years. Mark–recapture studies have been unsuccessful, as nutria once marked are rarely recovered. Although the state of Louisiana maintains data on nutria trapping, no adequate data exist for nutria population sizes or densities, and trapping data may not be strongly correlated with population parameters. Comparable data for alligator populations are also missing. This may be even more problematic because alligator size and age are correlated with prey choice, and therefore the number of adult alligators (not just the number of individuals) must be known to examine correlations with herbivore population dynamics. Finally, data on plants, nutria, and alligators are not being collected simultaneously using standard protocols that can be compared across sites and regions. In addition, historical data are sketchy, particularly from the period when alligators may have been most common. Muskrats in particular have dramatic population fluctuations, tending to reduce confidence in historic reconstructions. Even if correlative trends exist, they do not demonstrate cause and effect, because many other human activities such as logging, trapping, commercial fishing, and coastal development might be implicated in contributing to such patterns. Areas along the coast exhibit wide variations in alligator and nutria densities (as determined by qualitative observations), creating patterns that are difficult to interpret. In Terrebonne Parish, an area of primarily freshwater marsh, alligator nest densities are the highest that have been documented along the coast, approximately one nest every thirty-five to forty acres (N. Kinler, LDWF, personal communication). Consequently, this parish has one of the highest regulated harvest rates of one alligator from every sixty acres. This is simultaneously an area of great nutria damage and high nutria harvest; approximately 50 percent of the nutria-caused marsh damage in 2007 was in this parish, while in the 2006–2007 trapping season, almost one

hundred thousand nutria were harvested (CWPPRA 2008). In 2002, 13 percent of alligators from this parish had nutria in their stomachs, while 37 percent did in 2004 (Gabrey 2005). There were concurrent increases in crayfish and insect frequencies and decreases in crab frequencies. These data suggest that when habitat conditions are right, both nutria and alligators can flourish, but many questions remain. The high harvest rates (one alligator per sixty acres), for example, may indicate that all but the smallest alligators were removed. In our experience, alligator hunters know the exact locations where the few remaining large alligators live, and they explicitly plan to remove them, leaving ever larger numbers of juveniles. The idea that there are still vast unknown areas of swamp where large alligators can hide is, in our experience, a misunderstanding of how familiar trappers are with local wetlands and how accessible most marshes are to trapping. A further complication is the mobility of nutria. Once nutria have damaged an area, they tend to move to new areas for food. With human harvesting of both nutria and alligators in the parish, both animal populations may be being maintained at levels that allow marsh damage to continue. It remains to be seen what would happen if alligators were not harvested in an area such as this. An appropriate test of the alligator trophic cascade hypothesis would involve manipulation of alligator densities (and possibly size classes) in properly randomized experiments. Owing to the size of alligators, the difficulty of constructing fences in large wetlands, the need for proper replication, and the long duration required for differences to become large enough to be detectable, such experiments require multidisciplinary collaboration among scientists with expertise in sediments, vegetation, and animals. Therefore, we present two different approaches to testing this hypothesis in new field efforts. First, ongoing manipulations of alligator numbers that are in progress could be paired and analyzed. Pairs of wetlands along the Gulf Coast having similar habitat, but differing in

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presence or absence of alligator harvest, or pairs of wetlands known to already differ in alligator density for other reasons, such as past management regimes, could be compared. In each pair of sites, data would be needed on densities of herbivores (particularly, but not exclusively, nutria and muskrat) as well as alligators. Alligator nest density data are available for most coastal regions of Louisiana, but the relationship between nest numbers and densities of adult alligators (and particularly large alligators) is not well understood. Obtaining adult densities and size distributions would be crucial for an adequate test of this hypothesis. Data should simultaneously be collected for biomass and species composition of plants. Ideally, sites that had been trapped or not trapped for at least five years would be paired because it may take many years to reach an alligator density in which large males presumably control recruitment of juveniles. Or pairs of sites could be selected, and then trapping could be imposed on half of them; this would, however, increase the expense and require a longer period of time to answer the question. A second approach would be to compare sites differing in rates of wetland loss. Similar areas currently experiencing high rates of wetland loss could be paired with those that are not, and nutria and alligator densities could be compared. This approach might be hampered by the rapid changes in vegetation composition that occur during rapid loss, potentially making it difficult to pick comparable pairs of sites. But again, without systematically collecting the data, we do not know if areas where marsh damage has not yet occurred are being protected from nutria by predation. Additional approaches involving herbivore exclosures and manipulation of alligator densities are ideal but logistically much less tractable.

MANAGEMENT IMPLICATIONS The most important management implication of the alligator trophic cascade hypothesis is the possible existence of another tool for slow130

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ing marsh loss from herbivore damage: reducing alligator harvest wherever nutria damage is documented. Harvested alligators could be regulated for size to ensure that large alligators capable of eating nutria remain in the population, or alligators could be harvested in proportion to their availability. Perhaps guidelines similar to those used in managed fisheries could be established, such as harvesting intermediatesized alligators to ensure larger animals with higher predation rates were able to persist and better control the nutria population. Our modeling exercise suggests that controlling for alligator size in a harvest would indeed have a dramatic effect on nutria populations. Needless to say, this is a difficult decision to propose for wetland managers in a state like Louisiana, where the wild alligator harvest can be worth $9 million annually to local trappers and processors (LFAAC 2004). In addition, the public is generally fearful of alligators and may not be sympathetic to allowing large animals to increase in number. Given the available data on alligator population size and diet, the severity of current damage from herbivores to coastal wetlands, and the large scale and long duration of the appropriate experiments, we suggest that a collaborative venture is needed to test the alligator trophic cascade hypothesis. In the interim, wetland managers should consider the cautious strategy of allowing for increased density and increased size of alligators in coastal wetlands where nutria and muskrat damage have been documented. As current coastal restoration plans are being evaluated and modified following the 2005 hurricanes (Committee on the Restoration and Protection of Coastal Louisiana 2006), management plans need to incorporate the potential role of alligators in indirectly controlling marsh damage and, perhaps more importantly now, in subsequently affecting restoration efforts. Initial evidence suggests alligators and nutria did not suffer severe mortality from the storms (fig. 7.10; LDWF 2008); therefore, the possibility for this trophic cascade still exists and may become more important given the land loss

caused by the hurricanes and the restoration efforts to come. Acknowledgments. Steven Gabrey generously shared his unpublished report of alligator stomach contents with us and provided us with constructive comments on a draft of this chapter. We thank Len Bahr, Jeff Bounty, and Paul Kemp for helpful discussions; Mark Bertness for comments on an early draft; and the anonymous reviewers for their suggestions. Additional insight into nutria and alligator populations was provided by several staff members at the Louisiana Department of Wildlife and Fisheries, including Noel Kinler. REFERENCES Abraham, K. F., and C. J. Keddy. 2005. The Hudson Bay Lowland. Pages 118–148 in L. H. Fraser, and P. A. Keddy (eds.), 2005. The World’s Largest Wetlands: Ecology and Conservation. Cambridge: Cambridge University Press. Arthur, S. C. 1928. The Fur Animals of Louisiana. Bulletin No. 18. New Orleans: Department of Conservation, State of Louisiana. Barras, J. A., P. E. Bourgeois, and L. R. Handley. 1994. Land Loss Rates in Coastal Louisiana 1956–90. National Biological Survey, National Wetlands Research Center Open File Report 94-01. Lafayette, LA: National Wetlands Research Center. Bartram, W. 1791. Travels through North & South Carolina, Georgia, East & West Florida, the Cherokee Country, the Extensive Territories of the Muscogulges, or Creek Confederacy, and the Country of the Chactaws; Containing an Account of the Soil and Natural Productions of Those Regions, Together with Observations on the Manners of the Indians. Philadelphia: James & Johnson. Electronic version: Documenting the American South, University of North Carolina at Chapel Hill, 2001. Bernard, S. K. 2002. M’sieu Ned’s rat? Reconsidering the origin of nutria in Louisiana: The E. A. McIlhenny Collection, Avery Island, Louisiana. Louisiana History 20: 281–293. Boesch, D. F., M. N. Josselyn, A. J. Mehta, J. T. Morris, W. K. Nuttle, C. A. Simenstad, and D. P. J. Swift. 1994. Scientific assessment of coastal wetland loss, restoration and management in Louisiana. Journal of Coastal Research, Special Issue No. 20. Bondavalli, C., and R. E. Ulanowicz. 1998. Unexpected effects of predators upon their prey:

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herbivores on wetland vegetation in Atchfalaya Bay, Louisiana. Estuaries 21: 1–13. Fahrenthold, D. A. 2004. Blackwater refuge now nutria-free. Washington Post, November 17, p. B1. Ford, M. A., and J. B. Grace. 1998a. Effects of vertebrate herbivores on soil processes, plant biomass, litter accumulation, and soil elevation changes in a coastal marsh. Journal of Ecology 86: 974–982. ———. 1998b. The interactive effects of fire and herbivory on a coastal marsh in Louisiana. Wetlands 18: 1–8. Fuller, D. A., C. E. Sasser, W. B. Johnson, and J. G. Gosselink. 1985. The effects of herbivory on vegetation in Atchafalaya Bay, Louisiana. Wetlands 4: 105–114. Gabrey, S. W. 2005. Impacts of the nutria removal program on the diet of American alligators (Alligator mississippiensis) in south Louisiana. Report to Louisiana Department of Wildlife and Fisheries, New Orleans. Geho, E. M., D. Campbell, and P. A. Keddy. 2007. Quantifying ecological filters: The relative impact of herbivory, neighbours, and sediment on an oligohaline marsh. Oikos 116: 1006–1016. Gosselink, J. G., R. Hatton, and C. S. Hopkinson. 1984. Relationship of organic carbon and mineral content to bulk density in Louisiana marsh soils. Soil Science 137: 177–180. Gough, L., and J. B. Grace. 1998a. Effects of flooding, salinity, and herbivory on coastal plant communities, Louisiana, United States. Oecologia 117: 527–535. ———. 1998b. Herbivore effects on plant species density at varying productivity levels. Ecology 79: 1586–1594. Grace, J. B., and M. A. Ford. 1996. The potential impact of herbivores on the susceptibility of the marsh plant Sagittaria lancifolia to saltwater intrusion in coastal wetlands. Estuaries 19: 13–20. Gunderson, L. H., and W. F. Loftus. 1993. The everglades. Pages 199–255 in W. H. Martin, S. G. Boyce, and A. C. Echternacht (eds.), Biodiversity of the Southeastern United States: Lowland Terrestrial Communities. New York: Wiley. Jeffries, R. L., and R. F. Rockwell. 2002. Foraging geese, vegetation loss and soil degradation in an Arctic salt marsh. Applied Vegetation Science 5: 7–16. Joanen, T., and L. McNease. 1987. The management of alligators in Louisiana, USA. Pages 33–42 in G. J. W. Webb, S. C. Manolis, and P. J. Whitehead (eds.), Wildlife Management: Crocodiles and Alligators. Chipping Norton, Australia: Surrey Teatty.

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Joanen, T., L. McNease, G. Perry, D. Richard, and D. Taylor. 1984. Louisiana’s alligator management program. Proceedings of the Annual Conference of the Southeastern Association of Fish and Wildlife Agencies 38: 210–211. Keddy, P. A. 2000. Wetland Ecology: Principles and Conservation. Cambridge: Cambridge University Press. Keddy, P. A., D. Campbell, T. McFalls, G. P. Shaffer, R. Moreau, C. Dranguet, and R. Heleniak. 2007. The wetlands of lakes Pontchartrain and Maurepas: Past, present and future. Environmental Reviews 15: 43–77. Louisiana Department of Wildlife and Fisheries. 2008. Nutria: Wetland Damage. New Orleans: Author. http://www.nutria.com/site4.php. Louisiana Fur and Alligator Advisory Council. 2004. 2003–2004 Annual Report. Shreveport: Author. http://www.alligatorfur.com. Loveless, C. M. 1959. A study of the vegetation in the Florida everglades. Ecology 40: 1–9. Lowery, G. H. 1974. The Mammals of Louisiana and Its Adjacent Waters. Baton Rouge: Louisiana State University Press. Lynch, J. J., T. O’Neil, and D. W. Lay. 1947. Management and significance of damage by geese and muskrats to Gulf Coast marshes. Journal of Wildlife Management 11: 50–76. McCaffrey, R. J., and J. Thomson. 1980. A record of the accumulation of sediment and trace metals in a Connecticut salt marsh. Advances in Geophysics 22: 165–236. McFalls, T. 2004. Effects of disturbance and fertility upon the vegetation of a Louisiana coastal marsh. Unpublished master’s thesis, Southeastern Louisiana University, Hammond. McGinnis, T. E., III. 1997. Shoreline movement and soil strength in a Louisiana coastal marsh. Unpublished master’s thesis, University of Southwestern Louisiana, Lafayette. McIlhenny, E. A. 1935. The Alligator’s Life History. Boston: Christopher. Republished by the Society for the Study of Amphibians and Reptiles, miscellaneous publications, Facsimile Reprints in Herpetology, 1976. Myers, R. S., G. P. Shaffer, and D. W. Llewellyn. 1995. Bald cypress (Taxodium distichum (L.) Rich.) restoration in southeastern Louisiana: The relative effects of herbivory, flooding, competition and macronutrients. Wetlands 15: 141–148. Nyman, J. A., and R. H. Chabreck. 1995. Fire in coastal marshes: History and recent concerns. Pages 135–141 in S. I. Cerulean and R. T. Engstrom (eds.), Proceedings of the Nineteenth Tall Timbers Fire Ecology Conference—Fire in Wetlands: A Management Perspective. Tallahassee, FL: Tall Timbers Research.

Nyman, J. A., R. D. DeLaune, H. H. Roberts, and W. H. Patrick Jr. 1993. Relationship between vegetation and soil formation in a rapidly submerging coastal marsh. Marine Ecology Progress Series 96: 269–279. O’Neil, T. 1949. The Muskrat in the Louisiana Coastal Marshes. New Orleans: Federal Aid Section—Fish and Game Division, Louisiana Department of Wildlife and Fisheries. Paine, R. T. 1980. Food webs: Linkage, interaction strength and community infrastructure. Journal of Animal Ecology 49: 667–685. Palmer, M. L., and F. J. Mazzotti. 2004. Structure of Everglades alligator holes. Wetlands 24:115–122. Patuxent Wildlife Research Center. 1999. South American nutria destroy marshes. PWRC Fact Sheet 1999–01. http://www.pwrc.usgs.gov/ resshow/nutria.htm. Penland, S., and K. E. Ramsey. 1990. Relative sea-level rise in Louisiana and the Gulf of Mexico: 1908–1988. Journal of Coastal Research 6: 323–342. Platt, S. G., C. G. Brantley, and R. W. Hastings. 1990. Food habits of juvenile American alligators in the upper Lake Ponchartrain Estuary. Northeast Gulf Science 11: 123–130. Reyes, E., M. L. White, J. F. Martin, G. P Kemp, J. W. Day, and V. Aravamuthan. 2000. Landscape modeling of coastal habitat change in the Mississippi Delta. Ecology 81: 2331–2349. Schmitz, O. J., V. Krivan, and O. Ovadia. 2004. Trophic cascades: The primacy of trait-mediated indirect interactions. Ecology Letters 7: 153–163. Shoop, C. R., and C. A. Ruckdeschel. 1990. Alligators as predators on terrestrial mammals. American Midland Naturalist 124: 407–412. Silliman, B. R., and M. D. Bertness. 2002. A trophic cascade regulates salt marsh primary production. Proceedings of the National Academy of Sciences of the USA 99: 10500–10505. Sinclair, A. R. E., C. J. Krebs, J. M. Fryxell, R. Turkington, S. Boutin, R. Boonstra, P. Seccombe-

Hett, P. Lundberg, and L. Oksanen. 2000. Testing hypotheses of trophic level interactions: A boreal forest ecosystem. Oikos 89: 313–328. St. Amant, L. S. 1959. Louisiana Wildlife Inventory and Management Plan. New Orleans: PittmanRobertson Section—Fish and Game Division, Louisiana Wild Life and Fisheries Commission. Tarver, J., G. Linscombe, and N. Kinler. 1987. Fur Animals, Alligator, and the Fur Industry in Louisiana. Baton Rouge: Fur and Refuge Division, Louisiana Department of Wildlife and Fisheries. Taylor, D., and W. Neal. 1984. Management implications of size-class frequency distributions in Louisiana alligator populations. Wildlife Society Bulletin 12: 312–319. Teller, J. T., and L. H. Thorleifson. 1983. The Lake Agassiz–Lake Superior connection. Pages 261–290 in J. T. Teller and L. Clayton (eds.), Glacial Lake Agassiz. Geological Association of Canada Special Paper 26. St. John’s, Newfoundland: Geological Association of Canada. Turner, R. E. 1997. Wetland loss in the northern Gulf of Mexico: Multiple working hypotheses. Estuaries 20: 1–13. U.S. Geological Survey. 2000. Nutria, Eating Louisiana’s Coast. USGS FS-020-00, updated April 20, 2001. Washington, DC: Author. ———. 2005. USGS Reports New Wetland Loss from Hurricane Katrina in Southeastern Louisiana. Washington, DC: Author. http://www.usgs.gov/ newsroom/article.asp?ID=997. Valentine, J. M., J. R. Walther, K. M. McCartney, and L. M. Ivey. 1972. Alligator diets on the Sabine National Wildlife Refuge. Journal of Wildlife Management 36: 809–815. White, T. C. R. 1993. The Inadequate Environment: Nitrogen and the Abundance of Animals. Berlin: Springer. Wolfe, J. L., D. K. Bradshaw, and R. H. Chabreck. 1987. Alligator feeding habits: New data and a review. Northeast Gulf Science 9: 1–8.

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Shoreline Development and the Future of New England Salt Marsh Landscapes Mark D. Bertness, Brian R. Silliman, and Christine Holdredge In spite of the important role that salt marshes play in the ecology of shoreline ecosystems, nearly 50 percent of southern New England salt marshes have already been lost to human development. Here we review recent work that shows that the remaining salt marshes in southern New England are being rapidly degraded by shoreline development. Shoreline development, operationally defined simply as the removal of the woody vegetation bordering marshes, increases nutrient input to marshes and lowers soil salinities as a consequence of increased freshwater runoff that is not intercepted and processed by bordering woody vegetation. On the seaward border of these marshes, nitrogen eutrophication stimulated by local shoreline development is shifting the competitive balance among marsh plants by releasing plants from nutrient competition, leading to the displacement of natural high marsh plants by low marsh cordgrass. On the terrestrial border of these same marshes, shoreline development is precipitating the invasion of the common reed, Phragmites, via nitrogen eutrophication and reduced salinities. Eutrophication of southern New England marshes is also leading to increased consumer control of primary production. A recent fertilization experiment showed that nitrogen eutrophication in southern New England salt marshes has the potential to trigger consumer control of salt marsh primary production by insect herbivores. Quantification of insect grazing on cordgrass at marshes with a range of nutrient inputs reveals that current levels of nitrogen input into southern New England salt marshes have already increased herbivore grazing pressure. Thus, as a consequence of shoreline development, traditional New England salt marsh plant communities and the animals that are dependent on these habitats are being displaced by monocultures of weedy species that are increasingly vulnerable to consumer control. Preserving the woody buffers that border undisturbed marshes and limiting freshwater runoff from the terrestrial system are simple but effective conservation measures that can be taken to preserve New England salt marsh plant communities.

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While salt marshes provide a number of important ecological services in coastal ecosystems, human populations historically have abused them. Salt marshes serve as critical nursery grounds for mobile marine and estuarine species (Boesch and Turner 1984), buffer coastlines from erosion and storm damage (Frey and Bason 1985), and act as chemical filters of terrestrial runoff (Todd and Todd 1994). In spite of these important societal services, humans have long exploited salt marshes in many ways (Ranwell 1967; Dreyer and Niering 1995; Adam 2002). In New England during colonial times, marshes were heavily used for cattle grazing and farmed for the marsh hay, Spartina patens. To encourage marsh hay growth, colonial farmers drained marshes, which reduced marsh water logging and led to higher marsh hay production. In the late nineteenth century and twentieth century, New England salt marshes were extensively drained again to reduce standing water on the high marsh to control mosquito populations. Over 90 percent of southern New England marshes have been ditched (fig. 8.1; K. Bromberg, personal communication). The draining of southern New England salt marshes for both marsh hay farming and mosquito control has been

FIGURE 8.1 Infrared aerial photograph of a heavily ditched southern New England salt marsh.

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hypothesized to have led to the simplicity of these marshes by encouraging the competitive dominance of marsh hay at the expense of the forb community that is characteristic of waterlogged marsh pans in unditched marshes (Ewanchuk and Bertness 2004). In addition to ditching, southern New England marshes have also been filled and drained for land reclamation. It has been estimated using late colonial era maps that nearly 50 percent of the marshes that existed when Europeans colonized southern New England have already been lost to human activities (fig. 8.2; Bromberg and Bertness 2005). Clearly, New England salt marshes are already very heavily impacted systems and far from pristine. Given how much southern New England salt marshes have already been altered by human disturbance and their importance in coastal ecosystems, we need to understand the current condition of southern New England marshes in order to prudently conserve and manage the remaining marshes and the services they provide. In this chapter, we outline recent work that suggests that continuing shoreline development is currently driving massive and potentially irreversible changes in southern New England salt marshes. These findings warn that without strong conservation and management the remaining native southern New England salt marsh systems and many of the important societal services they still provide could easily be lost during our lifetime. This chapter is not a comprehensive review of human impacts on or the general conservation and management of New England salt marshes. In addition to shoreline development, increasing levels of carbon dioxide from the burning of fossil fuels (Curtis et al. 1989) and sea-level rise (Orson, Warren, and Niering 1998; Donnelly and Bertness 2001) are other global human impacts that will affect New England salt marshes over the next century and beyond. These impacts will not be the focus here. We will also not deal explicitly with the issue of what actually constitutes a pristine New

FIGURE 8.2 Salt marsh loss and urbanization of Boston from 1777 to 1999. From Bromberg and Bertness 2005.

England salt marsh and what the target of New England salt marsh conservation and restoration efforts should be. While these are interesting and potentially important topics, we will limit our discussion to conserving and maintaining ecological services and habitat provided by the salt marshes that remain in New England.

ASSEMBLY RULES FOR NEW ENGLAND SALT MARSH PLANT COMMUNITIES New England salt marshes plant communities are simple systems that have long attracted the interest of ecologists (Miller and Egler 1950; Redfield 1972; Niering and Warren 1980; Bertness and Ellison 1987). They have striking vertical plant zonation. The low marsh, or that part of the marsh that is flooded daily by tides, is dominated by a monoculture of the Atlantic cordgrass, Spartina alterniflora. The salt marsh hay, Spartina patens, dominates the seaward border of the high marsh, the part of the marsh that is not flooded on a daily basis. The black rush, Juncus gerardi, dominates the terrestrial border of the high marsh but is replaced by the woody shrub, Iva frutescens, on the extreme terrestrial edge of the high marsh. The spikegrass, Distichilis spicata, and the glasswort, Salicornia europaea, are typically found at low densities in the high marsh, but they can dominate disturbance-generated bare patches before being displaced over time by the zonal dominants (Bertness and Ellison 1987). The mechanisms generating this zonation have been elucidated by transplant experiments (Bertness 1991a, 1991b). When any of these plants are moved to a lower zone, they die with or without plant neighbors. This demonstrates that they are prevented from living at lower elevations by harsh physical conditions. In contrast, when any of these plants are moved to a higher zone, they do just as well or better than they do in their natural zone when plant neighbors are removed; but when plant neighbors are present, they are competitively excluded from the higher zone. Thus, the 140

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upper borders of these marsh plants are set by interspecific competition. The striking zonation in these plant communities is the result of a competitive hierarchy where competitively dominant plants at higher elevations displace competitively subordinate plants to lower elevations or ephemeral disturbance patches. This simple explanation of the striking zonation in these systems is entirely driven by competition for belowground nutrients. Fertilization of zonal borders in these salt marsh plant communities resulted in a complete reversal of the competitive relationships among these plants (Levine, Brewer, and Bertness 1998). With fertilization, the low marsh dominant, Spartina alterniflora, and the high marsh fugitive, Distichlis spicata, became the competitive dominant plants on the marsh. The mechanism of this dramatic shift in competitive dominance is the result of a shift from a nutrient-limited system without fertilization to a light-limited system when nutrients were added (Emery, Ewanchuk, and Bertness 2001). Under nutrient-limited conditions, plants with the best belowground nutrient harvesting roots are the dominant competitors; however, under nutrient-enriched conditions, light becoming limiting, and plants with the most effective aboveground light harvesting structures prevail. In practical terms, this means that under nutrientlimited conditions plants that invest heavily in a dense root mat win competitive encounters; while when nutrients are not limiting, plants that invest heavily in aboveground biomass (i.e., tall plants) win competitive encounters. This mechanistic understanding of the zonation of these marsh plant communities provides an experimentally generated framework for predicting how anthropogenic nitrogen loading will impact New England salt marsh plant communities. Increased nitrogen supply to these marshes should lead to a reversal of the competitive relations of marsh plants and lead to the marsh cordgrass, Spartina alterniflora, invading the high marsh. In the high marsh, the invasive reed, Phragmites australis, is the tallest plant and has also been

shown to competitively dominate under highnitrogen conditions (Minchinton and Bertness 2003). Thus, increased nitrogen supply should lead to Phragmites dominating the high marsh. Phragmites has been part of the New England salt marsh plant community for at least five thousand years. Until recently, it was a minor component of high marsh vegetation, found in low densities in the Iva and Juncus zones (Dreyer and Niering 1995). Over the past number of decades, however, Phragmites has aggressively invaded fresh- and saltwater wetlands in New England (Chambers, Meyerson, and Saltonstall 1999). Saltonstall (2002) has shown that this recent aggressive range expansion of Phragmites in New England has been the result of the introduction of an invasive genotype into North America. Our work (Bertness, Ewanchuk, and Silliman 2002; Silliman and Bertness 2004) reveals that shoreline development and eutrophication are important drivers of this invasion.

SHORELINE DEVELOPMENT AS A DRIVER OF LANDSCAPE CHANGE IN NEW ENGLAND SALT MARSH PLANT COMMUNITIES The woody vegetation that borders undisturbed salt marshes intercepts and processes water and nutrients potentially entering marshes as runoff, and thus buffers them from freshwater and nutrient additions (Valiela et al. 1992; McClelland and Valiela 1998). Our work suggests that shoreline development, simply defined as the removal of the woody plant border buffering marsh wetlands from terrestrial runoff, can increase the flow of nutrient-rich freshwater runoff from terrestrial watersheds onto the marsh surface (Bertness et al. 2002; Silliman and Bertness 2004). To examine the impact of shoreline development on salt marsh nitrogen supply and salinity, we sampled twenty-five Narragansett Bay salt marshes exposed to full-strength seawater (Silliman and Bertness 2004). We found that shoreline development explained 60 percent of intermarsh variation in soil salinity on the high marsh

border, with shoreline development decreasing soil salinity. We further found that shoreline development increased nitrogen supply to Narragansett Bay marshes and explained 66 percent of intermarsh variation in marsh nitrogen supply. Marsh nitrogen supply was quantified as the total aboveground nitrogen in cordgrass at the terrestrial border of the cordgrass zone. This is a good conservative measure of nitrogen supply because cordgrass is nitrogen limited and responds to increased nitrogen supply with increased production and foliar nitrogen concentrations. Thus, total aboveground nitrogen provides an integrated measure of nitrogen supply among marsh sites. By increasing nitrogen and freshwater supply to Narragansett Bay salt marshes, shoreline development explained 39 percent of the movement of cordgrass to higher marsh elevations on shorelines where the woody buffers has been removed (Bertness et al. 2002) and also led to the dominance of Phragmites, explaining 93 percent of the among-site variation in the dominance of Phragmites in Narragansett Bay salt marshes (Silliman and Bertness 2004). Together, these results reveal how shoreline development is dramatically altering southern New England salt marsh landscapes (fig. 8.3). Shoreline development, by removing the woody buffers surrounding pristine salt marshes, increases the flow of freshwater and nitrogen onto marsh surfaces. At the seaward edge of the marsh, increased nitrogen supply shifts S. alterniflora from being a competitively subordinate plant to being a competitive dominant, allowing cordgrass to invade the high marsh and displace S. patens. At the terrestrial border of the marsh, Phragmites also displaces high marsh plants under increased nitrogen supply and decreased soil salinities conditions that bit release from nutrient competition and turn it into the high marsh competitive dominant. The end result is that southern New England salt marshes are turned into clonal monocultures of S. alterniflora and P. australis, at the expense of the local extinction of all the remaining native New England salt marsh plants.

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Spartina alterniflora invades higher elevations

Phragmites australis dominates the terrestrial border and invades lower elevations

Phragmites australis

High marsh forbs - Aster tenuifolitus - Atriplex patula - Limonium nashii - Salicornia evropaea - Solidago sempervirens - Plantago maritima - Glaux martima Spartina alterniflora Zone

Spartina patens Zone

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Iva frutescens Zone

FIGURE 8.3 Conceptual model illustrating how shoreline development impacts Narragansett Bay salt marshes. For details, see Bertness et al. 2002 and Silliman and Bertness 2004.

PHRAGMITES DOMINANCE OF NEW ENGLAND SALT MARSH LANDSCAPES Once Phragmites has overtaken a high marsh habitat, it is able to invade physically harsh lower marsh habitats that it is unable to recruit to by seed. This is because, as a large clonal plant moving across the landscape, Phragmites ramets can spread vegetatively into anoxic, waterlogged soils assisted by clonal integration among ramets (Amsberry et al. 2000). Ramets invading physically harsh habitats are supported by clone mates in less stressful habitats (Shumway 1995). Once Phragmites has successfully invaded a low marsh, it has large ecosystem engineering effects on its habitat (Meyerson, Chambers, and Vogt 1999; Meyerson, Vogt, and Chambers 2000). Phragmites roots and rhizomes are much more extensive than other New England salt marsh plants, often extending over a meter into the substrate, nearly twice that of cordgrass. The extensive production of a dense belowground root and rhizome network by Phragmites 142

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increases the elevation of the marsh, and the large aboveground production of Phragmites increases sediment deposition and the production of aboveground plant biomass and debris. The extensive aboveground biomass of Phragmites also leads to extremely high transpiration rates. The end result is that marsh sediment dries out and the water table drops, reducing soil waterlogging and anoxia stress on Phragmites. This leads to positive feedback where, by ameliorating potentially limiting edaphic conditions, Phragmites performance increases even further, and Phragmites converts invaded marshes into drier habitats that are no longer frequently flooded low marsh habitats. The ecosystem engineering of salt marsh landscapes by Phragmites has a number of important consequences. Reduction of the low marsh habitat reduces the value of the marsh for marine species that rely on intertidal habitats, and the drying of the marsh surface eliminates the high marsh tidal pools and ponds on the marsh surface that are important

nursery areas for juvenile marsh fish (see Kneib 1997; Weinstein et al. 2003). Ecosystem engineering by Phragmites also generates severe problems for the restoration of marshes that have been invaded by Phragmites. Simply killing the Phragmites on a Phragmites-dominated marsh leaves a higher, drier habitat without a tidal low marsh habitat. This habitat will be colonized by high marsh and terrestrial plants rather than the low marsh plants that originally occupied the site. Thus, restoration of a Phragmites invaded marsh to its original state requires not just killing live Phragmites but physically removing the roots, rhizomes, and elevated marsh surface that has been engineered by Phragmites. Because of the extensive ecosystem engineering by Phragmites, the invasion and dominance by Phragmites of southern New England salt marshes can lead to irreversible community changes. In pristine southern New England salt marshes, anoxic salty soils at low marsh elevations provide a competitive refuge habitat for competitively subordinate salt marsh plants (Chambers, Mozdzer, and Ambrose 1998; Crain et al. 2004). The plants that occupy the high marsh, including Phragmites, are unable to colonize low marsh elevations because they are unable to cope with anoxic salty soils (Bertness 1999). Thus, physical stress maintains the strong spatial patterns on undisturbed New England salt marshes, and, when naturally disturbed by ice scour, storm overwash, or wrack deposition, these communities recover to a typically zoned New England salt marsh plant community. Once Phragmites invades a New England salt marsh landscape, however, ecosystem engineering by Phragmites leads to the loss of low marsh elevations and the drying out of the marsh surface, alleviating the waterlogging and anoxia stresses that maintain the spatial structure and zonation of undisturbed New England salt marshes. If a Phragmitesdominated salt marsh is disturbed, it will not recover into a traditional New England salt marsh, but it will be recolonized by Phragmites or other high marsh plants. Thus, colonization

and dominance of New England salt marsh landscapes by Phragmites is an example of habitat modification by an invading ecosystem engineer that leads to a community state change. If disturbed, the Phragmites marsh community will return to its new community state rather than its original community state.

DOES SHORELINE DEVELOPMENT LEAD TO CONSUMER CONTROL OF NEW ENGLAND SALT MARSH PLANT COMMUNITIES? One of the most basic questions about the ecology of natural ecosystems is whether the distribution, abundance, and production patterns are controlled by physical factors and nutrient supply or bottom-up forces, or by consumers or top-down forces (Power 1992). Salt marsh plant ecosystems have long been considered a classical example of an ecosystem under strong bottom-up control where physical soil conditions and nutrient availability controlled primary production patterns (Teal 1962). The role of consumers in salt marsh systems has long been thought to be trivial. Work in a variety of salt marsh systems, however, has begun to challenge this view. Initial evidence that consumers could play a leading role in controlling salt marsh primary production came from systems where human intervention could be seen as leading to consumer impacts (feral horses, Furbish and Albano 1994; invasive nutria, Gough and Grace 1998; snow geese, Jefferies 1997). Recent work in the extensive salt marshes of southeastern North America, however, has shown that the common marsh periwinkle, Littoraria irrorata, long thought to be a detritivore, can exert strong top-down control over marsh cordgrass production (Silliman and Zieman 2001; Silliman and Bertness 2002; Silliman and Bortolus 2003). Similarly, on the Atlantic coast of South America, the grapsid crab, Chasmagnathus granulata, has been shown to be capable of strongly impacting salt marsh primary production and plant distribution patterns (Bortolus

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4 FIGURE 8.4 Total live dried cordgrass biomasses per one hundred square centimeters from our preliminary fertilization experiment at Cogshall Marsh on Prudence Island in the summer of 2004.

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and Irbarne 1999; Costa, Marangoni, and Azevedo 2003). In spite of the fact that herbivory has not received much critical attention in New England salt marshes, it has long been assumed to play an insignificant role in the community and ecosystem dynamics of these habitats (Nixon 1982; Teal 1985). Insect herbivores such as grasshoppers and beetles have been shown to feed very little on the leaves and other somatic tissue of the numerically abundant clonal turf plants that dominate New England salt marsh landscapes and feed almost exclusively on the flowers and seeds of the clonal plants that dominate New England salt marshes (Bertness, Wise, and Ellison 1987) and the less abundant forbs that live inconspicuously hidden in the matrix of unpalatable clonal plants (Rand 1999). Vince, Valiela, and Teal (1981), however, found that insect abundance increased dramatically in salt marsh plots that had been enriched by nitrogen fertilizer. This suggests that eutrophication could lead to increasing herbivory on New England salt marsh plants. We performed a fertilization experiment at Cogshall marsh on Prudence Island, Rhode Island. This experiment was designed to examine the importance of nutrients in helping New 144

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Fertilized

England salt marshes keep up with sea-level rise, and it was not originally intended to capture plant productivity or herbivory. The experimental design consisted of four large (4 ⫻ 4–meter) plots separated by over ten meters in intermediate cordgrass (about ten meters from the creek bank). Two of the plots were left as control, and two of the plots were fertilized monthly from May to August 2004 with Scott’s Turfbuilder NPK (see Levine et al. 1998 for methods). Cordgrass responded vigorously to fertilization, and by mid-June, S. alterniflora in the fertilizer plots was conspicuously taller and more robust than in adjacent unfertilized controls so that the differences in productivity were apparent from fifty meters away. By July, however, insect damage was evident in the fertilized plots though virtually absent in control plots. By August, grazing on cordgrass in the fertilized plots was extensive and had clearly depressed plant biomass. By the end of the summer, grasshoppers and leafhoppers had damaged over 90 percent of the cordgrass stems in the fertilized plots, while in the control plots less than 15 percent of the cordgrass stems were damaged, and the extent of insect damage differed dramatically between fertilized and control plants.

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This damage translated into a severe suppression in end-of-the-season cordgrass biomass (fig. 8.4). While the biomass of fertilized plots had been conspicuously higher than the control plots in the early summer (probably double, but not measured), by the end of the growing season, insect herbivory, mostly from grasshoppers, led to a dramatic decrease (40 percent) in cordgrass biomass. Thus, fertilization appeared to trigger consumer suppression of salt marsh primary production. Interestingly, we had not seen any evidence for increased herbivory in any of our previous nutrient addition experiments in New England salt marshes (Levine et al. 1998; Emery et al. 2001). Since all of our previous nutrient addition experiments had been done at smaller, 0.5 ⫻ 0.5–meter plot, spatial scales, this suggests that consumer responses to nutrient additions may be size-dependent and may need a critical size threshold to occur. Nutrient enrichment at large spatial scales is already occurring in marshes within Narragansett Bay, and this gradient in marsh eutrophication has been quantified (Silliman and Bertness 2004). Relatively small-scale experimental studies within a single marsh may misrepresent ecological processes at larger spatial scales as enriched plots could attract insect herbivores that would forage more diffusely were nutrients to be added at larger spatial scales, as predicted with whole-marsh eutrophication. The established eutrophication gradient

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FIGURE 8.5 Plant damage (total length in centimeters of leaf damage per cordgrass culm) at seventeen Narragansett Bay marshes as a function of marsh nitrogen supply.

in Narragansett Bay marshes provides a unique opportunity and ideal system to examine this question of whether small-scale experimental results actually scale up to landscape-scale patterns. To examine whether current eutrophication of southern New England salt marshes could already be triggering increased insect herbivory, in August 2004, we visited the seventeen Narragansett Bay marshes where we have quantified relative nutrient status (see Silliman and Bertness 2004). At each site, we scored twentyfive random cordgrass culms at the upper border of the cordgrass zone for extent of insect damage. We found that increased nitrogen input led to increased herbivore damage and that intermarsh variation in total nitrogen explained over 40 percent of intermarsh variation in plant damage due to herbivory (fig. 8.5). These results provide strong preliminary evidence that current levels of eutrophication in Narragansett Bay may already be triggering consumer control of southern New England salt marsh production.

THE FUTURE OF SOUTHERN NEW ENGLAND SALT MARSHES Together, our recent results paint a grim future for New England salt marshes. Nearly 50 percent of southern New England salt marshes have already been lost to development

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(Bromberg and Bertness 2005), and our recent studies reveal that continuing shoreline development is rapidly and predictably degrading the remaining marshes. Our work suggests that shoreline development, by removing the woody vegetation buffers that process terrestrial runoff entering salt marshes, is increasing nitrogen loading and decreasing soil salinities in southern New England. This is resulting in the cordgrass and Phragmites competitive dominance of these salt marsh landscapes and the local and regional extinction of traditional salt marsh landscapes that were historically maintained by nitrogen limitation. Moreover, since Phragmites is a formidable ecosystem engineer, once it dominates a salt marsh, it increases the elevation of the marsh and draws down the water table, removing low marsh habitats and the surface heterogeneity that many salt marsh animals utilize and converting wetlands into habitats more suitable for dominance by terrestrial plants. Thus, the invasion of Phragmites ultimately leads to a community state change where a habitat once supporting wetland plants is converted to a habitat supporting a community favorable to terrestrial plants. Evidence also suggests the nitrogen loading in southern New England salt marshes is triggering increased herbivore pressure and potentially could lead to consumer control of marsh primary production. This is of particular concern since increased marsh primary production has already been shown in the arctic marshes of Hudson Bay (Jefferies 1997) and the southern marshes of Georgia (Silliman and Bertness 2002) to trigger salt marsh die-off driven by runaway consumption. Could eutrophication of New England salt marshes trigger similar marsh die-offs? Population growth and shoreline development are leading to a total transformation of New England salt marshes. Is this something we should care about? Shoreline development is clearly leading to the local and regional extinction of the native salt marsh plant community, but what is most important is not the native plants but the ecological services provided by salt marshes. In terms of ecological 146

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services, some appear to be in jeopardy with the conversion of traditional New England marsh landscapes to cordgrass/Phragmites monocultures. The nursery ground function of salt marshes may be largely lost when marshes are converted to cordgrass/Phragmites monocultures. Many estuarine fish and shellfish rely on standing water on marsh surfaces and intertidal marsh habitats to recruit to as young refuges from their water column predators (Kneib 1997). Both of these habitats are largely lost in cordgrass/Phragmites monoculture marshes (see Weinstein et al. 2003). The aesthetic value of marshes and many of the birds that live in traditional salt marsh landscapes may also be largely lost with cordgrass/Phragmites monocultures. In contrast, the chemical processing of terrestrial runoff (Todd and Todd 1994), the buffering of shorelines from erosion and storm damage (Frey and Basan 1985), and increased marsh accretion to keep pace with accelerating sea-level rise (Donnelly and Bertness 2001) may actually be enhanced by the dominance of cordgrass/Phragmites marshes. The simple management message of this work is that maintaining natural woody vegetation borders around marshes buffers them from acute human impacts. When woody borders need to be removed, diverting terrestrial runoff from the marsh surface will likely serve the same purpose as natural woody borders. Conserving the woody vegetation bordering salt marshes is a simple management strategy that can be used at local and regional spatial scales to conserve traditional salt marsh landscapes. This is a particularly powerful message for homeowners who often clear woody vegetation surrounding marshes to create a water view and then complain about invading Phragmites obstructing their view. By cutting woody vegetation borders and replacing them with watered and fertilized grass, homeowners for all practical purposes are farming Phragmites. Public education of the direct causal link between intact marsh borders and Phragmites invasion could limit the spread of Phragmites in New England salt marshes.

Acknowledgments. We gratefully acknowledge Kenny Reposa and the staff of the Narragansett Bay National Estuarine Research Reserve for facilitating the fieldwork presented in this chapter. This chapter benefited from the comments of Keryn Bromberg, Caitlin Crain, Ted Grosholz, and three anonymous reviewers. Our work was supported by grants from the Ecology Program of the National Science Foundation and Rhode Island Sea Grant. REFERENCES Adam, P. 2002. Saltmarshes in a time of change. Environmental Conservation 29: 39–61. Amsberry, L., M. A. Baker, P. J. Ewanchuk, and M. D. Bertness. 2000. Clonal integration and the expansion of Phragmites australis into New England salt marsh plant communities. Ecological Applications 10: 1110–1118. Bertness, M. D. 1991a. Interspecific interactions among high marsh perennials. Ecology 72: 125–137. ———. 1991b. Zonation of Spartina spp. in New England salt marshes. Ecology 72: 138–148. ———. 1999. The Ecology of Atlantic Shorelines. Sunderland, MA: Sinauer. Bertness, M. D., and A. M. Ellison. 1987. Determinants of pattern in a New England salt marsh plant community. Ecological Monographs 57: 129–147. Bertness, M. D., P. Ewanchuk, and B. R. Silliman. 2002. Anthropogenic modification of New England salt marsh landscapes. Proceedings of the National Academy of Sciences of the USA 99: 1395–1398. Bertness, M. D., C. Wise, and A. M. Ellison. 1987. Consumer pressure and seed set in New England marsh perennials. Oecologia 71: 191–200. Boesch, D. F., and R. E. Turner. 1984. Dependence of fishery species on salt marshes: The role of food and refuges. Estuaries 7: 460–468. Bortolus, A., and O. Iribarne. 1999. Effects of the SW Atlantic burrowing crab Chasmagnathus granulata on a Spartina salt marsh. Marine Ecology Progress Series 178: 78–88. Bromberg, K., and M. D. Bertness. 2005. Reconstructing New England salt marsh losses using historical maps. Estuaries 28: 823–832. Chambers, R. M., L. A. Meyerson, and K. Saltonstall. 1999. Expansion of Phragmites into tidal wetlands of North America. Aquatic Botany 64: 261–273. Chambers, R. M., T. J. Mozdzer, and J. C. Ambrose. 1998. Effects of salinity and sulfide on the distrib-

ution of Phragmites australis and Spartina alterniflora in a tidal marsh. Aquatic Botany 62: 161–169. Costa, C. S. B., J. Marangoni, and A. G. Azevedo. 2003. Plant zonation in irregularly flooded salt marshes: The relative importance of stress tolerance and biological interactions. Journal of Ecology 91: 951–965. Crain, C. M., B. R. Silliman, S. L. Bertness, and M. D. Bertness. 2004. Mechanisms of the spatial segregation of plants across estuarine salinity gradients. Ecology 85: 2539–2549. Curtis, P. S., B. G. Drake, P. W. Leadley, W. J. Arp, and D. F. Whigham. 1989. Growth and senescence in plant communities exposed to elevated CO2 in situ. Oecologia 78: 20–26. Donnelly, J., and M. D. Bertness. 2001. Rapid shoreward encroachment of salt marsh vegetation in response to sea-level rise. Proceedings of the National Academy of Sciences of the USA 98: 14218–14223. Dreyer, G. D., and W. A. Niering. 1995. Tidal marshes of Long Island Sound: Ecology, history and restoration. Connecticut College Arboretum Bulletin No. 34. Emery, N., P. Ewanchuk, and M. D. Bertness. 2001. Nutrients, mechanisms of competition and the zonation of plants across salt marsh landscapes. Ecology 82: 2471–2485. Ewanchuk, P. J., and M. D. Bertness. 2004. Maintenance of high diversity pans in northern New England salt marshes. Ecology 85: 1568–1574. Frey, R. W., and P. B. Basan. 1985. Coastal salt marshes. Pages 225–301 in R. A. Davis (ed.), Coastal Sedimentary Environments. New York: Springer. Furbish, C. E., and M. Albano. 1994. Selective herbivory and plant community structure in a midAtlantic salt marsh. Ecology 75: 1015–1022. Gough, L., and J. B. Grace. 1998. Herbivore effects on plant species density at varying productivity levels. Ecology 79: 1586–1594. Jefferies, R. L. 1997. Long-term damage to sub-arctic coastal ecosystems by geese: Ecological indicators and measures of ecosystem dysfunction. Pages 151–166 in R. M. M. Carawford (ed.), Disturbance and Recovery in Arctic Lands: An Ecological Perspective. Boston: Kluwer Academic. Kneib, R. T. 1997. Tidal marshes offer a different perspective on estuarine nekton. Annual Review of Oceanography and Marine Biology 35: 1–120. Levine J., S. J. Brewer, and M. D. Bertness. 1998. Nutrient availability and the zonation of marsh plant communities. Journal of Ecology 86: 285–292. McClelland, J. W., and I. Valiela. 1998. Linking nitrogen in estuarine producers to land-derived sources. Limnology and Oceanography 43: 577–585.

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Meyerson, L. A., R. M. Chambers, and K. A. Vogt. 1999. The effects of Phragmites removal on nutrient pools in a freshwater tidal marsh ecosystem. Biological Invasions 1: 129–136. Meyerson, L. A., K. A. Vogt, and R. M. Chambers. 2000. Linking the success of Phragmites to the alteration of ecosystem nutrient cycles. Pages 817–834 in M. Weinstein (ed.), Concepts and Controversies in Tidal Marsh Ecology. Philadelphia: Academy of Natural Sciences. Miller, W. B., and F. E. Egler. 1950. Vegetation of the Wequetequock-Pawcatuck tidal marshes, Connecticut. Ecological Monographs 20: 143–172. Minchinton, T. E., and M. D. Bertness. 2003. Disturbance-mediated competition and the spread of Phragmites australis in a coastal marsh. Ecological Applications 13: 1400–1416. Niering, W. A., and R. S. Warren. 1980. Vegetation patterns and processes in New England salt marshes. BioScience 30: 301–307. Nixon, S. W. 1982. The ecology of New England high salt marshes: A community profile. Biological Report 81, no. 55. Washington, DC: U.S. Fish and Wildlife Service. Orson, R. A., R. S. Warren, and W. A. Niering. 1998. Interpreting sea level rise and rates of vertical marsh accretion in a southern New England tidal salt marsh. Estuarine, Coastal and Shelf Science 47: 419–429. Power, M. E. 1992. Top-down and bottom-up forces in food webs: Do plants have primacy? Ecology 73: 733–746. Rand, T. A. 1999. Effects of environmental context on the susceptibility of Atriplex patula to attack by herbivorous beetles. Oecologia 121: 39–46. Ranwell, D. S. 1967. World resources of Spartina townsendii and economic use of Spartina marshland. Journal of Applied Ecology 4: 239–256. Redfield, A. C. 1972. Development of a New England salt marsh. Ecological Monographs 42: 201–237. Saltonstall, K. 2002. Cryptic invasion by a nonnative genotype of the common reed, Phragmites australis, into North America. Proceedings of the

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National Academy of Sciences of the USA 99: 2445–2449. Shumway, S. W. 1995. Physiological integration among clonal ramets during invasion of disturbance patches in a New England Salt marsh. Annals of Botany 76: 225–233. Silliman, B. R., and M. D. Bertness. 2002. A trophic cascade regulates salt marsh primary production. Proceedings of the National Academy of Sciences of the USA 99: 10500–10505. ———. 2004. Shoreline development drives the invasion of Phragmites australis and the loss of New England salt marsh plant diversity. Conservation Biology 18: 1424–1434. Silliman, B. R., and A. Bortolus. 2003. Underestimation of Spartina alterniflora production in western Atlantic salt marshes. Oikos 101: 549–554. Silliman B. R., and J. C. Zieman. 2001. Top-down control of Spartina alterniflora production by periwinkle grazing in a Virginia salt marsh. Ecology 82: 2830–2845. Teal, J. 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43: 614–624. ———. 1985. The Ecology of New England Low Marsh Habitats: A Community Profile. Washington, DC: U.S. Fish and Wildlife Service. Todd, N. J., and J. Todd. 1994. From Eco-Cities to Living Machines. Berkeley, CA: North Atlantic Books. Valiela, I., P. Peckol, C. D’Avanzo, C. H. Sham, and K. Lajtha. 1992. Couplings of watersheds and coastal waters: Sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts. Estuaries 15: 443–457. Vince, S. W., I. Valiela, and J. M. Teal. 1981. An experimental study of the structure of herbivorous insect communities in a salt marsh. Ecology 62: 1662–1678. Weinstein, M. P., J. R. Keough, G. R. Gutenspergen, and S. Y. Litvin, eds. 2003. Phragmites australis: A sheep in wolf’s clothing? Estuaries 26, no. 2B: 397–630.

9

Tidal Restrictions and Mosquito Ditching in New England Marshes case studies of the biotic evidence, physical extent, and potential for restoration of altered tidal hydrology Caitlin Mullan Crain, Keryn Bromberg Gedan, and Michele Dionne Hydrology is the principal physical driver of pattern and process in intertidal coastal marshes. Tidal hydrologic regimes in marshes, including inundation frequency, duration, and depth, control the physicochemical environment (i.e., waterlogging, salinity, and nutrient flux) and ultimately determine the species of plants able to colonize, persist, and create habitats in these biogenic systems. A long history of human use of coastal areas has impacted natural tidal regimes in numerous ways with consequent alteration of physical parameters, the biota, and ecological interactions in coastal marshes. Here we summarize research investigating the importance of hydrology to marsh ecology and then highlight two major hydrologic alterations within coastal marshes of New England. Specifically, we examine the history, physical and biological impacts, and potential for restoration of tidal restrictions and mosquito ditching in New England. Both of these alterations have extensively changed natural tidal marsh hydrology and left widespread signatures on species assemblages and marsh function. The synthesis of anthropogenic impacts to marsh hydrology, and concomitant changes in marsh structure and function, described here, provides a conceptual framework for understanding the need, importance, and potential for hydrological restoration to coastal marshes of North America.

Coastal marshes are by definition intertidal biogenic habitats; the environment drives vegetation patterns, and the vegetation, in turn, creates the ecosystem. Arguably, the single most important physical driver of the coastal marsh ecosystem is the hydrologic regime (Mitsch and Gosselink 2000). Tidal inundation not only

directly defines suitable marsh habitat but indirectly controls numerous physical variables important for marsh structure and functioning (fig. 9.1). Given this fundamental importance of tidal hydrology to marsh pattern and process, it is no surprise that the extensive human manipulations

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Biotic feedbacks Peat accretion Hummock formation Evapotranspiration Decomposition-induced subsidence

HYDROLOGY Tidal inundation Water Table level Water flux

Marsh Ecosystem EDAPHIC FACTORS Oxygen/Sulfides Salinity pH

Biotic Feedbacks Shading Evapotranspiration Oxygenation of rhizosphere

BIOTIC STRUCTURE Species Composition Zonation

Biotic Feedbacks Species interactions

ECOSYSTEM PROCESSES Nutrient cycling Productivity Diversity Decomposition

FIGURE 9.1 A simple conceptual model of the way in which hydrology influences important physical and biotic components, including drivers and feedbacks, of salt marsh structure.

of hydrologic regimes have had major impacts on coastal marshes. Humans drain, dam, divert, channelize, and tidally restrict water resources in estuarine marshes (Dionne, Short, and Burdick 1998). These pervasive human impacts have left significant anthropogenic signatures on North American marshes. Hydrologic alterations have not only altered the physical environment of marshes but additionally impact species interactions, community processes, and marsh ecology. In this chapter, we review major ways in which tidal hydrology drives the structure and function of coastal marshes. We then focus on human impacts due to two predominant types of hydrologic alterations in Atlantic Coast marshes of northeastern North America: tidal restrictions and ditching for mosquito control. For both of these human impacts, we review the historical motivation and present data examining the spatial extent of hydrologic modifications and their impacts in New England. We then describe what is known about how modifications alter physical conditions and concomitant impacts on marsh biota. Finally, we examine projects aimed at marsh restoration 150

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and highlight gaps in ecological understanding that need to be addressed to improve restoration success. This synthesis begins to reveal the full spatial extent and ecological impacts of these hydrological alterations, hinting at the widespread alteration of marsh structure and function that historical human modifications have left. Lessons from this synthesis can be applied to coastal marsh systems elsewhere in North America and globally, highlighting the need for well-researched hydrologic restorations to protect these valuable ecosystems.

HYDROLOGY AND PHYSICAL AND BIOLOGICAL MARSH PROCESSES Coastal marshes occupy a narrow zone of intertidal habitat on protected middle- and high-latitude shorelines worldwide. Marsh distribution is determined by tidal amplitude: the low border of the low marsh is generally determined (and defined as) mean low tide, since salt marsh macrophytes cannot tolerate constant flooding; while the upland border is determined by highest high tides, since upland vegetation has a competitive advantage under nonflooded

conditions (Adam 1990). Thus, tidal flooding sets the stage for marsh establishment and drives the distribution, zonation, and dynamics of dominant marsh plants. Hydrologic regimes not only directly define suitable habitat but indirectly drive the numerous edaphic or physical variables that determine marsh structure by controlling the physicochemical environment (fig. 9.1). Distance from tidal creeks and relative tidal elevation determine inundation frequency and duration, drainage rates, and tidal flushing (Zedler et al. 1999). Poorly drained soils become waterlogged, with low oxygen diffusion rates driving low oxidation-reduction (redox) potential in the soil (Armstrong 1979). Low redox, in turn, leads to buildup of toxic sulfides, making nutrients less accessible and therefore reducing plant productivity (Bradley and Dunn 1989; Koch and Mendelssohn 1989; Koch, Mendelssohn, and McKee 1990). In contrast, tidal flushing can refresh sediments, removing toxins and delivering nutrients. Therefore, creek banks tend to be well drained and subsidized environments while the inner marsh is more waterlogged (Mendelssohn and Morris 2000). The physical work of tidal flushing alone has been proposed to provide an energy subsidy to marshes that stimulates growth (Steever et al. 1976) and may be partly responsible for making salt marshes one of the most productive ecosystems in the world (up to 3,900 g C/m2/yr). In addition to marsh morphology, soil waterlogging can be accelerated by ice scouring (Ewanchuk and Bertness 2003) or subsidence (Delaune, Nyman, and Patrick 1994) that effectively lowers the relative height of the marsh surface. Soil waterlogging decreases when ditches (Buchsbaum 2001) or tidal creeks penetrate into the marsh, when plants engineer elevated mounds (Fogel et al. 2004), or when the groundwater table is lowered. Hydrologic regimes also drive a second defining physical variable in the marsh: salinity. Tidal inundation brings salt into marsh sediments and—together with freshwater inputs, soil texture, depth to water table, and vegeta-

tion—drives salinity patterns in salt marsh soils (Mitsch and Gosselink 2000). Regular tidal flushing in low marsh elevations results in more consistent salinity regimes, while higher marsh elevations are subject to greater variation in soil salinity. Particularly in lower-latitude marshes, high marsh elevations are subject to high temperatures and high evaporative stress that concentrates salts in the soil. The hypersalinities that accumulate in these marsh pannes can kill vegetation, leaving bare salty sediments. Gradients in edaphic factors driven by hydrologic regime form the foundation for marsh plant and animal community dynamics. Variation in soil redox and salinity with distance from tidal channels drives species zonation (Bertness 1991) and shifts from tall to short growth forms within monospecific stands of Spartina alterniflora (salt marsh cordgrass) (Mendelssohn 1979; Howes, Dacey, and Goehringer 1986). Variation in water discharge from upland sources also influences edaphic conditions in the high marsh, contributing to species invasions (Silliman and Bertness 2004). In addition to determining dominant space holder identity, edaphic gradients influence species interactions between primary producers and the whole marsh community. Both waterlogging and salinity have been shown to influence competitive hierarchies (Bertness 1991; Pennings and Callaway 1992; Bertness and Ewanchuk 2002; Crain et al. 2004), dependence on engineering plants (Fogel et al. 2004), direction of species interactions (Bertness and Shumway 1993; Bertness and Ewanchuk 2002; Crain 2008), and herbivory rates and trophic structure (Hacker and Bertness 1995, 1996; Srivastava and Jefferies 1996; Gough and Grace 1998; Rand 1999; Silliman and Zieman 2001; Silliman and Bortolus 2003). For example, trade-offs in tolerance to physicochemical environment and competitive ability drives marsh plant zonation across tidal elevations in salt marshes (Bertness 1991) and across salinity gradients in estuaries (Crain et al. 2004). Saltstressed plants attract increased grazing by

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invertebrate herbivores, leading to shifting plant–herbivore interactions with varying salinity stress (Silliman et al. 2005). Once intertidal marshes develop, the biotic communities themselves can feedback and influence hydrological variables. Organisms known as ecosystem engineers (Jones, Lawton, and Shachak 1994) alter their physical or chemical environments and can have major impacts on environmental variables and associated organisms. Most basically, plants transpire water, potentially altering hydrologic budgets (Mitsch and Gosselink 2000). Plant structure slows water movement, altering sediment deposition patterns (Stumpf 1983; Reed 1989; Morgan and Short 2002). Wetland plants often oxygenate their rhizosphere (Armstrong 1979) or lower the water table through evapotranspiration (Dacey and Howes 1984), making anoxic sediments more tolerable, and shade the substrate, reducing salt accumulation (Shumway and Bertness 1994). Finally, plants create raised rooting mounds in poorly drained sediments to evade waterlogging (Crain and Bertness 2005).

ANTHROPOGENIC ALTERATIONS TO MARSH HYDROLOGY Humans have a long history of physically altering coastal marsh environments. In Europe, marshes have been diked and drained since the early 1500s. When Europeans colonized North America, the presence of salt marsh was an important factor in the establishment of coastal New England towns (Russell 1976) due to the abundance of important resources. European colonists brought with them hydrologic management techniques including ditching and diking. With modernization, hydrologic alterations became progressively more invasive and extensive. Not until the 1970s, when ecosystem services of coastal marshes were increasingly recognized by scientists, were efforts made to protect, and more recently restore, proper marsh functioning. Despite these recent efforts, we are left with a legacy of major hydrologic alterations that have changed physical and ecolog152

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ical processes in marshes that we now attempt to quantify and rectify. In northeastern North America, estuarine marshes are the dominant coastal habitat type, perform essential ecosystem functions, have been degraded by different types of human interventions, and are now the focus of coastal restoration efforts. Here we outline two major anthropogenic modifications of tidal hydrologic regime in New England marshes in order to expose the far-reaching effects of these impacts, examine restoration potential, and foster insights on the impacts of hydrologic alterations in marsh systems worldwide.

TIDAL RESTRICTION A variety of engineering techniques have resulted in reduced tidal influence to New England marshes (fig. 9.2), most notably the creation of berms or dikes for road and railroad causeways, flood control, or wildlife enhancement (Buchsbaum 2001). Tidal restrictions were implemented as early as the 1600s when colonists brought coastal engineering techniques from Europe to improve agricultural use of the marsh. Salt marsh farmers created berms, mounded fill across the marsh platform that restrict sheet flow of water across the marsh surface, to reduce waterlogging above the berm and promote grazing areas and production of salt marsh hay (Smith and Bridges 1982; Sebold 1998). Tide gates, which limit creek channel flow, were also employed for marsh management and for tide mills (Rozsa 1995). Tidal restrictions became more widespread and restrictive when coastal development increased in the 1900s, especially during and after World War II (Nixon 1982). During this period, extensive causeways and dikes were created for roadways and coastal flood management, respectively. These restrictions not only reduced sheet flow across the marsh surface but reduced to varying degrees or completely eliminated tidal water to upriver marshes due to inadequately sized culverts (Buchsbaum 2001).

(A)

(C)

SPATIAL EXTENT OF TIDAL RESTRICTIONS IN NEW ENGLAND

The spatial extent of various types of tidal restrictions has never been fully quantified in New England or elsewhere. Regional assessments, however, make inferences possible (see table 9.1). Historical diking, draining, and filling are thought to have eliminated half of the coastal marshes of Massachusetts (Nickerson 1978) and Connecticut (Niering and Bowers 1966). Of the remaining marshes, diking is estimated to impact 10 percent of the marshes in Connecticut (Roman, Niering, and Warren 1984) and 25 percent in New Hampshire (U.S. Department of

(B)

FIGURE 9.2 A long history of tidal restrictions can be found in New England estuaries, such as these examples from Wells, Maine. (A) An early dike and berm were constructed to improve salt marsh farming; (B) a box culvert (left) replaces an undersized culvert (right) that restricts sheet and channel flow in Drake’s Island Marsh leading to (C) impounded water, species, and ecosystem changes upriver. In this aerial photo, remnants of the old dike can be seen in front of the causeway. Photos courtesy of M. Dionne.

Agriculture, Soil Conservation Service 1994). In Cape Cod, hundreds of acres of salt marsh were diked and subsequently ditched by the 1930s (Portnoy and Giblin 1997b). Extensive tidal restrictions extend to the north, into the Canadian Bay of Fundy, where of the nearly eighty thousand acres of original tidal marsh referenced by Ganong (1903), more than 75 percent have been impounded by dikes for agriculture beginning in the early 1600s (Ganong 1903; Gordon and Cranford 1994; Percy 1998). The majority of these dikelands currently lay fallow. While the number of tidal restrictions has been assessed in various New England regions,

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TABLE 9.1 Summary table of regional assessments of tidal restrictions and their impacts on coastal marshes in northeastern North America compiled from various sources

Region Canada Massachusetts New Hampshire Maine Maine (midcoast) Maine (southern)

Total Shoreline (km)

Total Number of Restrictions

Number of Restrictions Impacting Marsh

Area of Marsh Impacted

1,300 2,445 211 5,597 1,905 584

307 621 100 1,197 272 283

——— ——— ——— 500 121 57

——— ——— 526 ha (20%) 2,044haa 902 ha (28%)

NOTE: Survey techniques and data collected vary by region, and thorough assessment of restrictions impacting marsh and areal extent of impacts has only been done accurately for southern Maine estuaries, analyzed here (see table 4.2). Sources: for Canada, Van Proosdij and Dobek (2005), for Massachusetts, Cape Cod Commission (2001), Buzzards Bay National Estuary Program (2002), and Massachusetts Wetlands Restoration Program (2002, 2003); for New Hampshire USDA (1994); for Maine, Bonebakker et al. (2000) and USACOE (2004); for midcoast of Maine, USACOE (2004); for southern Maine, current study. a

Likely an overestimate since marsh upriver of multiple restrictions was counted more than once.

the spatial extent of these impacts has rarely been systematically quantified. Here, we present such an analysis for the southern coast of Maine (fig. 9.3 and table 9.2). A tidal restriction database was created using U.S. Geological Survey topographical maps followed by ground truthing as possible (75 percent of the map-identified sites). All road crossings within the twenty-nine estuarine bodies (encompassing 3,193 hectares of tidal wetland) between Kittery to Cape Elizabeth were identified. Of the 283 total restrictions, 57 were determined to be affecting a total of 902 hectares of tidal wetland on the upstream side. In this analysis, the area of wetland affected by each tidal restriction was bounded by the next upstream restriction, assuring that affected area was not overestimated. Our analysis shows that 28 percent of the halophytic tidal marsh in the region is affected to some degree by tidal restriction. Thirty-three additional restrictions were located at the upper limits of saltinfluenced wetlands, with halophytic marsh occurring immediately downstream and fresh marsh immediately upstream. The widespread occurrence of restrictions near the “head of tide” indicate potentially significant reductions in the extent of freshwater tidal wetlands in this region. This analysis from southern Maine is impressive considering tidal restrictions are 154

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likely more common and impact more marsh in southern New England and the mid-Atlantic. These results highlight the need for regional assessments that offer large-scale perspectives to help identify the main factors threatening marsh structure and function.

PHYSICAL IMPACTS OF TIDAL RESTRICTION

Tidal restrictions, due to a variety of human modifications, result in reduced tidal influence to upriver or inland parts of the marsh. Marsh systems upriver of restrictions have reduced salinities, since marine influence is low and freshwater is impounded, and in extreme cases are transformed from tidal to standing water systems. The influence of restrictions on upriver marshes depends both on the type of restriction and on additional drainage patterns. Open restrictions, such as culverts, have less impact on upriver marshes than complete restrictions, like tide gates, that can lead to dry impounded marshes when river discharge is low. Complete restrictions that are nondrained can convert to standing freshwater marshes (Montague et al. 1987), while those that are additionally ditched for mosquito control can result in subsiding acid sulfate marshes or terrestrial habitats (Portnoy 1999). In addition to

FIGURE 9.3 Boundaries of the twenty-nine tidal watersheds in southern Maine (from Cape Elizabeth to the New Hampshire border) used to estimate the extent of marsh area impacted by 283 tidal restrictions. Insets for two example watersheds (1) York River and (2) Spruce Creek show the distribution of individual tidal restrictions (filled circles).

TABLE 9.2 Metrics describing the numeric and spatial extent of tidal restrictions in the arcuate bay region of the Maine coast (Kittery to Cape Elizabeth; see fig. 9.4)

Regionwide

tidal restriction metrics Total salt marsh area identified (ha) Total number of restrictions identified Total number of head tide restrictions Total number of restrictions affecting marsh Mean number of restrictions affecting marsh (per system) Mean area affected (per system) (ha) Total marsh area affected by restriction (ha) Total salt marsh area affected by restriction (%)

Within Example Estuaries

southern maine

spruce creek

york river

3,193 283 33 57 2

35 11 3 3 n/a

206 43 13 9 n/a

76.9 (⫾20) 902 28.3

n/a 6 16

n/a 134 65

NOTE: National Wetlands Inventory Maps were used to delineate wetland distribution and identify wetland area impacted by tidal restrictions. Note that the number of head tide restrictions are not included in any of the other metrics, since wetland habitat above these restrictions has been totally converted, making the areal extent of impacts difficult to quantify. Two example estuarine systems (corresponding to fig. 9.4) provide examples of metric range and variation observed in the 29 estuarine systems in southern Maine.

reducing the total tidal influence, tidal restrictions change both the tidal amplitude and vertical position of the tide relative to mean low water, for instance, in impounded marshes that never drain completely or marshes below restrictions that experience increased flooding. These altered hydrologic regimes, in turn, influence all physical parameters characteristic of intertidal salt marshes. Portnoy and Giblin (1997a, 1997b) describe the numerous effects tidal restrictions have on marsh biogeochemistry, particularly highlighting reduction in porewater nutrients, sulfides, and increased acidity in restricted marshes. Biogeochemical responses in tidally restricted marshes depend on whether restricted marshes are flooded through freshwater inputs or are additionally ditched and drained. Flooded fresh marshes develop more organic soils than comparable salt marshes since inorganic sediment supply from flood tides is reduced and since decomposition by methanogenesis in fresh marshes is slower than sulfur reduction in salt marshes (Portnoy and Giblin 1997a). In contrast, drained freshwater soils have a low water table, enhanced by increased transpiration rates of fresh marsh plants (Hussey and Odum 1992). In these marshes, marsh peat aeration 156

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leads to both higher decomposition rates, pyrite oxidation, and subsequent soil acidification (Portnoy and Giblin 1997b). Marsh elevation loss due to subsidence will occur in all marshes that have been sufficiently restricted, especially those completely impounded by the use of one-way gates that prevent tidal influx. Subsidence occurs through oxidation of the organic component of salt marsh peat, which increases decomposition rates and reduces soil bulk. The mineral component of the peat is then released and ultimately can fill in former tidal creeks and channels, reducing the overall elevation gradient of the restricted marsh system. High decomposition rates lead to substantial (up to a meter documented) marsh surface subsidence in restricted marshes (Burdick et al. 1997). In addition to the actual loss of existing peat, creating new peat in tidegated marshes is a challenge. For one, restricted marshes are deprived of marine sediments that typically fuel the process. Additionally, peatbuilding salt marsh plants decline in abundance over time, generally overtaken by other wetland plants that do not accrete peat in the same way (e.g., Typha, Phragmites). These changes eliminate the ability of the impounded marsh to respond to sea-level rise through the normal

plant-mediated peat-building process (Warren and Niering 1993; Boumans, Burdick, and Dionne 2002). The vegetative and geomorphological changes resulting from subsidence must be considered when designing projects to restore natural patterns of tidal flooding and drainage. BIOTIC RESPONSE TO TIDAL RESTRICTION

The most obvious organismal response to tidal restriction is the conversion of dominant marsh plants from salt marsh halophytes to fresh marsh species. The shift happens gradually and results in conversion to brackish or fresh marsh species, more often than not, dominated by the invasive genotype of Phragmites australis (Roman et al. 1984). Salt-tolerant halophytes find competitive refuge in salt marshes, but under nonsaline conditions, fresh marsh species that do not invest in physiological mechanisms of salt stress tolerance are competitively superior (Crain et al. 2004) and will slowly suppress salt marsh vegetation. Beyond the conversion to fresh marsh plant species, the inevitable shift in complex biotic interactions under tidally restricted conditions have not been investigated. One exception is a study by Konisky and Burdick (2004) that experimentally investigated the relative competitive ability of plant species pairs in relation to flooding and salinity. More studies of the shifts in species interactions under altered hydrologic regimes are needed. Tidal restrictions limit marine-estuarine energy exchange, including movement of transient and migratory nekton; however, the degree to which tidal restriction alters marsh use by nekton is less clear. Evidence exists that removing tidal restrictions improves marsh fish passage and connectivity (Eberhardt 2004) and increases nekton density and richness (Roman et al. 2002), but some studies have shown that tidally restricted marshes have similar nekton use as nonrestricted reference marshes (Dionne et al. 1999; Raposa, Roman, and Heltshe 2003). Results may be skewed due to sampling techniques (Dionne et al. 1999) or predominance by just one or several generalist species. Evidence

from studies conducted in the Bahamas shows that the degree of hydrologic connectivity between estuarine and marine waters directly affects fish communities; specifically, fragmented estuaries have highly variable fish assemblages (Layman et al. 2004). This suggests that partial connectivity through culverts, common in New England estuaries, likely results in highly variable fish assemblages with consequent inconsistency in trophic interactions. Tidal restrictions may lead to the loss of important ecosystem services of estuarine systems. Impounded marshes dominated by fresh marsh and invasive plants undoubtedly perform different ecosystem functions than lost salt marshes, though whether new functions sufficiently replace those lost is unknown. Reduced fish passage through tidal restrictions is particularly important since it can disrupt the trophic relay whereby the nekton food web links small forage fish in marsh creeks and channels with larger predator fish in estuarine and nearshore marine waters (Kneib 1997). This subsidy supports many of the large fisheries humans depend on. In addition, once salt marshes degrade or transform, changes to trophic structure and dependent plant and animal communities cascade through the whole ecosystem, with the concomitant loss of numerous ecosystem functions (Roman et al. 1984). Impacts of tidal restrictions depend on the degree to which marine connectivity is reduced (Layman et al. 2004), but they inevitably lead to degradation of upriver marsh systems and their services. TIDAL RESTORATION

Due to increased awareness of the importance of coastal marshes, attempts have been made to regain functioning salt marshes through tidal restoration (Turner and Lewis 1997). Restoration has proceeded due to both planned restoration and unplanned breaches of dams or tide gates. Numerous studies document shifts back from fresh to salt marsh vegetation with tidal restoration (Sinicrope et al. 1990; Burdick et al. 1997; Warren et al. 2002). Recovery rate may be correlated with degree of flooding rather than salinity

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(Warren et al. 2002); however, greenhouse studies show fresh marsh plants die of salinity stress before waterlogged-induced sulfides could accumulate (Portnoy and Giblin 1997a). Comparisons between planned and opportunistic restorations (Burdick et al. 1997) and survey work (Montalto and Steenhuis 2004) highlight the importance of restoring an appropriate hydrological regime to accomplish marsh recovery. Portnoy and Giblin (1997a) caution that abrupt return of tidal flushing to freshwater marshes could produce detrimental and undesirable effects on the marsh and nearby coastal waters. Reintroduction of salt water to fresh flooded marshes could lead to increased subsidence, root death, and sulfide toxicity and in fresh drained marshes could lead to massive nutrient release (Portnoy and Giblin 1997a). For these reasons, they advocate gradual restoration of tidal flushing. Beyond dominant species shifts, the health and functioning of restored marshes is largely unknown but of utmost importance. Neckles et al. (2000) advocate use of a series of indicators and developed a protocol for assessing the success of tidal marsh restoration. These indicators have been applied to more than thirty paired salt marsh restoration and reference sites throughout the Gulf of Maine (Konisky et al. 2006). Results indicate that tidally restricted salt marshes have lower salinities and greater abundance of brackish marsh plant species prior to restoration. After restoration, tidal prisms increase, and marshes become more saline and have greater abundance of salt marsh plant species. Effects of tidal restoration on nekton and birds are inconclusive due to insufficient data. Coordinated regional monitoring of restored and reference marsh sites continues in the Northeast, with improvements in methods and implementation of avian and nekton assessment.

MOSQUITO DITCHING Another of the most apparent and extensive human alterations to New England tidal marshes is ditching, which was done to drain 158

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standing water from the marsh to control mosquito populations. The practice of marsh ditching was introduced in New England by colonists and early American farmers, to encourage the growth of commercially valuable salt marsh hay, Spartina patens, which they noted grew better and was easier to harvest on well-drained soils (Sebold 1992). However, the majority of the ditching was done in the first half of the twentieth century, as mosquito control techniques were being developed to limit the frequency of human disease within residential and urban areas (Smith 1904). Many large marsh systems were ditched during the 1930s by the Civilian Conservation Corps (CCC), when men were employed to hand-dig mosquito ditches in parallel grid patterns, primarily to boost the national economy and secondarily for mosquito control (Reiley 1936; Stearns and MacCreary 1934). During this time, many areas were ditched that were undoubtedly never suitable mosquito larval habitat (Provost 1977). Salt marsh mosquitoes breed in shallow surface waters in areas of the high marsh and on moist peat, where eggs will hatch only when inundated by tidewater. The rational behind ditching is to drain breeding habitat and also control mosquitoes by biological control, by allowing larvivorous fish, such as the mummichog, Fundulus heteroclitus, and the sheepshead minnow, Cyprinodon variegatus, access to their mosquito prey within the marsh (Smith 1904). Suggested sizes for mosquito ditches range from fifteen to twenty-five centimeters wide and sixty to seventy-five centimeters deep (Daiber 1986), but the dimensions of ditches dug throughout the last century are much more variable, ranging anywhere from ten centimeters to two meters wide and just as deep. Mosquito ditching in New England has been generally confined to the high marsh zones of Spartina patens, Distichlis spicata, and Juncus gerardi, but occasionally ditches have been dug in the low marsh S. alterniflora zone as well. Although some ditches have been widened by erosion, many others have filled in with S. alterniflora in the years since their construction

(A) FIGURE 9.4 Aerial photos of two heavily ditched marshes (ditching severity index ⫽ 8; see fig. 9.5) with ditches highlighted in (A) Barnstable, Massachusetts (Great Sippewesset Marsh) and (B) Marshfield, Massachusetts. (B) Photos from Massachusetts Geographic Information Systems (MassGIS).

(see Miller and Egler 1950 for a description of ditch aggradation processes). SPATIAL EXTENT OF DITCHING IN NEW ENGLAND

Mosquito ditches are a conspicuous feature of New England salt marshes (fig. 9.4). Bourne and Cottam (1950) estimated that as early as 1938, mosquito ditches had been dug in more than 90 percent of all salt marshes between Maine and Virginia. We surveyed digital orthophotos of thirty-two New England salt marshes (21 percent of all marsh area in the region), selected randomly along the coast from Connecticut to Maine, and found that 94 percent of these marshes had been ditched (fig. 9.5). Ditching intensity, quantified by ditch area divided by total marsh area, varied greatly from marsh to marsh. Ditching intensity varied significantly by state (fig. 9.6) and was significantly lower in Maine than Massachusetts marshes, with the only unditched marshes (two) found being in Maine (fig. 9.6). These survey results suggest that ditching impacts are differentially distributed throughout New

England. While greater ditching intensity clearly disrupts marsh structure, the long term impacts of ditching on salt marsh structure and function are poorly understood. In fact, our current understanding of New England marshes may be based on systems that have shifted significantly to altered states due to extensive human manipulation. How ditching has fundamentally altered the nature of marshes and their functioning must be investigated to fully quantify the impacts and effectively move toward integrated restoration and mosquito management techniques. Here we outline the known physical and biotic impacts of ditching to synthesize what is currently known about ditching impacts and identify gaps in knowledge that must be addressed. PHYSICAL IMPACTS OF MARSH DITCHING

The immediate impact of ditching is removal of a large volume of peat from the marsh, which is then spread more diffusely across the marsh surface. From a very basic perspective, this results in a loss of marsh habitat and an increase in tidal aquatic habitat. The sudden aeration of

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ME VT

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peat, both ditch spoil and on ditch banks, likely results in a large carbon and nutrient release. After this initial impact, the main effect of ditching is to increase the marsh area inundated and drained on each tidal cycle as tidewaters access and leave the marsh via ditches (Daiber 1986). Despite the dense, rhizophorous peat of the high marsh, mosquito ditches have been shown to increase peat drainage up to five meters from a ditch (Daigh, MacCreary, and Stearns 1938; Resh 2001; K. Bromberg, unpublished data). The water table level in marsh adjacent to ditches has been observed to drop five to ten centimeters relative to unditched control areas (Stearns, MacCreary, and Daigh 1940; Whigham, O’Neill, and McWethy 1983). Depth of drainage is likely dependent on both the depth of the ditch and the tidal elevation. Many ditches run nearly dry at low tide, exposing both the adjacent marsh and subsurface peat to atmospheric conditions. The soils 160

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4 7 5 4 6 6

FIGURE 9.5 Ditching frequency and severity across marshes in New England. A survey of orthophotos of thirty-two marshes showed that 94 percent of New England tidal marshes are ditched. Marshes were ranked by ditching intensity on a scale of 0 to 10, with 0 being unditched and 10 being heavily ditched. In the most heavily ditched marshes, greater than 2.5 percent of the marsh area has been converted to ditches. The figure shows the spatial distribution of ditching intensity that varies by region with the greatest ditching impacts in Massachusetts.

along ditch banks are aerated during low tides and flushed during high tides, ameliorating the hydrogen sulfide and sodium chloride buildups that waterlogged areas of the marsh typically experience (King et al. 1982; Ewanchuk and Bertness 2004). The amelioration of these stresses likely leads to increased plant productivity. Indeed, increased productivity of short-form S. alterniflora was observed by Shisler and Jobbins (1977) near recently constructed ditches. The authors attributed the productivity increase to the increased marsh tidal subsidy caused by ditching and a possible increase in nitrogen imports, which may also play a role in increasing plant productivity near ditches (Steever et al. 1976). Effects of ditching on the peat-building capability of a marsh have not been investigated, but ditches potentially affect four major factors that regulate the height of the marsh surface relative to sea level: belowground plant production, peat decomposition, sediment deposition, and

number of surface water bodies in ditched marshes (Adamowicz and Roman 2005). BIOTIC RESPONSE TO MARSH DITCHING

FIGURE 9.6 Ditching intensity (as a percentage of total marsh area converted to ditch area) by state. State is a significant factor explaining differences in ditching intensity (ANOVA, F ⫽ 5.582, p ⬍ 0.01). A Tukey post hoc test shows that Maine marshes are significantly less ditched than Massachusetts marshes.

sediment erosion. Adjacent to ditches, belowground plant productivity may increase due to favorable plant conditions described earlier, while increased substrate aeration adjacent to ditches may result in faster microbial decomposition of organic peat. In addition, sedimentation and erosion processes are also affected by increased tidal flux near to ditches. The effect of ditches on these factors that interactively determine the rate of peat accretion over the long term is unknown, despite the persistence of ditches over many decades and the growing concern over whether salt marshes can keep pace with sea-level rise. Side effects of ditch construction include spoil deposition both on the marsh surface and in the surrounding estuary (Bourn and Cottam 1950), as well as natural levee formation on the sides of larger ditches (Miller and Egler 1950). In many cases, spoils have been deposited directly beside ditches, resulting in unnatural high-elevation zones inhabited by high marsh vegetation in the middle of the marsh system. Recolonization rate and plant species composition of spoil deposits depends greatly on the thickness of deposition (Burger and Shisler 1983). Spoils have also been used to fill in smaller marsh depressions (Smith 1904), a practice that further contributed to reducing the

As ditching disrupts many of the hydrodynamic and biogeochemical processes of the salt marsh, it is not surprising that shifts in salt marsh vegetation have resulted. The drainage of large areas of marsh allowed, at least initially, invasion of formerly low marsh areas by high marsh vegetation (Headlee 1935). In a detailed study of vegetation shifts in a Delaware marsh in response to ditching, Bourn and Cottam (1950) found a precipitous decline in the spatial extent of Spartina alterniflora accompanied by an increase in S. patens and the woody species Iva fructescens and Baccharis halimifolia. In fact, the typically high marsh shrub expanded so extensively along ditch banks that ecologists referred to the “Iva problem” (Miller and Egler 1950). Later research demonstrated that Iva distribution is limited by waterlogging and therefore can grow well along well-drained ditch banks often further raised by ditch spoils (Bertness, Wikler, and Chatkupt 1992). Because mosquito ditching drains surface water from the high marsh, ditching may have decreased the extent of salt marsh forb pannes, areas of marsh with high diversity and lowdensity plant cover. Forb pannes in New England are maintained by waterlogging stress, which keeps out dominant grasses and makes pannes a competitive refuge for stress tolerant forb species (Ewanchuk and Bertness 2004). A slight lowering of the water table can result in incursion of forb pannes by S. patens (Ewanchuk and Bertness 2004). Forb pannes are less common in southern New England marshes, a pattern that may be a product of more extensive ditching in these marshes (Ewanchuk and Bertness 2004). Although published monitoring data is rare, ditching does seem to successfully suppress mosquito populations, with population reductions of an order of magnitude (DeBord, Carlson, and Axtell 1975; Wolfe 1996). Due to high ditch frequency, grid ditches may control

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the entire salt marsh mosquito population; however, selective ditching may only displace mosquito breeding to other parts of the marsh. More landscape-scale data is needed to know whether ditching is truly an effective mosquito control. In addition to the targeted effect on mosquito populations, ditching impacts other invertebrate groups. Bourn and Cottam (1950) reported reductions in invertebrate populations in ditched areas compared to unditched areas in the three years following ditch construction, and the authors suggest that this could have negative implications for higher trophic levels. Ditching effects on invertebrates are often seasonally and species-dependent. Lesser, Murphy, and Lake (1976) found an increase in fiddler crab density in a Delaware ditched marsh, which they attributed to an increase in ditch bank habitat, a mimic of creek banks where fiddler burrows are normally found. In a study of a Pacific Coast salt marsh, Resh and Balling (1983) found a decrease in terrestrial arthropod diversity adjacent to newer ditches during the wet season, but not the dry season. The authors suggest winter washout tides reduced arthropod diversity near newer ditches, while higher plant productivity adjacent to older ditches protected arthropods from tides and maintained diversity. Aquatic invertebrates, many of which overlap in ecological requirements with mosquitoes, seem to be especially hard-hit by ditching. Data from ditched marshes on the Pacific coast show that aquatic invertebrate diversity, but not biomass, is reduced near ditches, where rare species are lost and common species dominate (Resh 2001). Negative effects on aquatic invertebrates are probably a result of both a reduction in habitat for species that utilize surface waters and an increase in top-down control by fish and other marine invertebrates that have increased access to high marsh waters via ditches. Judging by the magnitude of several reported reductions in mosquito populations, the effect on other aquatic insect populations has probably been underestimated. As aquatic invertebrates are a major component of salt marsh trophic webs (Weisberg and Lotrich 1982), this 162

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(A)

(B)

effect could have larger implications for other salt marsh taxa. Ditching has positive repercussions for fish populations, by increasing the extent of tidal aquatic habitat, improving fish access to invertebrate prey, and providing a predation refuge from larger fish in shallow ditches (Valiela et al. 1977; Wolfe 1996). In a ditched Pacific Coast salt marsh, fish diversity increased twofold, and fish density increased threefold (Wolfe 1996). As fish that forage on the marsh surface represent a vital link between salt marsh and pelagic

(C)

(D)

food webs, the increase in fish populations could result in increased connectivity between these two ecosystems. The effect of ditching on wildlife, particularly bird and mammal species of commercial or recreational value, has been inconclusive and highly contentious (Daiber 1986). Although negative impacts of ditching on bird abundances have been predicted based on inferential evidence such as shifts in prey populations (aquatic invertebrates and terrestrial and marine plants; Bourn and Cottam 1950), others have attributed

FIGURE 9.7 Efforts to restore salt marshes are becoming more common, though successful return of marsh function is often unknown. Tidal restoration by (A) replacing one-way tide gates and (B) dam removal generally lead to salt marsh plants in previously impounded marshes. The long-term impacts of ditch plugging (C and D, from Sprague River, Phippsburg, Maine) to reverse the impacts of mosquito ditching are less well understood. Photos courtesy of M. Dionne and C. M. Crain.

observed decreased bird use of ditched marshes to the loss of salt marsh pools (Clarke et al. 1984; Reinert, Goley, and DeRagon 1981). DITCHING RESTORATION AND NEW APPROACHES TO MOSQUITO CONTROL

Current efforts in mosquito control have shifted to targeting larval source reduction at specific areas of high mosquito production. Ditching is done selectively to limit environmental impacts and deepwater fish reservoirs are created in the high marsh to encourage fish residence in

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mosquito breeding habitat (Daiber 1986; Wolfe 1996). This newer control approach has been termed quality ditching (as in “quality, not quantity,” to emphasize the departure from the haphazard and extensive ditching done in the 1930s), but it is more commonly termed open marsh water management, or OMWM (Ferrigno and Jobbins 1968). In OMWM, the goal continues to be to drain the surface water where mosquitoes breed and to increase predatory fish access, but expansive ditches are replaced with shorter ditches connecting depressions and shallow, transient pools (breeding grounds) to deeper, permanent ponds that act as fish reservoirs. Additionally, grading of spoils and careful deposition on the marsh surface is emphasized. In OMWM, large volumes of peat are removed for the creation of fish reservoirs, and this peat is sometimes used for ditch restoration, also called ditch plugging. Ditches are plugged with peat and lined with a marine plywood barrier to create a long, narrow fish reservoir and to prevent further draining effects of the ditch (fig. 9.7). This is a very new practice, and there is little information on how the ditch ponds function or affect the rest of the marsh. Turning the former ditch into an artificial pond will not necessarily remove the prior effects of the ditch on the adjacent salt marsh, although certainly further drainage is prevented. In addition, ditch-plugging restoration goals are not explicit or not known, making evaluation of success difficult. OMWM projects are not uncommon in New England. Many states have a department or agency responsible for mosquito abatement, and these agencies often own their own construction equipment, specially outfitted for low-impact marsh access (A. Gettman, personal communication). A report in 1998 by the Gulf of Maine Council on the Marine Environment found thirty-five separate OMWM sites (including ditch plug sites) in Massachusetts, New Hampshire, and Maine alone (Cornelisen 1998). OMWM tries to reduce physical alteration of the marsh and use an understanding of salt

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marsh ecology to accomplish mosquito control. This seems an improvement over more destructive mosquito control efforts, and some of the drainage problems associated with extensive ditching may be avoided. However, many projects are progressing before a proper scientific understanding of OMWM has been developed. A recent U.S. Geological Survey study of experimental OMWM projects in five National Wildlife Refuges (including Parker River National Wildlife Refuge and Stewart B. McKinney National Wildlife Refuge in New England) found mostly positive effects of OMWM on nekton and birds, but negative effects on vegetation at some sites, including an increase in bare space and a decrease in Iva frutescens (James-Pirri, Erwin, and Prosser 2005). However, the study also found no detectable effect of OMWM on mosquito populations. These data were collected a year after establishment of the OMWM projects, and more time may be needed to clarify the effects of OMWM on flora and fauna. But, clearly, long-term monitoring and evaluation for mosquito control efficacy of these experimental projects are necessary before continued use of this technique.

RESEARCH NEEDS Further research on the biological effects of tidal alterations is needed that document the far-reaching effects, in both space and time, of these physical modifications. Quantifying the extent and cascading biotic impacts of anthropogenic impacts will document “shifting baselines,” the altered state of marsh structure and function within a historical context (Jackson 2001), and inform restoration efforts. Despite increasing societal investment, numerous improvements could be made to increase salt marsh restoration success. First, comprehensive surveys of reference, restricted/ditched and restored sites where physical variables and indicators of marsh function are measured, would establish a baseline

understanding of ecosystem services lost and gained in various anthropogenic environments to enable restoration goal setting and success evaluation. Monitoring studies should be designed as “quasi-experiments” that can be statistically analyzed (Block et al. 2001). Further research should be question driven and designed to examine mechanisms of change, rather than impact detection alone. Research on long-term restoration success and natural marsh recovery from tidal alterations is also essential. The thresholds of hydrologic change that lead to loss and recovery of marsh self maintenance have yet to be determined (Boumans et al. 2002) but are critical to understanding when systems need restoration. For example, some mosquito ditches seem to be filling in, and restoration is unwarranted in these cases. Until more research is done, restoration projects should proceed cautiously. In other cases, restoration potential is unknown. Biotic feedbacks in systems may drive altered states, even when physical conditions are restored. For instance, the subsidence that occurs in tidally restricted marshes potentially makes recovery to a previous state very difficult, but biotic feedbacks are rarely considered in restoration projects. Finally, undesirable side effects of restoration may outweigh benefits and must be considered. For instance, unvegetated areas resulting from OMWM projects, ditch plugging, and tidal restoration may provide competitive refuge for invasive species to colonize and establish under low-salinity conditions. All of these factors need to be investigated to assure salt marsh restoration achieves desired goals, improves on natural recovery, and is as successful as possible.

CONCLUSION As evidenced from New England, hydrologic alterations to salt marsh systems in North America (1) are widespread, (2) have far-reaching impacts in space and time, (3) substantially alter coastal marsh structure and function, and (4) are inadequately understood. Efforts to reverse the effects

of historical human impacts are well intended, but they often have little scientific guidance, goals, or measures of success. In this chapter, we have outlined the importance of marsh hydrology to ecosystem structure and function and have reviewed the historical impetus, physical and biotic consequences, spatial extent, and restoration efforts surrounding two major hydrologic alterations in New England. This synthesis leads to a number of important research questions that must be addressed to better understand the current state of our coastal marshes and improve restoration efforts. Currently, the political, managerial, and ecological environment is ripe to tackle the issues outlined here that could lead to increased coastal marsh protection and restoration success. Acknowledgments. This work was supported by funding from Brown University and the Wells National Estuarine Research Reserve. We would like to thank Ray Konisky, Matthew McBride, and Lynne Carlson for supplying photographs and GIS mapping support and analysis, and two anonymous reviewers who helped improve the manuscript greatly. Thanks to Brian Silliman and Mark Bertness for editorial help, ecological insight, and many fun years exploring salt marshes around the world.

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Konisky, R. A., and D. M. Burdick. 2004. Effects of stressors on invasive and halophytic plants of New England salt marshes: A framework for predicting response to tidal restoration. Wetlands 24: 434–447. Konisky, R. A., D. M. Burdick, M. Dionne, and H. A. Neckles. 2006. A regional assessment of saltmarsh restoration and monitoring in the Gulf of Maine. Restoration Ecology 14: 516–525. Layman, C. A., D. A. Arrington, R. B. Langerhans, and B. R. Silliman. 2004. Degree of fragmentation affects fish assemblage structure in Andros Island (Bahamas) estuaries. Caribbean Journal of Science 40: 232–244. Lesser, C. R., F. J. Murphy, and R. W. Lake. 1976. Some effects of grid system mosquito control ditching on salt marsh biota in Delaware. Mosquito News 36: 69–77. Mendelssohn, I. A. 1979. Nitrogen-metabolism in the height forms of Spartina alterniflora in North Carolina. Ecology 60: 574–584. Mendelssohn, I. A., and J. T. Morris. 2000. Ecophysiological controls on the productivity of Spartina alterniflora Loisel. Pages 59–80 in M. P. Weinstein and D. A. Kreeger (eds.), Concepts and Controversies in Tidal Marsh Ecology. Dordrecht: Kluwer Academic. Miller, W. R., and F. E. Egler. 1950. Vegetation of the Wequetequock-Pawcatuck tidal marshes, Connecticut. Ecological Monographs 20: 143–172. Mitsch, W. J., and J. G. Gosselink. 2000. Wetlands. New York: Wiley. Montague, C. L., A. V. Zale, and H. F. Percival. 1987. Ecological effects of coastal marsh impoundments: A review. Environmental Management 11: 743–756. Montalto, F. A., and T. S. Steenhuis. 2004. The link between hydrology and restoration of tidal marshes in the New York–New Jersey Estuary. Wetlands 24: 414–425. Morgan, P. A., and F. T. Short. 2002. Using functional trajectories to track constructed salt marsh development in the Great Bay estuary, Maine/New Hampshire, U.S.A. Restoration Ecology 10: 461–473. Neckles, H. A., M. Dionne, D. M. Burdick, C. T. Roman, R. Buchsbaum, and E. Hutchins. 2002. A monitoring protocol to assess tidal restoration of salt marshes on local and regional scales. Restoration Ecology 10: 556–563. Nickerson, N. H. 1978. Protection of Massachusetts’ wetlands by order of conditions issued by local Conservation Commissions. Pages 69–76 in Proceedings of National Wetland Protection

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Symposium. Washington, DC: U.S. Department of the Interior. Niering, W. A., and R. M. Bowers. 1966. Our disappearing tidal marshes. Connecticut Arboretum Bulletin 12: 1–36. Nixon, S. W. 1982. The Ecology of New England High Marshes: A Community Profile. FWS/OBS-81/55. Washington, DC: U.S. Fish and Wildlife Service, Office of Biological Services. Pennings, S. C., and R. M. Callaway. 1992. Salt-marsh plant zonation: The relative importance of competition and physical factors. Ecology 73: 681–690. Percy, J. A. 1998. Land-based activities and their physical impacts on marine habitats in the Gulf of Maine. Global Programme of Action Coalition for the Gulf of Maine, Working Paper, May 1998, Commission for Environmental Cooperation, Montreal. Portnoy, J. W. 1999. Salt marsh diking and restoration: Biogeochemical implications of altered wetland hydrology. Environmental Management 24: 111–120. Portnoy, J. W., and A. E. Giblin. 1997a. Biogeochemical effects of seawater restoration to diked salt marshes. Ecological Applications 7: 1054–1063. ———. 1997b. Effects of historic tidal restrictions on salt marsh sediment chemistry. Biogeochemistry 36: 275–303. Provost, M. W. 1977. Source reduction in salt-marsh mosquito control: Past and future. Mosquito News 37: 689–698. Rand, T. A. 1999. Effects of environmental context on the susceptibility of Atriplex patula to attack by herbivorous beetles. Oecologia 121: 39–46. Raposa, K. B., C. T. Roman, and J. F. Heltshe. 2003. Monitoring nekton as a bioindicator in shallow estuarine habitats. Environmental Monitoring and Assessment 81: 239–255. Reed, D. J. 1989. Patterns of sediment deposition in subsiding coastal salt marshes, Terrebonne Bay, Louisiana: The role of winter storms. Estuaries 12: 222–227. Reiley, F. A. 1936. The CCC in mosquito work in southern New Jersey. Proceedings of the New Jersey Mosquito Extermination Association 23: 129–134. Reinert, S. E., F. C. Goley, and W. R. DeRagon. 1981. Avian use of ditched and unditched salt marshes in southeastern New England: A preliminary report. Transactions of the Northeastern Mosquito Control Association 27: 1–23. Resh, V. H. 2001. Mosquito control and habitat modification: Case history studies of San Francisco Bay wetlands. In R. B. Rader, D. P. Batzer, and S. A. Wissinger (eds.), Bioassessment and

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Management of North American Freshwater Wetlands. New York: Wiley. Resh, V. H., and S. S. Balling. 1983. Tidal circulation alteration for salt marsh mosquito control. Environmental Management 7: 79–84. Roman, C. T., W. A. Niering, and R. S. Warren. 1984. Salt marsh vegetation change in response to tidal restriction. Environmental Management 8: 141–150. Roman, C. T., K. B. Raposa, S. C. Adamowicz, M. J. James-Pirri, and J. G. Catena. 2002. Quantifying vegetation and nekton response to tidal restoration of a New England salt marsh. Restoration Ecology 10: 450–460. Rozsa, R. 1995. Human impacts on tidal wetlands: History and regulations. Pages 42–50 in G. D. Dreyer and W. A. Niering (eds.), Tidal Marshes of Long Island Sound: Ecology, History and Restoration. New London: Connecticut College Arboretum. Russel, H. S. 1976. A Long, Deep Furrow: Three Centuries of Farming in New England. Hanover, NH: University Press of New England. Sebold, K. R. 1992. From Marsh to Farm: The Landscape Transformation of Coastal New Jersey. Washington, DC: National Park Service, U.S. Department of the Interior. ———. 1998. The low green prairies of the sea: Economic usage and cultural construction of the Gulf of Maine salt marshes. Unpublished PhD diss., University of Maine, Orono. Shisler, J. K., and D. M. Jobbins. 1977. Salt marsh productivity as affected by the selective ditching technique, open marsh water management. Mosquito News 37: 631–636. Shumway, S. W., and M. D. Bertness. 1994. Patch size effects on marsh plant secondary succession mechanisms. Ecology 75: 564–568. Silliman, B. R., and M. D. Bertness. 2004. Shoreline development drives invasion of Phragmites australis and the loss of plant diversity on New England salt marshes. Conservation Biology 18: 1424–1434. Silliman, B. R., and A. Bortolus. 2003. Underestimation of Spartina productivity in western Atlantic marshes: Marsh invertebrates eat more than just detritus. Oikos 101: 549–554. Silliman, B. R., J. van de Koppel, M. D. Bertness, L. E. Stanton, and I. A. Mendelssohn. 2005. Drought, snails, and large-scale die-off of southern US salt marshes. Science 310: 1803–1806. Silliman, B. R., and J. C. Zieman. 2001. Top-down control of Spartina alterniflora production by periwinkle grazing in a Virginia salt marsh. Ecology 82: 2830–2845.

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Marshes in New Hampshire. Durham, NH: U.S. Department of Agriculture. Valiela, I., J. E. Wright, S. B. Volkmann, and J. M. Teal. 1977. Growth, production and energy transformations in the salt marsh killifish Fundulus heteroclitus (L.). Marine Biology 40: 135–144. Warren, R. S., P. E. Fell, R. Rozsa, A. H. Brawley, A. C. Orsted, E. T. Olson, V. Swamy, and W. A. Niering. 2002. Salt marsh restoration in Connecticut: Twenty years of science and management. Restoration Ecology 10: 497–513. Warren, R. S., and W. A. Niering. 1993. Vegetation change on a Northeast tidal marsh: Interaction of sea-level rise and marsh accretion. Ecology 74: 96–103. Weisberg, S. B., and V. A. Lotrich. 1982. The importance of an infrequently flooded intertidal marsh surface as an energy source for the mummichog Fundulus heteroclitus: An experimental approach. Marine Biology 66: 307–310. Whigham, D. F., J. O’Neill, and M. McWethy. 1983. The Effect of Three Marsh Management Techniques on the Ecology of Irregularly Flooded Chesapeake Bay Wetlands: Parts I and II. Edgewater, MD: Smithsonian Environmental Research Center. Wolfe, R. J. 1996. Effects of open marsh water management on selected tidal marsh resources: A review. Journal of the American Mosquito Control Association 12: 701–712. Zedler, J. B., J. C. Callaway, J. S. Desmond, G. VivianSmith, G. D. Williams, G. Sullivan, A. E. Brewster, and B. K. Bradshaw. 1999. Californian salt-marsh vegetation: An improved model of spatial pattern. Ecosystems 2: 19–35.

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Impacts of Global Climate Change and Sea-Level Rise on Tidal Wetlands J. Court Stevenson and Michael S. Kearney It has been almost fifty years since the carbon dioxide measurements begun at Mauna Loa led to our understanding of the greenhouse effect of warming of the Earth’s surface and sea-level rise. Shortly thereafter, Albert Redfield began coring sediments at Barnstable Marsh using 14C dating to determine it had kept pace with sea level for almost four millennia. Redfield’s early work led to the paradigm that marshes were in equilibrium with sea level. But by the 1970s, it became clear that large amounts of marsh worldwide had disappeared in areas as diverse as Louisiana, Chesapeake Bay, and Venice Lagoon. It is now an open question as to what extent the world’s salt marshes are actually in equilibrium and able to keep abreast of sea-level rise. To be sure, this is a difficult problem to address, even using detailed sea-level histories during the Holocene period. The list of marshes that have deficits in accretion or are affected by lateral erosion process appears to be growing. At the same time, in other instances, high marsh zones are being encroached upon by the low marsh. Moreover, it is highly likely that sulfide accumulation in the root zone from increasing water levels due to rising sea levels will ultimately reduce the potential for vertical accretion. We conclude that, without additional sediment inputs, marshes are forced to accrete more organically to keep up with rising sea level. The highly organic sediments near the marsh surface are more susceptible to oxidation during periodic drought (or when subsurface groundwater inputs are diminished as a result of human activities), leading to ephemeral acidification events, especially after rehydration occurs via acid rain. Managers need to be especially careful that when creating impoundments, marsh surface sediments do not dry out, shrink, and oxidize. By the same token, every effort should be made to reduce fire (an oxidation process) on highly organic marshes. Additional steps that can be taken to improve marsh survival include reducing overgrazing of marshes by geese, muskrats, nutria, and snails. Ultimately, marshes can survive high upward excursions of sea-level rise if mineral sediment inputs are enhanced. Studies have documented that jet-spraying material on the marsh surface is an effective, if possibly a limited-scale, solution. In addition, new marshes can be created using dredged materials to offset areas where losses are unavoidable, though large projects may cost billions of dollars. Over the long run, stabilization of greenhouse gas emissions may be the most 171

cost-effective way to save marshes at risk because of increasing sea-level rise. We suggest that as much as 90 percent of the tidal marshes worldwide could be in jeopardy by 2100—if no action is taken to curtail greenhouse gas emissions and population growth.

Humans have invaded and exploited almost every available habitat on the planet to sustain an enormous population now estimated at over 6.5 billion people by the U.S. Census Bureau. Assuming a land area of 148.85 million square kilometers (Weast 1981), this yields a present density of almost forty-four people per square kilometer. While the 12.76 million square kilometers of the Earth’s surface now classified as wetlands under the RAMSAR definition (Finlayson et al. 1999) have not been the main focus of development attention, they have not been overlooked, either. Many wetlands have been drained, diked, burned, filled, used for hunting and grazing of animals, and otherwise altered to satisfy growing anthropogenic needs for food, shelter, recreation, and transportation corridors (Chapman 1960; Stevenson et al. 1977; Roman, Niering, and Warren et al. 1984; Stevenson et al. 1999; Allen 2000; Isacch et al. 2004; Geslin, Eybert, and Radureau 2006). In some areas such as Europe, an estimated 90 percent of the wetlands have been simply converted to other uses, compared to slightly over 50 percent in the United States and Australia (Mitsch and Gosselink 2000). Salt

FIGURE 10.1 The marshes surrounding Barnstable Harbor on Cape Cod in Massachusetts early in the year appear healthy, with substantial wrack deposits in the foreground. Photo by J. C. Stevenson, May 2007.

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marshes have been especially affected by urbanization and other anthropogenic factors that concentrate growth in the coastal zone. Recently, Bromberg and Bertness (2005) have estimated from historical maps that 37 percent of the salt marshes have been lost in New England, mostly in and around the growing cities of Boston and Providence (fig. 10.1). Beyond physical damage to a wide variety of wetlands, there are many significant indirect (and often difficult-to-assess) consequences of human activities. One of the potentially problematic anthropogenic alterations associated with modern industrialization is the dramatic increase in carbon dioxide in the atmosphere (fig. 10.2), which has risen from approximately 280 parts per million in 1750 to over 380 parts per million at present (Körner 2006). This CO2 rise results in a myriad of global environmental changes, ranging from higher air and water temperatures to accelerating rates of sealevel rise (Houghton et al. 2001; Church and White 2006). Over the long term, indirect impacts of raising CO2 could have far-reaching implications for marsh survival through a variety of mechanisms. One of the prime concerns is that in response to

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global warming, glaciers in Greenland and Antarctica will melt. If they both were completely deglaciated, global sea level would rise by about sixty-five meters (Cazenave 2006). Currently, the Greenland ice sheet appears to be melting at a rate of 113 gigatons per year (Luthcke et al. 2006), with both Greenland and Antarctica contributing a combined total of about 125 gigatons of water per year to the global ocean, adding significantly to its mass (Shepherd and Wingham 2007). Although there are still uncertainties (due mostly to only a few years of data collection), most recent analysis (Lombard et al. 2007) of satellite altimetry and gravity data suggests that the in-

FIGURE 10.2 Mean monthly change in CO2 at Mauna Loa since 1958. From NOAA.

crease in oceanic mass results in 1.2-millimeter higher sea level per year, while thermal (i.e., steric) expansion accounts for an additional rise of about 1.9 millimeters per year. Nicholls, Hoozemans, and Marchand 1999 and Nicholls (2004) had previously estimated a 5 to 20 percent wetland loss by 2080 along the coasts due to sea-level rise. However, these are sea-level rise rates that some consider too low (Rahmstorf 2007; Rahmstorf et al. 2007). Also of concern, Nicholls (2004) did not take into account feedback processes that may allow some marshes to accrete enough material to keep abreast of rising sea level (fig. 10.3). Vertical

Sediment supply Elevation

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Allochthonous sedimentation Hydroperiod

Root zone Soil volume

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Consolidated basement FIGURE 10.3 Principal factors governing marsh elevation and adjustment to sea level (from French 2006).

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accretion can be rapid enough to keep abreast where plant canopies provide enough drag to capture incoming sediments that can be incorporated into a robust rhizomatous matrix capable of resisting erosion when waves action is significant (McCaffrey and Thomson 1980; Bricker-Urso et al. 1989; Friedrichs and Perry 2001). If marsh vegetation remains healthy, canopy density, height, and flexibility can often interact to dissipate energy and capture additional sediment potentially available during excursions of sea-level rise (Leonard and Luther 1995; Nepf 1999). Unfortunately, Nichols et al. 1999 and Nicholls (2004) did not address which tidal marshes were most likely to fall behind rising sea levels. Another mechanism that is often neglected in considering global change impacts on marshes is that enhanced levels of CO2 in the atmosphere has little effect on C4 plants, but most C3 plants increase their productivity (Körner 2006), which could offset some of the negative effects suffered by marsh plants when sea level rises. For example, Curtis et al. (1990) found that doubling atmospheric CO2 levels resulted in an 83 percent increase in the belowground biomass of a C3 sedge species, Olney’s three square (Schoenoplectus americanus). Increased biomass may help C3 species lay down more peat, keeping them abreast of rising sea levels. However, there are other complications in this evolving story. Recently, Marsh et al. (2005) reported that when Schoenoplectus was exposed to two times ambient CO2 concentrations, more dissolved organic carbon (DOC) is available for export out of the root zone. While greater DOC export may supplement production for the estuarine plankton community downstream, Marsh et al. (2005) also noted increased production of methane from the marsh dominated by C3 species when exposed to elevated CO2. Methane is twenty times more potent a greenhouse gas than CO2 (Matthews and Fung 1987), ultimately making feedback loops difficult to predict. However, it would appear that C3 species more typical of the high marsh should have an in174

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creasingly competitive edge than C4 species such as Spartina alterniflora, which dominates the low marsh. In addition to CO2, increased temperatures themselves may be key in influencing not only species interactions in coastal wetlands but community interactions as well. The boundaries between mangroves and salt marshes are significant ecotones where species compete for dominance (Clarke and Hannon 1969). Mangroves are restricted to tropical and subtropical shorelines because their evergreen leaves cannot tolerate freezing conditions very long (Chapman 1976). Steadily increasing temperatures may be the most plausible explanation for the recent loss of salt marshes in favor of mangrove vegetation (i.e., Avicennia marina) along the east coast of Australia (Saintilan and Williams 1999). However, Rogers, Saintlan, and Heijns (2005) do not discount the impact of local anthropogenic changes that lower water tables (increasing the potential for compaction of sediments) in giving mangroves the edge over salt marshes in Western Port Bay (38⬚21⬘ S, 145⬚13⬘ E), which is at the southern limit of survival for A. marina. Indeed, studies at the northern part of the range of Avicennia germinans in the United States, where it can be found with Spartina alterniflora, suggest that the competition between the two may be complicated by biotic interactions as well as temperature. Using transplants at the edge of Bay Champagne in Louisiana and greenhouse experiments, Patterson et al. (1993) have hypothesized that Avicennia may be physiochemically inhibited by S. alterniflora when in direct competition, especially under more flooded conditions. This inhibition likely involves hydrogen sulfide (H2S) in the root zone, which S. alterniflora tolerates more readily than A. germinans (Patterson and Mendelssohn 1991). Much more needs to be known about these kinds of details if the scientific community hopes to make credible projections necessary for future planning. Twenty years ago in a series of papers (Stevenson, Kearney, and Pendleton 1985; Stevenson, Ward, et al. 1985; Stevenson, Ward,

and Kearney 1986, 1988), we speculated that sea-level rise could have increasingly negative impacts on tidal marshes, forcing some of them to be increasingly stressed and ultimately submerged. At the time, Orson, Panageto, and Leatherman (1985) had come to similar conclusions. Despite early reports of marsh erosion in South Carolina (Settlemyer and Gardner 1975), most marsh researchers were not focused on this issue, because marsh deterioration had only been well documented in a small number of tidal marshes in Chesapeake Bay and Louisiana. In 1983, Delaune, Baumann, and Gosselink reported that at least one major marsh system in the Mississippi Delta was not keeping abreast of sea level, but overall marsh losses in the region were seen more as a result of deltaic subsidence and sediment starvation rather than global (i.e., eustatic) sea-level rise (Boesch et al. 1982). This view was developed by Morgan and his colleagues in Louisiana (Coleman, Roberts, and Stone 1998) and still is thought to be instrumental in driving wetland loss in the Mississippi Delta (Reed 1990, 1995, 2002). Recently Tornqvist et al. (2004) have questioned the dating of lobe formation and subsidence patterns in the late Holocene history of the Mississippi Delta. But subsidence on clearly more local and shorter temporal scales (from oil and gas development) remains critical to understanding present marsh loss in the area, as Morton, Tiling, and Ferina (2003) have documented. Turner (1997, 2004) has consistently underscored the importance of the effects of hydrological changes on wetlands due to dredging of canals and creation of spoil banks (to facilitate oil and gas extraction) in the Mississippi Delta, which account for 0.84 percent of marsh loss per year. Nevertheless, the hypothesis Stevenson et al. (1986) suggested was that marshes that were accreting mostly by in situ organic deposition (because they had low inorganic sediment inputs or low tidal energies) were most at risk to rising sea levels. Furthermore, the most at risk marshes were located where tides were ebb dominated and in areas where subsidence

was high due to negative isostatic readjustment following glaciation (i.e., forebuldge collapse), excessive groundwater removals, and sedimentstarved deltaic systems where sediment compaction is high (Stevenson et al. 1988). A reassessment of the impacts of global change on coastal marshes is perhaps now warranted, because recent studies confirm an acceleration in sea-level rise (Church and White 2006; Gehrels et al. 2006), and new tools and studies are available. Thomas and Ridd (2004) have reviewed the latest developments in measuring sediment accumulation over short time scales. These methods are numerous and include simple approaches such as deployment of anchored tiles (Pasternack and Brush 1998) and anchored rubber pads (Rooth and Stevenson 2000) to elaborate deployment of sedimentation erosion tables (SETs; Boumans and Day 1993) or, as they are now called, surface elevation tables (Cahoon et al. 2002). In addition, considerable strides have been made in quantifying flow regimes and wave climates in tidal marsh canopies, which can account for sedimentation patterns (Leonard and Luther 1995; Möller et al. 1999; Neumeier and Ciavola 2004). Furthermore, Kearney, Grace, and Stevenson (1988) have pioneered the use of remote sensing to detect marsh deterioration from aerial photography. A particularly useful remote sensing technique that complements tide gauge measurements is satellite altimetry, which has been operational on a global basis since 1993 (Nerem and Mitchum 2001). The most recent synthesis of the Poseidon/Topex and Jason satellite data indicates eustatic sea level currently rising about 2.9 millimeters per year (fig. 10.4). This appears to be 50 percent higher rate than that estimated using tide gauge data over the last one hundred years, 1.8 to 2.0 millimeters per year (Douglas, Kearney, and Leatherman 2001). Nevertheless, greenhouse gas emissions continue to increase, and there is a growing concern that global temperatures eventually will reach what they were 130,000 years ago when the Greenland ice cap melted, releasing enough water to raise sea levels more than six meters (Overpeck et al.

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30

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TOPEX Jason 60-day smoothing Inverted barometer applied

10

0

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Univ of Colorado 2005_rel5

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1996

1998

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FIGURE 10.4 Global sea-level changes from 1993 to 2005 from Topex/Posiedon and Jason Satellite altimetry.

2006). If the existing ice sheets in Greenland and Antarctica (Cook et al. 2005) were to undergo a catastrophic collapse within less than a century, most tidal wetlands around the world could simply drown. It is equally certain that even with rates of global sea-level rise comparable to those that followed waning of the Wisconsin ice sheets after the last Glacial Maximum (Clark et al. 2004), marsh survival would be extremely limited. The first question we address here is whether marshes were able to flourish, even periodically, after global sea-level rise began to slow during the early Holocene (about one centimeter per year) despite the possibility of renewed acceleration during the mid-Holocene warm period (i.e., Hypsithermal), about five to eight thousand years ago (see Kearney 2001). Then we review the Redfield (1965) and Lucke (1934) models of marsh and lagoonal fill evolution, followed by a sampling of coastal marshes where recent studies show that declining wetlands are not restricted to Chesapeake Bay and 176

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Louisiana but are now increasingly documented throughout the mid-Atlantic and at several other locations in the United States and beyond. Thereafter, we suggest what management strategies might be employed to ameliorate the impacts of sea-level rise.

GEOLOGICAL BACKGROUND: THE HOLOCENE Orson (1996) emphasized that paleoecological research can offer insights into the past and allow us to more accurately model and predict future responses of marsh systems to sea-level rise. One of the best opportunities for tracing the long-term impact of sea-level rise on coastal wetlands lies on the seabed of passive trailing edge continental margins where tectonic activity is not pronounced (because there are no plate boundaries). Shackleton’s ␦O18 data (1987), which by inferring global ice volumes remains the best way to avoid the regional and local effects that characterize traditional sea-level curves

Leading edge of the holocene transgression

Wave erosion

Duneswind

Beach & berm erosion & deposition Atlantic Ocean

ar al m

Tid

sh

Overwash Landward Transport

Lagoonal mud

Pr

e-t 2780 BP ran 5640 BP sg res sio n Plaislocana

Feet

Meters

0

e op Sl sh Wa

0

Tidal Stream

6345 BP

Sand B Gravel

Shallow Marine

7075 BP

Su

rfa

24

Mean low sea level

ce

8325 BP

80 0

Miles

10

0

Kilometers

16

FIGURE 10.5 Ravinement surface (Kraft 1987).

(Kearney 2001), suggest overall Holocene sealevel rise was high enough (about one centimeter per year) to drown marshes that formed at maximum glaciation. As rates of sea-level rise slowed, there is evidence of eleven-thousandyear-old marshes in water fifty-nine meters deep off Georges Bank (Emery et al. 1965). Farther south, Rampino and Sanders (1981) reported several more old marsh deposits on the inner shelf of the mid-Atlantic coast, leading them to conclude that marsh development was episodic during the Holocene and was correlated with cooling periods (when sea-level oscillations were negative). The fact that very old marsh deposits are so scarce on the continental shelf of the Atlantic Coast, however, may not be because the sea was rising too rapidly for them to form, but because those marshes that did get established were simply eroded. During rapid transgressions, erosion undoubtedly played some role in the paucity of marsh depositional records older than 5,000 years BP. The extant studies of the inner continental shelf show that the occurrence of late Pleistocene marsh peat in transgressive sequences can be scarce where a combination of high wave energy and relatively slower rates of

sea-level rise limit preservation potential (Belknap and Kraft 1981; Field and Duane 1976). Antecedent stream valleys are often the only places on the inner shelf of the Delaware coast that contain Holocene and late Pleistocene marsh peat (e.g., Kraft et al. 1987). The ravinement surface (i.e., ravine-like) is covered by just a thin cap of sands on former interfluves (fig. 10.5). Seaward of headlands no marsh peats of Holocene or late Pleistocene age have been found. Similarly, there are no preHolocene marsh sediments on the shelf eastward of the Virginia barrier islands, and only a few mid-Wisconsin peats have been dated (Finkelstein and Kearney 1988; Oertel, Kearney, and Woo 1992). Furthermore, Riggs et al. (1992), using amino racemization dates on mollusk shells, demonstrated that marsh peats are largely absent on the inner shelf and shoreface of the North Carolina coast, which also consists of a complex of ravinement surfaces and earlier truncated Quaternary sequences. The predominantly bedrock substrate on the inner shelf in the Gulf of Maine is also another factor that limited long-term tidal marsh establishment at high latitudes along the U.S. shelf— until the late Holocene. Also complicating

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marsh formation was that isostatic depression of the crust resulted in an inland intrusion and marine conditions, which subsequently reemerged and is now fastland along the Maine coast (Scott, Medioli, and Miller 1987; Kelley et al. 1992). Since the inner shelf of the Gulf of Maine was predominantly bedrock as a result of glacial scouring, tidal marsh establishment was limited there until the late Holocene. Marsh formation was also complicated by isostatic depression of the crust, which resulted in a transgression of the sea inland of the present shoreline of the Maine coast (Scott et al. 1987; Kelley et al. 1992). During the middle to late Holocene periods, however, sea levels appear to have risen slowly (one to two millimeters per year), enough for marsh survival along the Gulf of Maine and the rest of the mid-Atlantic shelf to support marsh growth (Belknap and Kraft 1977). Donnelly et al. (2004) have concluded that sea levels in Connecticut rose in the range of one millimeter per year until about 1850, after which they rose at nearly a threefold rate. At present, there is not enough evidence to decide whether the absence of coastal marshes from late Pleistocene and early Holocene times was due to erosion during transgression or from sea-level rise outpacing marsh development—though the rapid loss of coastal marshes in Louisiana and the middle Atlantic Coast in the twentieth century makes a persuasive case for rapid sea-level rise. In environments that are more favorable to marsh growth and accretion, where wave action is not severe, marsh survival depends on several factors. For example, the model developed by Morris et al. (2002) suggests that marshes in the mesotidal coast of South Carolina, where tidal flushing is robust, can accommodate rates of sea-level rise of a meter per century. This prediction certainly fits the late Holocene sea-level record of South Carolina, in which basal marsh peat data show marshes adjusting to sea-level fluctuations of over one meter (Colquhoun and Brooks 1986). The study of Morris et al. (2005) shows that S. alterniflora marshes in the North Inlet estuary are presently maintaining elevation 178

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at the low end of the tidal frame. According to the Morris et al. (2002) model, vertical accretion rates should increase and reestablish with the rate of sea-level rise, with marsh surface elevations rising to within the mean elevation of the tidal frame. It remains to be seen whether the North Inlet marshes can adjust upwardly if sealevel rise continues to accelerate. Kirwan and Murray (2007) have proposed a related tidal morphodynamics model whereby a “metastable” equilibrium is established under a constant but low rate of sea-level rise. The sealevel trend of the past half-century for the U.S. mid-Atlantic coast reveals that the notion of a constant rate of sea-level rise is a chimera, with considerable decadal and interannual variability. At least three major pulses in rate of one or more centimeters per year have occurred since 1966, the last spanning most of the 1990s. The widespread “sudden dieback” being reported from many areas of the U.S. Atlantic Coast suggest that physiological “disturbance” of the vegetation has precluded any simple recovery. Although the most recent marsh geomorphic evolution models (Kirwan and Murray 2007; D’Alpaos et al. 2007) have attempted to characterize the interplay between hydrodynamics and the development of the marsh platform, they still do not incorporate the interaction of other factors such as high-energy events (e.g., hurricanes), edaphic waterlogging (leading to hydrogen sulfide production) associated with increasing submergence, and other biological processes (e.g., grazing), which ultimately determine the stability of salt marsh ecosystems in a given environment. Clearly, more elaborate models will be necessary to provide a realistic assessment of how well marshes will survive future sea-level rise. Rahmstorf’s (2007) estimate of up to a 1.4-meter sea-level rise by 2100 (if we do nothing to address carbon emissions) would present serious problems for most tidal marsh survival, especially for marshes where sediment inputs are limited. While relative sea-level rise is a key driver in marsh survival, the presence or absence of sediment inputs from the upland or seaward edges

FIGURE 10.6 Marsh die-off at Wellfleet Bay Wildlife Sanctuary on Cape Cod, forty kilometers northeast of Barnstable Harbor. This photo was taken at the same time as figure 10.1 and normally would have substantial growth of marsh at this time of year. Photo by J. C. Stevenson, May 2007.

of the marsh is also critical. As the glaciers came and went, much of the land was left bare of vegetation, and loess materials were deposited in and around many river systems, ending up on the broad shelves. There is no question that if robust sediment supplies are available, marshes can accrete well beyond one centimeter per year. Yang (1998, 1999) found that sediment accretion in Scirpus marshes on the Jiuaduansha Shoal at the mouth of the Yangtze River in China ranged from 4 to 14 centimeters (mean ⫽ 7.9 centimeters) during 102 days in May through August in 1997. We imagine that during the Holocene, areas of high sediment input at the mouths of muddy rivers were undoubtedly the locations where marshes were able to survive the most extreme transgressive periods (when sea-level rise could have reached one to two centimeters cm per year).

MODELS OF MARSH FORMATION AND EQUILIBRIUM WITH RISING SEA LEVEL Redfield (1965, 1972) concluded that Barnstable Marsh formed behind a prograding spit and accreted by accumulation of organic and mineral materials in equilibrium with sea-level rise for at least four thousand years. Nonetheless, since Redfield’s work, more detailed information has become available on Holocene sea-level changes, marsh sedimentation, hydrography, and marsh

plant physiology, suggesting that Barnstable Marsh (fig. 10.1) should now be regarded as typical only when ample sediment supplies are available for development. Documentation of losses of coastal marshes along the mid-Atlantic coast in the late twentieth century—in some instances without evidence of other human intervention (Kearney et al. 2002)—undercuts notions of marsh equilibrium (Friedrichs and Perry 2001). Increasing reports (http:// wetlands.neers.org, www.inlandbays.org, www. brownmarsh.net, www.lacoast.gov/watermarks/ 2004-04/3crms/index.htm) of widespread “sudden marsh dieback” and “brown marsh dieback” from Maine to Louisiana e.g. Well fleet Bay (fig 10.6), along with published studies documenting losses of marshes dominated by S. alterniflora (as well as other halophytes), suggest that many are approaching or have actually gone beyond their “tipping point” where they can continue to accrete enough material to survive (Delaune et al. 1983; Stevenson, Kearney, et al. 1985; Stevenson, Ward, et al. 1985; Kearney et al. 1988; Mendelssohn and McKee 1988; Kearney et al. 1994; Hartig et al. 2002; McKee, Mendelssohn, and Materne 2004; Turner 2004). While Friedrichs and Perry (2001) acknowledge that sea level may overwhelm marsh systems, forcing them to become mudflats, they view this as being in sedimentary equilibrium. We maintain that this usage stretches the original “marsh sea-level

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equilibrium concept” advanced by Redfield (1972) beyond its original connotation and is little comfort for those concerned with marsh survival in a global environment where sea-level rise is accelerating (Church and White 2006). One particular factor that calls for a reexamination of Redfield’s conclusion regarding marsh development and sea-level rise concerns the revised thinking on the history of the late Holocene sea-level change. When Redfield’s original work was published (Redfield and Rubin 1962), the concept of an oscillatory sea level was very controversial—and remains so. Fairbridge’s (1961) famous global reconstruction for Holocene sea-level history showed oscillations, though the first curves for New England showed a smooth decelerating trend. The picture of late Holocene climate history has become less sketchy than it was when Flint (1971) published his seminal text on Quaternary and glacial geology. Moreover, based on work like that of Colquhoun and Brooks (1986), it has become passé to assume the upward trend in sea level during the late Holocene was modest and constant. If anything, the complexity of the late Holocene climatic record (Grove 1988) would argue against such a notion, and the evidence for equally complex sea-level variations, especially within the last thousand years (Varekamp et al. 1992; Kearney 1996; Gehrels et al. 2006), has become especially persuasive. Periods of regression, followed by episodes of transgression, appear to characterize the sea-level record since the late twelfth century (Varekamp et al. 1992). This is consistent with global temperature change after the Late Middle Ages Warm Period through the severe cold phases of the Little Ice Age in the late sixteenth and seventeenth centuries to the present warming trend after about 1820–1850, which most climate researchers now attribute to the burning of fossil fuels (Houghton et al. 2001). The finer-scale reconstructions of the sealevel history of the last millennium, combined with more detailed studies of the marsh vertical accretion covering most or all of this period (see Kearney et al. 1994), raise questions as to 180

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whether equilibrium is a condition that marshes achieve early on and maintain, as Redfield implied, or is at most a temporary circumstance, essentially becoming nonexistent with substantial transgressions or regressions. In the first instance, Kearney et al. (1994) showed that marsh vertical accretion rates in the eighteenth century at Monie Bay largely track these sea-level changes, but beginning in the late nineteenth century, though exhibiting a sharp upward trend in accretion, these rates actually fell behind the pace of coastal submergence. Before the eighteenth century, details on marsh development in many estuaries such as Chesapeake Bay are still largely conjectural. The depth of the regression during the Little Ice Age can be inferred by Varekamp et al. (1992); it is hard to envision any other marsh response other than oxidation of substrates (and perhaps older peats depending on the depth of the regression) as sea levels dropped. This is definitely not an equilibrium scenario. In general, based on Gehrels’s (1999) study data from Machiasport, Maine, and Iceland (Gehrels et al. 2006), it is possible to conceptualize marsh development over the last century, at least at high latitudes. Depending on the depth of regressions that probably occurred during the coldest phases of the Little Ice Age (e.g., during the early and late seventeenth century; Grove 1988), either one of two scenarios could apply. Hypothetically, prolonged significant regressions where there was a substantial drop in mean tide level should result in substantial oxidation and erosional unconformities. Yet, no evidence of this appears to have been published. Eventually, new marsh would have developed on a remnant surface left by peat decay (recurrence surface), which coincides with the present transgression beginning around 1850 (Kearney 1996). On the other hand, if the regressions were relatively shallow, of limited duration, and possibly amounting only to a still stand, loss of peat would be limited. Perhaps only humification of the marsh substrate would be obvious if the peat was exposed to aerobic conditions for longer time periods.

140 120

Sea level change (cm)

100 80 60 40 20 0 –20 1900

1950

2000

2050

2100

Year FIGURE 10.7 Sea-level projections to 2100 (Rahmstorf 2007) using several CO2 emission scenarios in the Third IPPC assessment (Houghton et al. 2001). For example, the A1FI scenario (the uppermost curve within the gray area) reaches more than 1 meter by 2100, which postulates rapid economic growth with global population peaking at midcentury where fossil intensive energy sources are utilized; whereas the B1 scenario (the bottom-most curve within the gray area) reaches about 0.7 meter by 2100, which postulates the same population growth as A1FI but assumes a global shift toward an information economy with reduction in material intensity and shift toward clean-resource, efficient energy sources.

Unfortunately, few studies have addressed the chemical composition of peat in enough detail to fully understand what may have occurred during past sea-level recessions. Ward, Kearney, and Stevenson (1998) did show changes in percent organics of marsh cores in Monie Bay on Maryland’s Eastern Shore that penetrated to the basal estuarine tidal flat clays, and for which there is no discernible hiatus in peat deposition. Unfortunately, no basal peat dates for the cores were obtained by Ward et al. (1998). It is possible that the recessions were too shallow and, hence, the marsh too young to record the episodes of sea-level regression that likely occurred in the sixteenth and seventeenth centuries. Though present concern about the fate of tidal marshes focuses on whether they will survive an acceleration of global sea-level rise, improved understanding of peat preservation might gives us more clues about coastal marsh responses to future sea levels. In addition to

field and modeling studies, simulations could be carried out using marsh mesocosms where water levels could be manipulated over time and marsh responses could be observed under a variety of sea-level scenarios. The most realistic projections (fig. 10.7) are those recently published by Rahmstorf (2007).

TIDAL ASYMMETRY AND ITS IMPACTS ON VERTICAL ACCRETION Building on the model of Lucke (1934), Stevenson et al. (1988) suggested that tidal velocities on ebb and flood were very influential in determining whether marshes import or export suspended material at their mouth and help keep abreast of sea level. Most East Coast marshes south of New York appear to be “ebb dominated”; that is, they have peak current velocities during ebb tides, which promote the export of suspended materials (Stevenson et al. 1988). On the

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181

contrary, those located in New England appear to be flood dominated, the classic being Nauset Inlet on Cape Cod (Aubrey and Speer 1985). Ebb domination in San Francisco Bay has also been reported by Leopold, Collins, and Collins (1993), Pestrong (1965), and Fagherazzi, Gabet, and Furbish (2004). Ebb domination occurs because the water surface slope from the top of the marsh surface to the mouth is greater at low tide than at high tide. At high tide, water in the creek system is minimally above the marsh surface; and when the tide turns, the drainage of water off the surface is delayed until water elevations in the creeks have dropped substantially, which increases velocities during the ebb tide (fig. 10.8). The difference in water surface slope essentially makes the ebb more like a free-flowing stream. Although not addressing marshes directly, Fitzgerald and Nummedal (1983) also found that the lag between ocean levels and those in the estuary (or tidal creeks) tends to promote high ebb velocities. What has drainage dynamics to do with marsh vertical accretion and sea-level rise? Differences in drainage dynamics influence the potential influx of substantial volumes of allochthonous (organic and mineral) sediment. What sediment does finally reach the headwaters of small, first-order tidal creeks is likely to be very fine grained (i.e., clay) and slow to settle out of suspension during the time the marsh surface is flooded. The prevalence of ebb domination throughout the world is not known. Nevertheless, it may be critical where tidal energies are low, such as in Chesapeake Bay (Stevenson, Ward, et al. 1985) and San Francisco Bay (e.g., Fagherazzi et al. 2004). Thus, it is little wonder that microtidal marsh systems are among the most vulnerable to acceleration in sea level, constrained as they are by their hydrographic dynamics to maintaining marsh surface elevation largely by organic accumulation. This may account for the extensive marsh losses noted early on in the Mississippi Delta (Boesch et al. 1982) and in the Blackwater Wildlife Refuge (Stevenson, Kearney, et al. 1985) where extremely microtidal conditions prevail. 182

land use and climate change

Subsequent studies in the Blackwater River have shown that the diminution of energy as the tidal wave moves upstream is dramatic. Stevenson et al. (2000) presented data to show severe truncation of tides in the center of Blackwater Refuge (which is now largely an open water lake where an extensive marsh existed around 1900). Recent data collected by the U.S. National Ocean Survey show that the mean tidal range at McCready’s Creek in Fishing Bay, where the Blackwater River debouches into the Chesapeake, is 0.6 meter; and by the time flood waters reach the Route 335 bridge, the mean tidal range plummets to 0.02 meter (A. Allen, National Oceanic and Atmospheric Administration, Silver Spring, MD, personal communication). Perhaps contributing as much as any other factor to this precipitous drop in tidal range has been dramatic losses of interior marsh (Stevenson, Kearney, et al. 1985). Since 1938, the year of the earliest aerial photography for the region, open water areas have become five to ten times larger at Blackwater Refuge, with a loss of 1,500 hectares of marsh (Rizzo 1995; fig. 10.9). The conversion of marsh to open water, which has been deepening over the last several decades, effectively increases the volume in the center of the marsh system and dilutes tidal energies. Unfortunately for the marsh vegetation, low tidal energy does not allow the flushing of toxic sulfides out of the system. Furthermore, microtidal estuarine marshes are potentially vulnerable to be continually submerged for days, or even weeks, when estuarine headwaters go hypoxic in early summer (Stevenson et al. 2000), limiting reaeration of the root zone and ultimately diminishing plant productivity (Mendelssohn, McKee, and Postek 1982; Mendelssohn and Morris 2000). Blackwater Refuge is one of the few large marsh systems where a sediment budget has been attempted, and this shows a large sedimentary imbalance. Ebb velocities are consistently higher than flood, and that disparity becomes more pronounced downstream (table 10.1). As a result, significantly more

Elev. above mean tide level

Overall water slope

Distance from marsh shoreline

Initial development of marsh drainage exploiting antecedent surface topography of tidal flat

Development of first tributary channels: increasing tidal prism with increasing flood energy loss from greater channel roughness and change in direction

Elev. above mean tide level

(A)

Overall water slope

Distance from marsh shoreline

Further in flood energy from great tidal prism, channel roughness, changes in direction, and denser vegetation

Elev. above mean tide level

(B)

Overall water slope

Distance from marsh shoreline

Further accelerated increase in tidal prism with the formation of interior ponds

(D)

Elev. above mean tide level

(C)

Overall water slope

Distance from marsh shoreline

FIGURE 10.8 Model of the development of marsh hydrography from evolution in marsh tidal creeks, interior ponds, and increasing water slope as they affect tidal prism and velocities.

FIGURE 10.9 A pond formed over the last two decades along the Blackwater River in Dorchester County that had been previously marsh. Photo by J. C. Stevenson, June 2002.

suspended sediment is exported than imported, with a net annual export of 7.2 ⫻ 108 kg yr⫺1 (Stevenson, Kearney, et al. 1985). The impact of this ebb-dominated annual tidal velocity and sediment export is the reliance of most interior marshes almost exclusively on organic accumulation for vertical accretion, which readily can be overwhelmed by strong upward excursions in relative sea level. In addition to regular tidal activities, storm events can be significant in the sediment budgets of wetlands. Baumann et al. (1984) and Turner et al. (2004) have stressed the importance of hurricanes in adding sediments to marshes of the Mississippi Delta. However, hurricane surge can also be very destructive to tidal marsh systems, causing severe erosion (Van de Plassche et al. 2006). If flood tide velocities in marsh creeks exceed ebb velocities (as appears to be the case in many macrotidal environments) and there is a rich downstream sediment supply, marshes can often keep up with rising sea levels. In contrast

to declining marshes at Blackwater, marsh development is rapid where tidal energies are greater and flood tides are dominant in Jiangsu Province in China. In a study of four creeks in Jiangsu Province off the East China Sea, where the tidal range is 3.9 to 5.5 meters, Wang, Zhang, and Gao (1999) have documented tidal velocities on the flood of 1 m s⫺1, or twice as high as on the ebb. Thus, fine-grained sediments in the water column are transported landward in surges during normal tidal cycles to the interior of the marsh. and annual accretion rates of four to eight centimeters have been measured (Zhang 1995). Perhaps not surprisingly, the Jiangsu marshes have prograded at a rate of twenty to thirty meters per year from 1954 to 1988 and do not appear to be adversely affected by rising sea levels along this portion of the Yellow Sea coast (Wang et al. 1999). The progradation of marsh has been going on for centuries but appears to be facilitated recently by the introduction of S. alterniflora in 1979. Wang et al. (2006) have recently reported that 137Cs and 210 Pb dating indicate that the mean surface elevation of S. alterniflora marshes has increased along the Jiangsu Coast in the range of 5.8 centimeters per year over the last half century. Ultimately, the luxuriant sediment supplies from the Yangtze and Yellow rivers have allowed about 2.5 million hectares of Jiangsu Province tideland to be reclaimed since the eleventh century (Zhang et al. 2004). Nevertheless, there is now high probability that marshes along this coastline will be adversely affected by another anthropogenic change over the next several decades: the massive Three Gorges Dam on the Yangtze River, which even during construction

TABLE 10.1 Tidal velocities (cm sec⫺1 ) in the Blackwater River systems as determined from averaged monthly measurements in 1980

Location

Kilometers Upstream

Flood Mean

Ebb Mean

12.0 25.5 26.5

25.0 10.2 3.7

34.0 11.9 7.8

Shorter’s Wharf Bridge Little Blackwater Bridge Upper Blackwater Bridge SOURCE: From Pendleton and Stevenson (1983).

184

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has reduced sediment loading from 466 million tons per year in the 1970s to 394 million tons per year in the late 1990s (Yang et al. 2003). Previously, the Yangtze River contributed about 95 percent of the sediment discharge going into the East China Sea, but Yang et al. (2007) predict that this will be reduced drastically as they expect sediment delivery of the Yangtze will continue to decline (but should not fall below one hundred million tons per year). The human impacts of sediment sequestration in reservoirs behind dams, at a time when sea level is accelerating, could have broad ecological and economic repercussions for this and other shorelines where particulate inputs were traditionally high, such as happened in the Nile Delta (Stanley and Warne 1998). We fear that many tidal marsh systems will decline as a result of sediment starvation, since Sylvitski et al. (2005) have recently calculated that there has been a reduction in the delivery of riverine sediments to the world’s coasts by about 1.4 billion metric tons per year.

DECLINING WETLANDS IN THE MID-ATLANTIC REGION Marshes along the Blackwater River are the most obvious examples of marsh loss in Chesapeake Bay; however, also evident are extensive marsh degradation and losses in the lower Nanticoke estuary downstream of the town of Vienna, Maryland (Kearney et al. 1988). Comparison of historical maps and aerial photographs document considerable expansion in tidal networks of mesohaline and oligohaline marshes of the Nanticoke River of Chesapeake Bay since the early twentieth century, with stream orders (using the Shreve system) of the tidal channels draining the marshes rising from four to more than twelve in many instances (Kearney et al. 1988). Channel widths of larger tidal creeks also substantially increased, with lowering of the creek bank slopes in many cases. These observed changes in tidal drainage coincided with a dramatic growth in the number and size of interior ponds (first creeks acting often as

loci for pond formation) and with vertical accretion rates that were markedly below the relative sea-level trend. In contrast, upstream marshes in the sediment trap reaches of the Nanticoke River (north of Vienna, Maryland) show no signs of deterioration, although rising sea levels, by increasing tidal prism, had increased the width of channels and the number of first-order creeks (Kearney et al. 1988). These upstream marshes were largely of the “estuarine meander” type adjacent to the channel where sediment and nutrient inputs are comparatively high due to agricultural inputs and where salt penetration causes flocculation. The estuary has a considerable amount of sediment resuspension from the bottom, which is transported over the marsh surface on spring tides. This general landscape pattern of healthy marshes upstream and deterioration downstream has also been observed in the adjacent Choptank River watershed. Upstream marshes in the meander portions of the tidal freshwater and oligohaline zones of the Choptank are keeping abreast of sea level, especially in areas where Phragmites australis has invaded. Rooth, Stevenson, and Cornwell (2003) found accretion rates varying from 4 to 9.5 millimeters per year at Kings Creek marsh using 210Pb dating. These rates are higher than relative sea-level rise (RSLR) at the NOAA tide gauge at the nearby town of Cambridge, Maryland (3.4 millimeters per year). One strategy to help combat sea-level rise is to allow Phragmites to persist in tidal marshes, instead of attempting to eradicate it with burning and herbicide applications (Stevenson et al. 2000). Downstream of Cambridge on the Choptank River, old maps reveal that isolated marshes were submerged over the last century, and marsh peat can still be found in shallow waters, particularly in Dickinson Bay. Such evidence supports the conclusion of Stevenson et al. (1988) that marshes are less likely to exhibit signs of significant deterioration where sediment inputs are significant—even if RSLR is high (in this case in the range of three to four millimeters per year). In studies at Monie Bay,

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Kearney et al. (1994) confirmed that marshes in a small estuary, where sediment supply from the watershed was relatively moderate, could show some signs of impacts of sea-level rise but nevertheless still maintain elevation with respect to sea level. Interestingly, sediment cores revealed that recent reductions in sediment loads (probably from sediment buffer strips along agricultural lands) had shifted marsh accretion in the youngest to mostly organic accumulation—often the first sign that the marsh is falling behind RSLR. Coastal wetlands on the western shore of Chesapeake Bay generally have more potential for input of sediment due to the higher elevation of surrounding landscape resulting in a greater energy gradient. The generalization has been that the marshes on the western shore of Chesapeake Bay have been keeping abreast of RSLR (Froomer 1980), and some even appear to have expanded during maximum land clearance in the 1800s. In a series of papers (Pasternack and Brush 1998, 2001; Pasternack, Brush, and Hilgartner 2001), marsh expansion was well documented in a prograding delta at the head of the Bush River, which historically had abundant sediment supplies due to land clearance. The observed net sedimentation in the Bush Delta averaged 1.0 gram per meter per year (Pasternack and Brush 2001). Farther south at the Smithsonian Environmental Research Center (SERC) surrounding the Muddy Creek off of the Rhode River, sediment inputs have also been historically significant from the eroding land in the watershed. At present, only a few acres of marsh have actually been lost at SERC, and these appeared to be due to overgrazing by muskrats (D. Whigham, SERC, 2006, personal communication). The most recent SET measurements suggest this balance between sea level and marsh surface elevation may be changing, as accretion in the brackish marshes at SERC appear to be falling about one millimeter per year short of RSLR, and aboveground biomass has been declining over the last fifteen years (B. Drake, SERC, 2006, personal communication). The apparent 186

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accretion deficit, coupled with lower biomass in recent years (fig. 10.10) at Rhode River (Erickson et al. 2007), is especially troubling, and more data need to be assembled to determine if this is a persistent downward trend suggesting that sea-level rise may be affecting this system. In Virginia, there is also increasing concern about the impacts of sea-level rise on marshes (C. Herschner, Virginia Institute of Marine Science, 2006, personal communication), especially at the head of the York River where the Pamunkey and Mattaponi rivers join near West Point. In the tidal freshwater region of the Mattaponi River, Dark and Megonigal (2003) reported that one of the most important factors in determining sediment availability to marshes was the distance to the high-turbidity zone. They concluded that upstream of this turbidity zone, marshes had much less access to sediments, and there was large variability in sediment dynamics depending on particular location. In fact, Dark and Megonigal (2003) reported an order-of-magnitude higher deposition downstream in the tidal freshwater zone of the Mattaponi than upstream. The survival of salt marshes northward of the mouth of the Chesapeake Bay, now incorporated into the Virginia Coast Reserve, has been of concern for the last couple of decades, especially since they comprise a Long Term Ecosystem Research (LTER) site. Oertel et al. (1992) measured 210Pb in five vibracores in a fringe marsh near Oyster, Virginia, and found accretion rates varied from 1.1 to 2.2 millimeters per year— considerably below the RSLR at the nearby Kiptopeke tide gauge (4.2 millimeters per year). Studies in the 1990s showed that the interior marshes were low in productivity, and sediments showed signs of water logging (low redox, high NH4), but there were areas where robust marshes were building on recently deposited materials (Osgood and Zieman 1993, 1998). More recently, Silliman and Zieman (2001) reported high densities and heavy grazing by periwinkle snails (Littoraria irrorata) on Spartina in this area. When left unchecked by a predator such as mud crabs, excessive grazing by

12

C3

E A

9

C

6 3 0 12

MX–C3

Peak standing shoot biomass (t ha-1)

9 6 3 0 12

C4

9 6 3 0 12

MX–C4

9 6 3 0 1986

1989

1992

1995 Year

1998

2001

2004

FIGURE 10.10 Mean (n ⫽ 5) aboveground biomass of four brackish marsh zones off the Rhode River: C3 ⫽ Scirpus olneyi, MX-C3 ⫽ Scirpus olneyi dominant, C4 ⫽ Spartina patens dominant, and MX-C4 ⫽ Spartina patens/Distichlis spicata mixture (from Erickson et al. 2007). Open circles are chambered plots with ambient CO2, black circles are chambered plots with elevated CO2, and triangles are controls.

Littoraria can result in complete denudation of marsh vegetation (Silliman et al. 2004). Kearney et al. (2002) suggested much of the marsh in this region appeared to be deteriorating in satellite imagery. Using SET measurements and marker horizons at Mockhorn Island and Curlew Pond near the town of Wachapreague, Virginia, Erwin et al. (2006) found that Spartina marsh accretion had not kept abreast of the long-term rate of RSLR for the mid-Atlantic coast. However, the approach of Erwin et al. (2006) in comparing SET measurements and marker horizon data taken over three to five years with long-term tide records is problematic, especially since they chose such a distant tide gauge (Sandy Hook, New Jersey). If the same study period is considered, sea level was actually dropping slightly at Sandy Hook as well as Atlantic City, New Jersey (fig. 10.11), so it could be argued that the same data set might be interpreted as actually showing a positive accretionary balance at Mockhorn Island and Curlew Pond during this short period. It is difficult to be definitive about the long-term status of accretion in the Virginia Coast Reserve because marker horizon and SET techniques do not take into account subsequent autocompaction caused by dewatering and compaction of organic materials. While little autocompaction is expected in mostly inorganic sandy marshes (French and Spencer 1993), some Holocene deposits retain as little as 10 to 20 percent of their original thickness if they are highly organic (Pizzuto and Schwendt 1997). Nevertheless, it is interesting that the bottoms of ponds may be deepening, suggesting a significant overall negative sedimentary balance in the marsh system of the Virginia Coast Reserve. This might corroborate the Lucke (1934) model discussed earlier and interpretation of sediment dynamics in this region by Kearney et al. (2002). More 210Pb and or 137Cs dated cores throughout the interior marshes of the reserve would certainly help resolve these issues. Farther north along this shoreline into Delaware and Maryland, the back barrier lagoons lack interior marshes, and most likely they have been submerged as old peats, which 188

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have been found in the bottoms of these embayments (J. C. Stevenson, personal observation). Not surprisingly, rates of erosion of existing fringe marshes along the western edge of the coastal bays in Maryland and Delaware are quite high. Schwimmer (2001) measured up to a half meter of lateral erosion per year in Rehoboth Bay. Much of this erosion was associated with undercutting of the marsh substrate by wave action. Since this undercutting is subtidal, normal summer wave activity from the southeast can cause significant erosion. The eroded sediments are usually deposited near the retreating marsh edge in the shallows, but they can be resuspended sporadically during storm events and provide a subsidy if they are redeposited on the marsh surface (Stumpf 1983). However, the short-term accretion Stumpf (1983) estimated (five millimeters per year) at Holland Glade marsh does not reflect the long-term accretion rates derived from more recent 210Pb data. Kim, Alleman, and Church (2004) measured sediment accretion at Wolf Glade marsh near Holland Glade using 210Pb and found the vertical accretion rate was 2.6 millimeters per year—considerably lower than the long-term RSLR at the Lewes, Delaware, tide gauge (~3.1 millimeters per year). Kim et al. (2004) did find considerably higher accretion (3.9 millimeters per year) at Oyster Landing in Great Marsh, which is better positioned to receive sediment from Delaware Bay during northeast storm events than Wolf Glade marsh. Since the bayside margin of these marshes erodes during northeast storm events, this is hardly a picture of equilibrium in terms of marshes keeping abreast of sea level (Kearney et al. 2002). Rather, we have previously regarded this phenomenon as better described as marsh cannibalism (Stevenson et al. 1988). In New Jersey, marshes appear to be equally dynamic. Philipp (2005) compared historical maps dating back to the 1840s and concluded that there have been extensive losses of salt marsh along Delaware Bay due to sea-level rise and other anthropogenic causes. Furthermore, Garofalo (1980) compared 1940 and 1972

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FIGURE 10.11 Tide records from Boston, Sandy Hook, and Atlantic City. Data from NOAA.

photographs of tidal marshes and established that there was considerable instability of the stream channels and hypothesized that this was due to storm events. He concluded that vegetation was the controlling factor over stream migration since there was a much greater rate (thirty-two centimeters per year) in nonsaline marshes, where vegetation does not hold the banks, than in saline marshes (twenty-one centimeters per year), where channel banks are more protected from erosion by plant roots that bind the muddy sediments. Farther north along the New Jersey coast, one of the first uses of 14C dating was at Brigantine National Wildlife Reserve, north of Atlantic City, now part of the wilderness area at the Edwin B. Forsythe Wildlife Refuge. Stuiver and Daddario (1963) found the oldest peat (5,890 radiocarbon years BP) was thirteen meters below high tide, yielding an overall accretion rate of 2.2 millimeters per year. This rate was close to that reported by Redfield and Rubin (1962) at Barnstable Marsh on Cape Cod (Massachusetts) and appeared to corroborate the concept that these marshes were in equilibrium with sea level. Since that early study, there has been a dramatic change in perspective in this region because of the latest study of accretion in the refuge. Recently, Erwin et al. (2006) reported that marshes west of Little Beach, New Jersey (using SET and marker horizon data), were only accreting at 1.7 millimeters per year, far short of the long-term average of the nearest NOAA tide gauge (4.1 millimeters per year). The lack of sufficient vertical marsh accretion at Little Beach is compounded by the deepening of ponds (–5.5 millimeters per year), suggesting a considerable sedimentary disequilibrium of the entire platform could exist (Erwin et al. 2006). Southern New Jersey has a particularly high RSLR, since it is just south of the limit of glaciation and is in the forebulge collapse region (Peltier 2001); in addition, subsidence in this area of New Jersey can be as high as two millimeters per year due to groundwater withdrawals (Sun, Grandstadd, and Shagam 1999). 190

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Particularly large uncertainties should be expected when predicting marsh submergence responses in areas where local subsidence is occurring (Rybczyk and Cahoon 2002). Moreover, as we pointed out previously, mean sea level (fig. 10.11), along the mid-Atlantic coast was actually decreasing over the course of the short period that Erwin et al. (2006) did their study, so it remains debatable how marshes in this region are faring. Again, longer-term SET measurements and 210Pb dating could better resolve how well the marshes of the New Jersey coast are keeping abreast of sea level. It might be unexpected because New York City and Long Island are situated on the terminal glacial moraine, but predicted from Peltier’s (2001) model of crustal readjustment (fig. 10.12), it is clear that a considerable amount of marsh has been lost in Jamaica Bay, New York. This area is adjacent to a large sanitary landfill, as well as John F. Kennedy International Airport, which has a runway extending into the bay. Using digitized aerial photos of Jamaica Bay, Hartig et al. (2002) found that about 12 percent of the marsh system has been lost since 1959, with smaller islands losing as much as 78 percent of their vegetative cover. The center of one 130-acre marsh island was transformed into a mudflat. However, aboveground biomass of existing vegetation was in the range of 700 to 1,500 g m⫺2, which is considered high for S. alterniflora at this latitude (Hartig et al. 2002). In addition, 210Pb measurements of accretion were high, with the long-term average for the low marsh about eight millimeters per year and high marsh around five millimeters per year (Zeppie and Duedall 1977). Further complicating issues, Franz (1997) found very high numbers of subtidal filter feeders (e.g., Geukensia demissa) on the mudflats and subtidal areas around the declining islands and has hypothesized that they have been intercepting sediment that might otherwise have nourished the surface of the marsh. Hartig et al. (2002) concluded that marsh losses in Jamaica Bay are driven by a lack of sediment inputs associated with urbanization of the entire watershed over the last 150 years.

UPWARD MOTION 12 9 6 3

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FIGURE 10.12 Holocene crustal movements: Peltier model. From Scientific American 1997.

Ultimately, marsh survival may be compromised by the fact that the prograding sand spit of Rockaway Beach, New York, deprives the lagoon of sediment. Intensive residential development on Rockaway Beach now prevents significant transport of sediment in overwash during storms into Jamaica Bay, which likely occurred before development. Moreover, groundwater sources have also been cut off by surrounding urban development. This is most likely the cause of higher subsidence rates than those recorded at the New York City tide gauge located at the southern tip of Manhattan. The U.S. Army Corps of Engineers has now committed $13 million to restore seventy acres of saltmarsh at Elders Point Island in Jamaica Bay using 270,000 cubic yards of sediment dredged from various channels of New York Harbor. Farther north and east of Jamaica Bay, several early studies (Armentano and Woodwell 1975; McCaffrey and Thomson 1980; BrickerUrso et al. 1989) had previously determined that tidal marshes on Long Island, in southern Connecticut, and in Rhode Island were keeping abreast of sea-level rise. However, Nydick et al. (1995) made a high-resolution study of peat

deposition using 14C over the last millennia in the marshes around Guilford, Connecticut, and concluded that recent accretion was less than RSLR and that if this trend continues, large areas of interior marsh could be submerged. The study of Cochran et al. (1998) reinforced this concern when they reported accretion (using 210Pb) in six marshes around Long Island Sound and found that three fell below RSLR in the area (2.8 millimeters per year). The lowest accretion rate (1.1 to 1.2 millimeters per year) that Cochran et al. (1998) found was at Hunter Island, at Pelham Bay Park in the borough of the Bronx in New York City. Obviously, a number of anthropogenic factors may be involved at this urban location. For example, Anisfeld, Tobin, and Benoit (1999) found recent sedimentation rates (measured with 210Pb) in Connecticut marshes were 50 percent lower when marshes were hydrologically restricted compared with nearby reference sites. The most recent 210Pb and 137Cs accretion rates combined with reinterpretation of past sea levels in Narragansett Bay, Rhode Island (Donnelly and Bertness 2001), suggest a more dire situation than that portrayed by

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Bricker-Urso et al. (1989). Donnelly and Bertness (2001) reported 210Pb rates varying from 2.0 to 4.9 millimeters per year at Nags Creek Marsh and 2.1 to 2.2 millimeters per year at Rumstick Cove, with low marsh (S. alterniflora) having the highest accretion. From plant macro-fossils in their cores, Donnelly and Bertness (2001) concluded that previous RSLR has spurred S. alterniflora migration into a narrowing zone once dominated by high marsh species (Spartina patens, Distichlis spicata, and Juncus gerardii) and that additional climate change could result in the extensive loss of coastal marshes in New England. Near the border of Rhode Island and Connecticut at the Barn Island Wildlife Management Area, high marsh accretion appears to have lagged sea-level rise by about 30 percent. Orson, Warren, and Neiring (1998) reported that 210Pb accretion rates varied from 1.8 to 2.0 millimeters per year at the two high marsh sites (Headquarters and Blooms Point) that they studied at Barn Island compared to the RSLR (3.0 mm per year) at New London, Connecticut, ten kilometers to the west. The accretionary deficits correspond to the loss of J. gerardii from the upper marsh border, along with the conversion of S. patens dominated high marsh to stunted S. alterniflora (Warren and Niering 1993). According to Orson et al. (1998), these vegetation changes could herald submergence problems along this coastline. In contrast, accretion (also measured with 210Pb ) at the low marsh site at Barn Island was 3.3 millimeters per year, which is in the range for marshes (2.4 to 6.0 millimeters per year) in Narragansett Bay, Rhode Island (Bricker-Urso et al. 1989), and in Waquoit Bay, Massachusetts (2.7 to 3.7; see Orson and Howes 1992). Farther out on Cape Cod, accretion has been measured with a variety of techniques in the 945-hectare Nauset Marsh located five kilometers south of the town of Eastham, Massachusetts. Roman et al. (1997) dated three cores with 210Pb and 137Cs and found close agreement between the two radiometric techniques. Highest accretion (4.2 to 4.4 millime192

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ters per year) was at Fort Hill Island, while the lowest was at Nauset Bay (2.6 to 3.8 millimeters per year). The Inlet 1 site was intermediate (3.8 to 4.0 millimeters per year) in terms of 137Cs and 210Pb (respectively), but a high of 24 millimeters per year was recorded there using feldspar marker horizons between June 1991 and April 1992 due to Hurricane Bob (August 1991) and the “Halloween Storm” of 1991 (Roman et al. 1997). Despite accretion rates exceeding sea-level rise at the Boston gauge (2.4 millimeters per year) from 1921 to 1993, Roman et al. (1997) mention that marsh accretion rates were spatially and temporally variable and dependent to a large degree on flooddominated sediment dynamics and overwash processes to supply sediment to the system (Aubrey and Speer 1985; Leatherman and Zaremba 1986). In addition, Roman et al. (1997) noted vegetation changes that could be related to sea level. For example, over the last forty years, S. patens has been replaced with Distichlis spicata, which has broader salinity tolerance (Bertness and Ellison 1987). More recently, Erwin et al. (2006) reported SET and marker horizon measurements at Nauset Marsh indicating that the Spartina marshes were just keeping abreast of sea level during their five-year study. Largely because of Redfield’s (1965, 1972) papers, Cape Cod marshes have often been viewed as among those least susceptible to sealevel rise, but in view of the most recent studies, they may be showing some signs that they are approaching a tipping point in terms of their resilience to sea-level changes. Due to the large amount of bedrock along the coastline, Maine has a relatively small aerial extent of salt marsh, only about seventy-nine square kilometers (Jacobson, Jacobson, and Kelley 1987). Wood, Kelley, and Belknap (1989) found accretion ranged from zero to thirteen millimeters per year as measured by marker horizons at twenty-four sites. Generally, marshes in back-barrier environments had the highest rates, while bluff-toe marshes had the lowest rates. The lowest accretion rate was zero in a bluff-toe marsh at Wharton Point, a peninsula intruding into

Casco Bay where sea-level rise was estimated at 2.4 millimeters per year. Interestingly, at a site a few kilometers away at the head of Maquoit Bay, a fluvial marsh accreted at 3.1 millimeters per year. Wood et al. (1989) found surprisingly little correlation of accretion with tidal amplitude, but they hypothesized that ice rafting might be key in understanding allochthonous sediment inputs, which can be crucial to surficial accretion in this harsh climate. Farther north, in a study utilizing 210Pb, 137 Cs, and pollen dating techniques, Chmura et al. (2001) also cited ice rafting as a possible mechanism to account for the historically high rates of marsh accretion in three marshes located on the Bay of Fundy between Eastport, Maine, and St. John, New Brunswick. Over the last one hundred years or so, they suggested that these marshes were in equilibrium with sea level, despite their finding that 210Pb accretion rates were only 1.5 millimeters per year at two sites (Little Lepreau and Dipper Harbor, New Brunswick), with a slightly higher rate (1.7 millimeters per year) at Chance Harbor, New Brunswick, compared to a comparable time span at the nearest tide gauges at Eastport (2.8 millimeters per year) and St. John (2.1 millimeters per year). Chmura et al. (2001) raised an important point: global warming might reduce the amount of ice rafting in the future, which in turn could reduce sediment inputs to the marshes on the Bay of Fundy. In another study, Chmura and Hung (2004) compared the 137 Cs data collected at the three Bay of Fundy sites by Chmura et al. (2001) with twelve other sites scattered widely in New Brunswick, Prince Edward Island, and Nova Scotia. All the other sites had accretion rates higher than the range in the Bay of Fundy (1.5 to 1.9 millimeters per year), with the highest accretion rates in Nova Scotia near Yarmouth (3.9 millimeters per year) and near Halifax (3.8 millimeters per year). Interestingly, Chmura and Hung (2004) concluded that the most significant predictors of accretion rates (in order of importance) were organic matter inventory, distance to the nearest creek, and range of mean tides.

In view of the predictions of global warming being more drastic at high latitudes, we are less sanguine than local Canadian researchers about the implications of these recent accretion rates in the Maritime Provinces. As we pointed out earlier, marshes that depend on a high proportion of organic material for accretion are generally at elevated risk for loss. Not only are marshes that depend on organic accretion more compressible and subject to compaction, but also as peats warm, the organic matter is more readily oxidized (during droughts) and if hydrated can be susceptible to acidification, promoting “sudden die-off” events (McKee et al. 2004). Since acid precipitation is a widespread problem in eastern Canada, it might present an even more dramatic acidification events on the marsh, especially if severe rain follows a drought (and the surface is not quickly buffered by seawater).

MARSH LOSS AND EXPANSION IN EUROPE Perhaps the most classic case of sediment starvation is in the 550-square-kilometer Venice Lagoon, where the major rivers (i.e., Brenta, Piave, and Sile) flowing into it were diverted several centuries ago to reduce sedimentation (Stevenson et al. 1999). The inlets of the lagoon are ebb dominated, so as much as a million cubic meters of sediment are lost to the Adriatic Sea per year (Bettinetti, Mattarolo, and Silva 1995). The marshes in the lagoon have considerable plant diversity, with Puccinellia palustris, Inula crithmoides, and Sueda maritima along the creeks, while Spartina maritima, Limonium narbonense, and Juncus maritimus are distributed more widely over the surface (Mariani et al. 2004). Due mostly to the overall sedimentary imbalance of the lagoon, it has deepened and substantial marsh acreage has been lost to mudflats (Fagaherazzi et al. 2006). In 1912, the lagoon had about 149 square kilometers of marsh compared with only 47 square kilometers in 1997 (Mariani et al. 2003). Much of the loss corresponded to the period of high RSLR

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associated with subsidence due to aquifer pump-out (Gambolati, Gatto, and Freeze 1974). In a series of studies, Day et al. (1998, 1999) concluded that sediment deposition at the marsh sites they studied in the southern part of the lagoon averaged three to seven grams dry weight per day, with highest rates where suspended sediments were available for transport to the marsh surface during wind-driven highwater (i.e., aqua alta) events. Recently, Rolinski and Umgiesser (2005) constructed a hydrodynamic model of the Venice Lagoon, which appears to confirm that the success of marshes in keeping abreast of sea level depends on a relatively small number of wind-driven events when the marsh surfaces are submerged. Day et al. (1999) concluded that without major rediversions of rivers into the lagoon, four of the seven marshes they studied would succumb to even a low range of RSL increase of fifteen centimeters over the next century, as projected in the Third Intergovernmental Panel on Climate Change Report (Houghton et al. 2001). Another area of concern in the Mediterranean is the Rhône Delta or “Camargue” in southern France. Tidal range is only about thirty centimeters, so the Rhône Delta is potentially susceptible to sea-level rise (Corre 1992). Using SET measurements in the marshes, Hensel,

Day, and Pont (1999) reported a mean sedimentation of 13.4 millimeters per year for riverine sites, which have a direct connection to the Rhône, in contrast to 1.1 millimeters per year for impounded sites and 1.2 millimeters per year for marine sites between the beach dunes and the sea dike. They concluded that the most important inputs over the last several decades occurred when the Rhône flooded in 1993 and 1994, leaving thirty millimeters of sediment in the riverine sites, but nothing in the impounded and marine sites. Another anthropogenic stress for the marshes of the Rhône Delta is the large amount of grazing, which lowers production significantly. Productivity in a nongrazed Phragmites australis–Scirpus maritimus site was 824 grams per year, whereas a grazed site nearby was stripped entirely of aboveground biomass, leading to complete loss of vegetation (Ibanez, Day, and Pont 1999). In sharp contrast to the Venice Lagoon and the Camargue, where riverine inputs are limited along the seaboard, on the shores of the MontSaint-Michel and Gulf of Saint-Malo bays of the northern coast of France, sediment deposition in the intertidal zone has been high in recent years, and salt marshes have been expanding (fig. 10.13; Tessier 2002; Haslett et al. 2003). This is in an area where contoured plots show that relative

FIGURE 10.13 Marshes west of the causeway to Mont-Saint-Michel in northern France have expanded in recent years and represent one of the few areas of the world where vertical accretion has outpaced sea-level rise. Photo by J. C. Stevenson, June 2006.

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sea-level rise is rather modest, in the range of 1.5 to 2.0 millimeters per year (Douglas 2001), but the Island of Jersey gauge in the middle of the Gulf of Saint-Malo, reveals a rate of 3.9 millimeters per year (Haslett et al. 2003). High sedimentation rates (7 to 14 millimeters per year) reported from the salt marshes around Mont-Saint-Michel (Jigorel 1995) have been largely attributed to the tourist causeway, which was constructed in the 1800s without any openings. Sediment has accumulated along the causeway to Mont-SaintMichel, changing mudflats previously inhabited by polychaete worms including Nereis diversicolor and Nereis virens (Olivier et al. 1995) to salt marsh dominated by Spartina anglica, Salicornia europaea, Salicornia dolystichostachia, Suaeda maritima, Aster tripolium, Halimione portulacoides, and Puccinellia maritima (Tessier 2002). Tidal range in Mont-Saint-Michel Bay reaches as much as fourteen meters (Lambeck 1997), and strong currents once scoured this section of coast before the construction of the causeway. There is now a large restoration project to build a new causeway with openings to facilitate tidal flow with the goal of reducing the acreage of tidal marsh in the area. However, Haslett et al. (2003) found salt marshes building in three small estuaries forty kilometers or more from Mont-Saint-Michel Bay: Havre de Carteret, Surville, and Lessay. They report accretion rates (determined with 210Pb and 137Cs) in the range of 4.1 to 7.7 millimeters per year, well in excess of RSL rise over the same period. This suggests that marsh accretion is a general phenomenon brought about by high sediment supplies and strong tidal currents. It would be surprising if changing the configuration of the causeway has a significant impact on marsh progradation in Mont-Saint-Michel Bay. As sea level rises, however, it may help reduce mudflat to marsh conversion and could eventually contribute to a net marsh loss in this area—without a large and expensive project, which was conceived at a time when global sea-level rise patterns were not well established (Werther 1984). Across the English Channel, marsh erosion has been sporadically reported—even though

relative sea-level rise has been minimal (one to two millimeters per year) over the last century (Woodworth et al. 1999). Van der Wal and Pye (2004) reported marsh erosion downstream of London has been as high as sixteen hectares per year in the Greater Thames Region (i.e., open coast of Dengie and Foulness, as well as the Blackwater and Thames estuaries). However, vertical accretion (two to three millimeters per year) of the mature marshes in this area was generally keeping abreast of sea-level rise. According to Van der Wal and Pye (2004), most of the marsh losses in the Greater Thames Region were associated with lateral erosion due to high waves during southeastern wind events, which occurred especially during the 1970s. This is at odds to the often-repeated generalization that storms result in marsh accretion (Stumpf 1983; Baumann et al. 1984). Earlier, we (Stevenson et al. 1988) disagreed with that generalization and argued that the direction of winds and ebb or flood domination was crucial to whether storms ultimately contribute to net accretion versus net erosion. In addition, storm energies are proportionally more important in accretion and erosion processes in microtidal systems than in mesotidal and macrotidal. Especially in the latter systems, strong tidal currents often can subsidize or erode marshes much more readily than occasional storms.

MARSH SURVIVAL IN A PERIOD OF ACCELERATED SEA-LEVEL RISE As one of the most productive components of the coastal ocean, tidal marshes are a subject of considerable concern as RSLR accelerates further due to global warming. Recent evidence of substantial melting of the Greenland ice cap (and possibly parts of Antarctica) gives notice of a substantial global rise in sea level from mass change, not just thermal expansion (Carton, Geise, and Grodsky 2005). The question arises: what can be done to alleviate a catastrophic loss of coastal marshes, which could not only disrupt critical food chains and habitat but also lead to species extinctions?

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First, it would helpful if we could provide managers with more insight to which marshes might be most at risk and unable to adjust to a potential RSLR not seen since the middle Holocene warm period about 8500 to 5000 BP. Despite high RSLR, robust sediment supplies and high tidal energies can promote marsh survival. Unfortunately, these conditions do not exist in many estuaries along the U.S. midAtlantic coast and in estuaries of the Adriatic Sea such as the Venice Lagoon. As we have seen, increasing numbers of marshes along microtidal (and possibly low mesotidal) coasts appear to be negatively affected by RSLR. If marshes go beyond the tipping point, they can erode, as at Blackwater, with expansion of tidal creek systems and formation of interior ponds. On the other hand, marshes at the heads of estuaries with ample sediment supplies (and higher tidal amplitudes) should remain intact and functional through the twenty-first century. A major obstacle in formulating policy is the sheer number of microtidal marshes potentially at risk. Perhaps even more disquieting, we have little in the way of indicators to warn of imminent degradation—until it happens. The reasons for “sudden marsh die-off” must be better understood before researchers can best advise the management community how to remedy the problem. The health classification of marshes using satellite imagery (Kearney et al. 2002) may help identify marshes at risk, but considerable ground truthing needs to done before this becomes an effective screening tool. As Turner et al. (2004) noted, aboveground biomass can remain nominally vigorous in plants like S. alterniflora, while the perennial part of the plant, the rhizome, is stressed. Recent data from Blackwater suggests that hydrogen sulfide levels are often elevated in the range of two millimoles per liter preceding “sudden dieback,” and this may provide managers with an early indicator of marsh distress. The rapid onset of marsh loss and its linkage to conditions related to stress in the belowground biomass require rethinking on how best to plan for future. In planning for future sealevel rise scenarios, we advocate two directions 196

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for further research. As we have discussed, one potentially fruitful direction would involve the development of methods for obtaining a broad regional assessment of marsh vulnerability to future sea-level rise. Obtaining more vertical accretion data can also help make predictions, but we should be more selective about using various techniques. In particular, SET and marker horizon measurements are often made for a few years and then compared with sea-level data on a different time scale. We need longer-term measurements to improve our assessment of marsh accretion, and more measurements need to be made per marsh to fully cover the inherent variability of the marsh surface. Furthermore, there is now no clear way at present to determine the compressibility (autocompaction) of recent layers of sediment, which contributes to marsh subsidence. Although 210 Pb or 137Cs radioisotopic measurements are an improvement, they still do not completely solve this problem. Both dating techniques, where vertical accretion rates are slow, can lie within the zone of marsh sediment dewatering and compaction and lead to underestimation of accretion deficits (Kearney and Ward 1986; Anisfeld et al. 1999). More research on the impacts of reduced groundwater tables in relation to marsh loss is essential. Although somewhat conjectural, it appears that the losses in Jamaica Bay and Blackwater may have been exacerbated by lowering of surficial groundwater tables. When flushing via groundwater and tidal action is weak, sulfides build up in the interstitial porewaters of the rhizosphere, slowing nitrogen uptake and plant productivity (Bradley and Morris 1990). Furthermore, when groundwater is lowered, peat can dry out more easily during droughts, promoting pyrite oxidation and acidification, which ultimately remobilizes toxic metals such as iron and aluminum, further jeopardizing the plant’s ability to assimilate other critical elements (McKee et al. 2004). At present, only a few studies (Mendelssohn and McKee 1988; McKee et al. 2004; Silliman et al. 2005; Ogburn and Alber 2006) have tried to investigate the ecological changes that lead to

stress and eventually, if continued, to “brown marsh die-off” or, as it is known along the Atlantic Coast, “sudden marsh die-off.” Turner (2004) demonstrated that aboveground biomass can continue to appear relatively vigorous despite increasing impairment of the physiological functioning of rhizomes in S. alternifora. Because clip plots traditionally have formed the basis for assessing marsh plant vigor, this apparent disconnect between conditions above- and belowground is disconcerting and, at worst, foster unfortunate management decisions doomed to fail. Though linkages between aboveground biomass and rhizome stress are not fully understood, progress has been made at identifying problems with metal uptake and sulfide toxicity, but more quantification is necessary on a wide variety of marsh systems before we understand the geochemical paths and nutritional implications to tidal marshes in high-RSLR regions. One problem is that often the root causes of the dieback are not actually present when the site is investigated (Ogburn and Alber 2006). If marshes can be identified as physiologically stressed and at high risk of die-off, managers can often help survival. One strategy that may help is to reduce grazing pressure on marshes by controlling herbivores (geese, nutria, snails, etc.), which can accelerate the demise of marshes. This strategy has been employed in a massive trapping effort to reduce the population of nutria (Myocastor coypus), which had been introduced in 1937, in and around Blackwater National Refuge on the eastern shore of Maryland. Ultimately, forecasting marsh survival in a future where global sea levels may rise more rapidly than at any time in the last several millennia will require an improvement of marsh models that truly couple hydrogeomorphic processes with physiological functioning. At present, we believe that recent models (Morris et al. 2002; Kirwan and Murray 2007; D’Alpaos et al. 2007) cover only the rudimentary geomorphic processes well. It is especially noteworthy that the D’Alpaos et al. (2007) model suggests an equilibrium in a lagoonal system where twothirds of the marshes have already been lost and

are still eroding at a rapid rate (Mariani et al. 2003). An obvious problem with Kirwan and Murray’s model (2007) is it does not predict widening of marsh tidal creeks, even under very high RSLR. Even in areas where RSLR is low, such as San Francisco Bay, movements of creek banks are detectable (Gabet 1998), and these movements are amplified when RSLR increases, as is the case in the declining marshes in Maryland and elsewhere in the mid-Atlantic (Kearney et al. 1988, 2002). Another shortcoming is the “geomorphic modeling” approach that incorporates plant communities mainly as biological architectures for capturing suspended sediment. These type of models often assume that marsh plant biomass will adjust forever to increasing levels of substrate flooding from rising sea level by ratcheting up vertical accretion rates without accounting for the sulfide toxicity incurred from increasing duration and depth of flooding (Koch and Mendelssohn 1989; Koch, Mendelssohn, and McKee 1990). Without taking into account such physiological feedback mechanisms (whereby energy reserves of plants are depleted to the point where less robust culms are produced and plants are less able to lay down organic material and trap sediments needed for accretion), these modeling exercises can only be regarded as heuristic. Although French’s (2006) model is still in the exploratory phase, his attempt to couple autochthonous and allochthonous processes to predict tidal marsh sedimentation with respect to sea level appears more potentially useful for management purposes. What else can be done to help manage marsh losses, other than support efforts to curtail greenhouse gas emissions? One of the more innovative strategies to emerge around Chesapeake Bay in recent years is the increasing use of living shorelines to protect fringe marshes from lateral erosion (fig. 10.14). These often depend on a low offshore breakwater to buffer the marsh from excessive wave action. As we have seen, sediment supply can be critical to survival. In several of these projects reviewed by Burke, Koch, and Stevenson

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FIGURE 10.14 Living shoreline project at the Aspen Institute on the Back Wye River, with marsh establishment behind a low sill, in Queen Annes County, Maryland. Photo by J. C. Stevenson, May 2004.

(2005), sediment that was either eroded from the toe of the scarp behind the marsh (if present) or transported landward from where the subtidal zone ended up, helping nourish the marshes behind the breakwaters. While these “soft technologies” may work well in a variety of low-energy shorelines for a while, allowing us to buy time, they do not work along high-energy shorelines and cannot be expected to stave off the impacts of sea-level rise indefinitely. Unfortunately, some of the current management practices sponsored by government agencies may actually contribute to marsh loss. Satellite images show that the marshes which are under federal and state jurisdiction in the Chesapeake and Delaware regions are often those most affected by marsh loss (Kearney et al. 2002). Common practices carried out to manage marshes on federal and state lands in this region include creating impoundments for waterfowl and extensive burning to create uniform grass and sedge swards free from woody shrubs. Both these practices may result in losing refractory carbon either as methane (from freshwater impoundments) or carbon dioxide (burning of marshes in winter) emissions to the atmosphere, causing more problems in terms of greenhouse warming. These practices need to be curtailed in areas where RSLR has stressed tidal marshes to their tipping points. Also, widespread efforts by various government agencies to reduce sediment inputs into 198

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estuaries tend to decrease the potential for marsh accretion. It would be beneficial if runoff containing high-sediment loads could be directed from the uplands to marshes where accretion deficits have been documented. Where there are farms in the watershed, it might be possible to divert agricultural ditches directly into marshes so that they could subsidize the marsh surface during storm events. This practice would help marshes and avoid wholesale contamination of estuaries. On a much larger scale, Day et al. (1998, 1999) have suggested that rediverting the sediment-laden rivers back into the Venice Lagoon (as they were naturally) would help restore the sedimentary imbalance there, enabling marshes to recover to previous levels. Subsidizing marshes with subtidal sediment has already been utilized where dredging projects can provide materials. Sediment subsidies have been carried out using thin-layer spray techniques (Cahoon and Cowan 1988) or simply by adding slurried material to the system (Mendelssohn and Kuhn 2003). In Chesapeake Bay, there have been several marsh creation projects dating back to the early 1980s (Perry et al. 2001). The largest is the restoration of Poplar Island and includes the creation of 625 acres of tidal marsh from dredged materials derived from the navigation channels of the Upper Bay (fig. 10.15). This approach appears to be largely successful thus far, and a plan is now emerging to supplement large areas of marshes around the Blackwater River with sediment

FIGURE 10.15 Spartina plantings in dredged material in cell #3D at Poplar Island, Maryland; the fencing and strings help reduce grazing by waterfowl and other wildlife. Photo by J. C. Stevenson, September 2005.

from main-stem Chesapeake channels. As sea level rises, however, these marshes may again have to be subsidized with sediment to keep them from becoming waterlogged. One might argue that this is a costly solution to the problem, but it may be one of the few ways we can have a safe refuge for species survival—until carbon dioxide emissions can be stabilized in the future and global sea-level rise relaxes. On the other hand, if nothing is done to alleviate carbon dioxide and other greenhouse gas emissions, and global warming results in world sea levels reaching 1.4 meters higher than present by 2100, our present understanding of marsh–sea level relations supports the strong possibility that almost all of Chesapeake Bay and other mid-Atlantic marshes will then be in severe decline, if not lost altogether. Likewise, it is highly doubtful that Louisiana coastal marshes could survive under such a sea-level scenario without the input of suspended sediments at levels of where they were a century ago—and even then, substantial losses could occur. Dire consequences are also probably in store for the marshes of the Venice Lagoon and the Nile Delta, as well as the China Coast where sediment supplies have been curtailed by dams and river diversions. Indeed, only a few select locations, where sediment focusing occurs naturally along coastline, such as the marshes of Mont-Saint-Michel, may be able to survive such a rapid rate of sea-level rise. In short, we believe that as much as 90 percent of the tidal marshes worldwide could be in jeopardy by 2100. Whatever the sea-level future holds for coastal marshes, the potential threats are now serious enough to warrant more global cooperation in assessing their present and future status, as well as in determining what impacts their decline will have on coastal ecosystems.

Acknowledgments. The authors would like to thank two anonymous reviewers for their comments on earlier drafts of the manuscript as well as Lorie W. Staver and Catherine Pieper Stevenson for editorial assistance and help in

drafting the figures. In addition, numerous colleagues freely shared their ideas and helped improve our understanding of global change and marsh processes. This is contribution number 4240 from the University of Maryland Center for Environmental Science.

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Potential Impacts of Elevated CO2 on Plant Interactions, Sustained Growth, and Carbon Cycling in Salt Marsh Ecosystems Jordan R. Mayor and Caitlin E. Hicks Global atmospheric concentrations of CO2 and other greenhouse gases have dramatically increased as a result of human activities and now far exceed preindustrial values. According to the February 2007 Intergovernmental Panel on Climate Change (IPCC) Fourth Assessment Report, global increases in CO2 are primarily due to fossil fuel use and have led to generalized global warming. Much debate has focused on the potential role of ecosystems to mitigate rising atmospheric CO2 through C sequestration into slowly cycling pools. Salt marsh ecosystems may be particularly well suited for C sequestration in ever-accruing sediments owing to their anoxic conditions, high productivity, and a potential for mineral nutrient inputs from adjacent terrestrial ecosystems. However, few studies have fully evaluated this potential. Here we summarize the body of literature surrounding experimental treatments of elevated CO2 in salt marshes and place it in context with other ecosystems to create a guiding framework for continued research on the ability of salt marsh ecosystems to act as C sinks under increasing CO2. We focus on the Smithsonian Environmental Research Center’s (SERC) project, which is the longest-running elevated CO2 experiment and the only one of its kind in a coastal salt marsh. Increased atmospheric CO2 concentrations in the SERC marsh has been shown to accelerate daytime net ecosystem exchange of C, to stimulate C3 plant shoot and root biomass accrual rates, to decrease C3 plant litter quality and decomposition, to increase soil CO2 and CH4 respiration, and to increase soil porewater C pool sizes. The role of these individual processes in determining the net ecosystem C balance of salt marshes remains to be determined. To this end, we provide research suggestions that will help constrain the possibility of coastal salt marsh ecosystems to act as long-term C sinks. Based on our synthesis of the extant literature, we recommend the preservation and expansion of existing coastal salt marsh ecosystems because their soils have the highest potential C accumulation rates of all coastal soils (Rabenhorst 1995). Specific management for marshes dominated by C3 plants is encouraged because they have the largest growth response to increased atmospheric CO2 concentrations due to their more responsive physiologies.

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Atmospheric carbon dioxide concentrations ([CO2]) have reached 380 parts per million as of 2006, a concentration that may be the highest in the last twenty million years (UNESCO-SCOPE 2006). The many negative impacts of increasing atmospheric [CO2] on our global climate, biological communities, and ecosystem processes have warranted a flurry of climate change research. Of particular interest to global change ecologists and policymakers is the potential capacity of ecosystems to act as long-term C sinks, thereby partially mitigating anthropogenic sources of atmospheric CO2 for decades to come. Understanding how ecosystems respond to increased [CO2] is of paramount importance to global change policy as terrestrial ecosystem C sinks are currently deemed unpredictable and generally unreliable (UNESCO-SCOPE 2006). Thus far, most elevated [CO2] experiments have reported direct CO2 stimulation of photosynthesis, plant growth, and soil respiration, yet it is unclear if these initial responses are temporary responses by nonacclimated or immature plants (Long et al. 2004; Nowak et al. 2004; Luo, Hui, and Zhang 2006). The unpredictable nature of terrestrial ecosystem C sequestration is partially due to the long time scale necessary to measure the amount of photosynthetically fixed C entering long-term pools of slowly decomposing materials, such as wood or soil humus, where it can be sequestered (Schlesinger and Lichter 2001). The unpredictability of long-term terrestrial C sinks are also due to the wide variability among ecosystem types, plant species (Bazzaz and Catovsky 2002), and methods used in elevated CO2 experiments (Poorter and Navas 2003). Furthermore, eventual cessation of continued C uptake in nitrogen (N)-limited terrestrial ecosystems is theoretically expected to occur as a result of progressive nitrogen limitation (PNL). In the absence of continued N additions or reduced N losses, the increase in plant productivity causes N (or other mineral nutrients) to be increasingly sequestered in plant biomass where it is unavailable for continued biomass accrual (Luo et al. 2004; Hungate et al. 2006). 208

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Salt marshes are the border between terrestrial and aquatic ecosystems, and therefore their response to increased [CO2] may differ from the response of terrestrial ecosystems. Salt marshes are generally N limited (Sundareshwar et al. 2003; Wijnen and Bakker 1999) but are unique with respect to PNL because they often receive additions of limiting nutrients from adjacent terrestrial systems and rivers (Martinetto, Teichberg, and Valiela 2006). This coupling of N cycles between land and coastal environments, while often viewed as pollution when in excess (Howarth, Sharpley, and Walker 2002), may act as a plant fertilizer that circumvents PNL, thereby allowing for continued C sequestration under elevated [CO2]. Whereas upland ecosystems tend to reach equilibrium when soil C inputs, via plant detritus, roughly equals C outputs, via respiration (Aber and Melillo 2001), leaching, or volatization (Chapin et al. 2006), salt marshes have the potential to accumulate soil C continuously owing to suppressed decomposition in anoxic soils. This chapter focuses on the experimentally observed and potential future impacts of rising atmospheric [CO2] on salt marsh ecosystems. We rely heavily in our discussion on the most recent published results of a long-term Department of Energy–funded Smithsonian Environmental Research Center (SERC) experiment subjecting a Chesapeake Bay tidal wetland plant community to elevated atmospheric [CO2]. We use examples of research in other, mainly terrestrial systems to add context to the results of the SERC experiments. In our discussion, we focus on the observed impacts of rising atmospheric [CO2] on tidal salt marsh plant interactions, biomass accrual, C mineralization, and C storage. Our discussion includes predictions of future plant community composition in light of multiple climate change factors and how PNL may potentially modify ecosystem processes that lead to C storage. In conclusion, we offer recommendations for future research in salt marsh ecosystems that may help settle the uncertainty of the role of salt marsh ecosystems as long-term C sinks.

SERC SITE DESCRIPTION Currently, there is only a single salt marsh where a long-term [CO2] enrichment study is being carried out. This site is a temperate, brackish salt marsh called Kirkpatrick Marsh, located on the South River subestuary of Chesapeake Bay, Maryland, referred to as the SERC marsh throughout this chapter. For every growing season (April to November) since 1987, plots at this site have been exposed to elevated [CO2] via open top chambers (figs. 11.1 and 11.2). This site has the distinction of being the longest-running CO2 enrichment site among all ecosystems. It has been running continuously since 1987, and as of this writing, its published results span eighteen growing seasons. While only a single site, results at the SERC marsh have the potential to be applied to salt marshes along the North Atlantic that contain similar plant communities comprised of both C3 and C4 species. Unfortunately, in terms of broad applications, the most common salt marsh

grass found along temperate shorelines of the western Atlantic, Spartina alterniflora, was not present at the SERC site. The vegetation at the SERC marsh can be divided into three categories: monospecific stands of a Scirpus olneyi, a C3 sedge; monospecific stands of Spartina patens, a C4 grass; and mixed stands of both species along with an additional C4 grass, Distichlis spicata (fig. 11.3). The C3 sedge, S. olneyi, typically inhabits areas of slightly lower elevation than the co-occurring C4 grass S. patens, which, being slightly less flood-tolerant, tends to inhabit higher elevations (Broome, Mendelssohn, and McKee 1995; Anastasiou and Brooks 2003). The open top chamber system (fig. 11.1) used at the SERC marsh is one of the most widely used methods of exposing ecosystems to increased [CO2] in situ. At the SERC marsh site, plots are in a threeblock design with each vegetation community containing five plots without chambers as controls, five plots with chambers receiving ambient [CO2], and five plots with chambers

0.8 m

Frustum

Remote blower

Main chamber

Mixing blower

1m

Lower plenum

Inlet plenum

FIGURE 11.1 A diagram of the open-top chamber used in the SERC marsh experiments. Air with elevated [CO2] was blown in through the inner plenum while the mixing blower recycles air from inside the chamber. Air can leave the chamber through the frustum. This figure was originally found in Drake et al. 1989.

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FIGURE 11.2 Image of the SERC elevated CO2 chamber system in Kirkpatrick Marsh, Maryland. Photo courtesy of Bert Drake, Smithsonian Environmental Research Center.

receiving elevated [CO2] at twice the 1987 ambient [CO2] (Drake, Gonzalez-Meler, and Long 1989). The chambers changed the microclimates of the plots slightly by increasing air temperature about 2°C and decreasing light input by 10 percent, but the effect on plant productivity is considered to be minimal (Drake et al. 1989). Similarly, open top chambers were found to warm a grassland community by 2.6°C (Morgan et al. 2001). However, chambers may affect soil processes and herbivory by decreasing macrofaunal abundance (Ball and Drake 1998), and they have been implicated in altering soil moisture contents (Morgan et al. 2001; Ainsworth and Long 2004) with unknown consequences.

RESPONSES OF SALT MARSH PLANTS TO ELEVATED CO2 The SERC research project has demonstrated the value of long-term CO2 fertilization experiments. After eighteen years of elevated [CO2], the capacity for continued stimulation of photosynthesis in the C3 plant S. olneyi community has been demonstrated. What is most remarkable about the SERC experiment is that unlike most experiments subjecting plants to elevated [CO2], the stimulated response of above- and belowground biomass has been sustained and even shows signs of increasing with time 210

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FIGURE 11.3 Multiple wetland plant species are present in the automated elevated CO2 system at SERC marsh. A mosaic of monospecific stands of the C3 sedge Scirpus olneyi and the C4 grass Spartina patens can both be seen here. Photo courtesy of Bert Drake, Smithsonian Environmental Research Center.

(Rasse, Peresta, and Drake 2005; Erickson et al. 2006). This response is significant to global change biologists as the response of plant communities to future increases in atmospheric [CO2] is often assumed to be transient (Bazzaz and Catovsky 2002; Poorter and Navas 2003; Ainsworth and Long 2005). Sustained increases in plant biomass under elevated [CO2] indicates the possibility of a long-term C sink, which could partially mitigate anthropogenic increases in atmospheric [CO2]. Therefore, understanding the response of tidal marsh ecosystems to elevated atmospheric [CO2] is of great importance to predicting the C sink capacity of our coastal ecosystems. In this section, we discuss the long-term trends in biomass responses, modifiers of these long-term trends, implications for competitive outcomes, and net ecosystem exchange (NEE) of C. OVERALL TRENDS IN BIOMASS RESPONSE

The biomass responses recorded at the SERC marsh during the summer months of each year include aboveground biomass, aboveground shoot density, shoot growth, and belowground biomass. Despite a general decline in biomass of all species and shoot density in the C4 species, over the past eighteen years at the SERC marsh, S. olneyi plots under elevated [CO2] have experienced the least decline. The cause of declining biomass and shoot density in both the unchambered control and ambient air chambered treatments may be due to the observed rise in sea levels, which increased by ten centimeters during the course of the SERC marsh study (Erickson et al. 2006). However, shoot biomass of the C3 sedge S. olneyi declined significantly less under elevated [CO2] treatments, and shoot density remained unchanged during the eighteen years of published data. The growth enhancement of S. olneyi due to elevated [CO2], hereafter referred to as stimulation (the standardized difference between the ambient air chambers and elevated CO2 treatments), averaged 35 to 45 percent for shoots (significant at a p value ⬍ 0.10; Rasse et al. 2005) and 26 percent for roots (Erickson et al.

2006). The biomass stimulation was 70 percent that at ambient [CO2] 2003 (fig. 11.4; Rasse et al. 2005). Similarly, shoot density was stimulated 120 percent in S. olneyi as of 2003 (Rasse et al. 2005). No stimulation in aboveground biomass was observed in neighboring S. patens plots, and only C3 shoot biomass increased in the mixed species plots (Erickson et al. 2006). Aboveground growth stimulation of S. olneyi under elevated [CO2] rose throughout the duration of the study with the exception of the initial two years, dubbed the “acclimation phase” (Drake et al. 1996; Rasse et al. 2005). This acclimation phase was characterized by a pronounced deviation from the overall trend and was subsequently excluded from regression analyses. Omitting the acclimation phase demonstrated that the stimulation in S. olneyi shoot biomass and shoot density explained 68 and 89 percent of the total yearly variance in S. olneyi aboveground biomass (Rasse et al. 2005). The responses of belowground biomass to elevated [CO2] at the SERC marsh differed slightly from the aboveground response. Root biomass was assessed from three 5 ⫻ 20–centimeter root ingrowth cores in each treatment plot. This method consists of removing the roots from a volume of extracted soil and monitoring the regrowth of roots into the replaced soil (Jordan and Escalante 1980). Elevated [CO2] significantly increased S. olneyi and the mixed species root biomass but had no effect on S. patens (p ⬍ 0.01; Erickson et al. 2006). The increased root biomass in the mixed species community was likely due to S. olneyi growth stimulation (Erickson et al. 2006). Long-term trends in root biomass response in monodominant S. olneyi plots were not evident; however, an increasing trend was evident in the mixed species plots, suggesting belowground competitive success of the C3 versus C4 plants. Absolute root ingrowth biomass was roughly equivalent between the C3 and mixed species communities (about 2.7 tons ha⫺1; Erickson et al. 2006). The magnitudes of the biomass responses at the SERC marsh were somewhat smaller than the 45 percent (C3 plants) and 12 percent

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(C4 plants) global average biomass responses determined from a meta-analysis of plant responses to elevated CO2 (Poorter and Navas 2003). The difference may be because the mean duration of the studies included in the metaanalysis was certainly shorter than that of the SERC site. Compared to average free-air CO2 enrichment (FACE) studies (Ainsworth and Long 2005), S. olneyi aboveground biomass response was greater (about 35 percent vs. 15 to 20 percent for C3 plants). MODIFIERS OF LONG-TERM TRENDS

Increasing biomass in response to elevated CO2 could potentially be of great importance because if the biomass responses continue to increase without an equivalent rise in marsh C losses, the marsh ecosystem may become a significant C sink. However, both salinity and competition can modify the effects of CO2 enrichment on plant productivity (Bertness and Ewanchuk 2002; Pennings, Grant, and Bertness 2005). In addition, changes in plant species or species nutrient status can result in drastic changes in herbivory pressure, reducing biomass accrual rates (Ngai and Jefferies 2004; Silliman et al. 2005). Therefore, it is important to understand how the biomass stimulation response varies with environment fluctuations and bottom-up versus top-down trophic interactions so that future predictions about C sequestration in coastal salt marshes can be better constrained. In general, the most immediate response of plants exposed to increased atmospheric [CO2] is an increased rate of photosynthesis and a decreased rate of leaf transpiration (Poorter and Navas 2003). Increased rates of C fixation are attributed to a greater amount of available substrate for CO2-fixing Rubisco enzymes and a decline in photorespiration owing to suppression of Rubisco oxygenation reactions at higher partial pressures of CO2 (Lambers, Chapin, and Pons 1998). Decreased transpirational water loss is due to partial closing of plant stomata that allow for equivalent amounts of CO2 diffusion into the plant under elevated CO2. At 192 parts per million CO2 enrichment, a 23 percent reduc212

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tion in stomatal conductance in thirteen prairie grassland species and a 40 percent increase in instantaneous water use efficiency were observed (Lee et al. 2001). In an analysis of thirteen field-based studies on tree species, stomatal conductance was on average 21 percent reduced under elevated [CO2] (Medlyn et al. 2001). With reduced transpirational water loss under high [CO2], salt marsh plants may be able to overcome periods of salinity-induced water stress. Salinity is an important environmental variable that can moderate growth owing to its physiological effects on plant osmotic balance (Penning et al. 2005). Excessive salinities in salt marshes create a situation of plant water stress, a process that can limit photosynthesis in ways equivalent to drought (Munns 2002). The salinity levels in the SERC marsh were lowest in years of high rainfall (low salinity) and generally highest during the years of lowest rainfall (high salinity). High-salinity years (arrows in fig. 11.4) elicited the greatest growth response in S. olneyi aboveground biomass to elevated [CO2], indicating that the C3 sedge was better able to ameliorate water stress when exposed to [CO2]. Yet even in the years of lowest salinity, biomass stimulation in S. olneyi remained relatively high (Rasse et al. 2005). In contrast, the C4 S. patens was not observed to respond to salinity levels (Erickson et al. 2006). Whereas the elevated [CO2] increased biomass stimulation of the C3 sedge, higher salinity augmented this stimulatory effect. In addition, S. olneyi biomass was positively correlated with mean sea level (p ⬍ 0.01), perhaps owing to this species’ greater flood tolerance, whereas S. patens biomass response was negatively correlated (p ⬍ 0.01; Erickson et al. 2006). Combined, this suggests that S. olneyi is better poised to adapt to the multiple changes predicted to occur due to rising atmospheric CO2 such as rising sea level and more intense or frequent droughts (UNESCO-SCOPE 2006). The effects of increasing [CO2] may greatly influence the outcome of C3 and C4 plant competitive interactions. Stimulation of shoot density and shoot and root biomass in response to elevated [CO2] only in S. olneyi monodominant

patches suggests that the C3 sedge S. olneyi was better able to capitalize on higher [CO2], possibly through greater photosynthetic rates (Rasse et al. 2005) and water use efficiency. Similarly, in an agricultural system, the biomass of C3 cotton (Gossypium hirsutum) was believed to better compete with C4 sorghum plants when grown at higher [CO2] (Derner et al. 2003). In the SERC marsh, however, the combined biomass stimulation in mixed species (both C3 and C4) plots ceased after ten years (Erickson et al. 2006), suggesting a competitive plateau eventually limited the C3 plants’ continued positive growth response. Similar distribution controlling factors (salinity tolerance, competition) were also described as interacting to determining plant zonation in New England salt marshes (Pennings and Bertness 2001). Despite this result, the sustained growth stimulation of S. olneyi over eighteen years, even with fluctuations in salinity and sea level, suggest that S. olneyi is better able to adapt to increasing atmospheric [CO2] than other co-occurring salt marsh grasses such as S. patens and Distichlis spicata and will thus come to dominate the site with time. IMPLICATIONS FOR COMPETITIVE INTERACTIONS

The physiological differences between the C3 and C4 marsh plants are substantial and will likely determine the long-term competitive success of species interactions. The underlying mechanisms responsible for the more pronounced positive response of C3 plants to elevated atmospheric [CO2] are due to biochemical differences in the two photosynthetic pathways. Numerous field and laboratory studies subjecting plants to elevated [CO2] have led to the general finding that C3 photosynthesis is stimulated by increasing atmospheric [CO2] and that this stimulation is greater in C3 than C4 plants (Sage 2002, Poorter and Navas 2003; Ainsworth and Long 2004). Although exceptions do occur (Ghannoum et al. 2000; Watling, Press, and Quick 2000), the finding that the dominant C3 plant S. olneyi responded more strongly and in more dynamic ways to elevated

TABLE 11.1 Summary of the plant responses to the 17-year elevated CO2 SERC experiment

plant community

Gross CO2 assimilation Net CO2 assimilation Shoot density Shoot biomass Root biomass Root productivity

C3

C4

⫹⫹ ⫹⫹ ⫹⫹ ⫹⫹ ⫹⫹ ⫹⫹

⫹ ⫹ ⫺ ⫺ ⫺ ⫺

NOTE: Symbols indicate the magnitude of positive response (⫹, ⫹⫹) or lack thereof (⫺). Table modified from its original form found at www.serc.si.edu/labs/co2/marsh_physiology.jsp.

[CO2] than co-occurring C4 plants is not surprising. The results of the elevated [CO2] experiment in the SERC salt marsh indicate that increasing atmospheric [CO2] should create a situation where C3 salt marsh plants such as S. olneyi are at an adaptive advantage, and cooccurring C4 species (table 11.1), such as Spartina patens and D. spicata, may eventually become displaced due to their inability to capitalize on elevated atmospheric [CO2]. It should be noted that S. olneyi, S. patens, and D. spicata are a small component of total plant biomass in many temperate western Atlantic salt marsh communities, which are often dominated by the highly productive C4 grass, Spartina alterniflora (Mitsch and Gosselink 2000; Voznesenskaya et al. 2006). Therefore, we caution extrapolating species-specific interactions in the SERC marsh to other coastal marshes and emphasize the need to conduct comparable elevated CO2 experiments on S. alterniflora–dominated marshes. Furthermore, another species that needs to be considered is Phragmites australis, whose invasion of many Atlantic salt marsh ecosystems has been well documented (Saltonstall 2002). Phragmites australis has variable photosynthetic pathways exhibiting both C3 and C4 characteristics throughout its broad global geographic distribution (Antonielli et al. 2002), with distinct ecotypes found to vary physiologically with habitat (Zheng, Zheng, and Zhang 2000). Under elevated [CO2], the C3-like

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pathway would most likely become dominant in P. australis, allowing it to dominate salt marshes, especially if the salt marsh is also experiencing N loading (Bertness, Ewanchuk, and Silliman 2002). Changes in competitive interactions due to increased [CO2] can have far-reaching effects on the ecological functioning of salt marshes, as has been documented with the recent P. australis invasions. Phragmites australis invasions of salt marshes have led to competitive displacement of dominant species and changes in plant species abundance (Chambers et al. 2003), with direct implications for plant and animal biodiversity and corresponding trophic interactions (Minchinton and Bertness 2003; Silliman and Bertness 2004). A reduction of palatable plant species (Pennings and Silliman 2005) could cause herbivore population crashes or shifts of grazing pressure to other plant species. More relevant to this chapter are the changes in biomass accrual and soil carbon inputs that can occur with changes in the dominant plant species, such as when a P. australis invasion of an S. patens–dominated marsh changed vegetation structure and soil biogeochemistry (Windham and Ehrenfeld 2003). NET ECOSYSTEM EXCHANGE

Net ecosystem exchange (NEE) is the measure of C exchange between the ecosystem and the atmosphere. It is comprised of the gross primary production (GPP), the amount of C fixed by the plants, minus total respiration (Recosystem), the amount of C leaving the system in the form of respired CO2 (Chapin, Matson, and Mooney 2002):

nighttime respiratory fluxes from heterotrophic decomposers and cannot be used a measure of net C exchange with the atmosphere. Nevertheless, its measure is an important starting point for determining trends in photosynthetic respiratory fluxes under elevated [CO2]. Daytime NEE (NEEday) of S. olneyi was measured during each summer season of the SERC experiment by enclosing the chamber systems for short durations of time and measuring the changes in [CO2] using a portable infrared gas chromatograph. Net ecosystem exchange was not measured in S. patens or mixed species plots. The findings were in many ways similar to those of S. olneyi biomass and shoot density stimulation. Disregarding the initial acclimation phase, NEEday was stimulated on average by about 35 percent, a magnitude that rapidly stabilized throughout the study period (p ⫽ 0.002; Rasse, Li, and Drake 2003; Rasse et al. 2005; fig. 11.4) despite fluctuating by as much as 20 percent between years. This stabilization of NEEday was unlike the stimulation in shoot biomass and density, which continued to rise over time. Similar to the interannual variability in growth, variability in NEEday was largely correlated with salinity. High salinities likely caused water stress during the drought years of 1995, 1999, and 2002 when the enhanced NEEday was depressed below the 35 percent mean stimulation (Rasse et al. 2005). Overall, the increase in NEEday due to elevated [CO2] could only be perceived as a negative feedback to atmospheric [CO2] if other respiratory fluxes were known so that net C exchange could be calculated.

CARBON CYCLING IN MARSH SOILS NEE ⫽ GPP ⫺ Recosystem

Its measure is the first step in determining if the SERC marsh ecosystem will function as a C sink in response to elevated CO2. In the SERC site, Rasse et al. (2005) measured NEE only during the day (NEEday), when photosynthetic respiratory fluxes dominate outputs of C to the atmosphere. Thus, NEEday does not include 214

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Plants are responsible for the majority of C inputs to salt marsh ecosystems, but the fate of this C is largely determined by microbial communities living in soils. When plants respond positively to elevated [CO2] by greater belowground C allocation, then more substrate is available for soil microbes due to greater inputs of fine roots (Drake et al. 1996; Marsh et al.

140 Day-time NEE 120

Shoot biomass Shoot density

Stimulation (%)

100

80

60

40

20

0 1986

1988

1990

1992

1994

1996

1998

2000

2002

2004

FIGURE 11.4 Stimulation of shoot biomass, shoot density, and daytime net ecosystem exchange (NEE) of Scirpus olneyi grown under elevated atmospheric [CO2] at the Smithsonian Environmental Research Site, Chesapeake Bay, MD. Arrows indicate the driest years, which exhibited a correspondingly higher salinity. Data points were extracted from there originally published form found in Rasse et al. (2005). Years of highest rainfall are inversely correlated with salinity (Erickson et al. 2006).

2005) and root exudates. The increased productivity can affect microbes by (1) increasing their numbers and biomass, (2) changing their community structure, or (3) increasing their metabolic activity, including respiration and enzyme production. All of these microbial processes can affect C residency in marsh soils. The lability (decomposability) of plant litter may also change with increased [CO2], which also affects C residency time. It is unlikely that elevated [CO2] will have a direct impact on microbial communities because air enriched in CO2 is still several magnitudes lower than in soils (Ball and Drake 1997). Furthermore, in SERC marsh mesocosms when the effects plants have on soil respiration were controlled for, elevated [CO2] was shown to have no direct effect on soil respiration (Ball and Drake 1997). Possible changes in microbial community and activity caused by increased plant productivity will in turn affect the residency time of organic C and mineral

nutrients and have lasting effects on ecosystem services such as productivity and C sequestration. Because of the dearth of soil C cycling studies in salt marshes subjected to elevated [CO2], examples from upland terrestrial ecosystems are used throughout this section to provide a wide perspective. MICROBIAL COMMUNITIES

If microbial biomass and community composition change under elevated [CO2], there may be subsequent changes in the way C is cycled within marsh soils. Microbial biomass is a measure of the amount of C stored in the microbial pool, whereas microbial community composition is a measure of the number and identity of microbial functional groups or species that are present in soils. Each functional group can have different enzymes, which affect the decomposition rates of various C pools. For example, bacteria and fungi preferentially decompose

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different pools of C, with fungi mineralizing organic C from complex structures like lignin and cellulose, and bacteria mineralizing simpler compounds like glucose (Paul and Clark 1996). To our knowledge, there are no studies of microbial biomass and community responses to elevated [CO2] in salt marshes; however, there are in terrestrial systems. There has not been a consistently significant response of microbial biomass to elevated [CO2] in many systems, including a shortgrass steppe (Kandeler et al. 2006), chaparral (Allen et al. 2005), or controlled experimental environments (Jones et al. 1998; Lipson et al. 2006). In a review of fortyseven studies, microbial biomass responses to elevated [CO2] varied widely, ranging from large increases to large declines (Zak et al. 2000). Bacterial communities and diversity have not been shown to change significantly with higher [CO2] (Jones et al. 1998; Lipson et al. 2006), but fungal community composition (Jones et al. 1998), abundance (Carney et al. 2007), and diversity in forested ecosystems have (Chung, Zak, and Lilleskov 2006; Parrent, Morris, and Vilgalys 2006). The inconsistent and often insignificant responses of soil microbial biomass and composition to elevated [CO2] may be due to the high spatial and temporal variation in soil microbial communities within and among systems as well as the variability in plant substrate responses to elevated CO2. Those inconsistencies, and a lack of data from wetland systems, make it hard to predict either how the size and composition of microbial communities will change with increased [CO2] in salt marshes, or how this would affect C residency time. However, there are indications that a reduction in marsh plant diversity and substrate complexity could alter fungal community composition and biomass (Clipson et al. 2006) that may have unknown C residency time feedbacks. RESPIRATION

Another way elevated [CO2] could affect microbes is by increasing their activity. Increased organic C substrates, owing to greater plant productivity, typically leads to greater CO2 respira216

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tion rates. Unfortunately, microbial respiration is usually not directly measured because it is difficult to partition from total soil respiration, which includes both heterotrophic respiration by microbial decomposers and autotrophic respiration by roots. Where increases in soil respiration are measured, it is impossible to know whether the increase is due to enhanced rhizosphere respiration or enhanced decomposition. The latter has direct implications for the C storage potential of salt marsh ecosystems. Some studies have partitioned total soil respiration into microbial and root respiration (Hanson et al. 2000; Schuur and Trumbore 2005). In a study involving maple tree seedlings, microbial respiration was not affected by increased [CO2], but root respiration was (Edwards and Norby 1999). Conversely, in a semiarid grassland system subjected to elevated [CO2], microbial decomposition rates doubled, while rhizosphere respiration remained constant (Pendall et al. 2003). In a Douglas fir mesocosm where heterotrophic respiration was itself partitioned, elevated [CO2] increased litter decomposition but not soil organic matter decomposition (Lin et al. 1999). Clearly, there are site-specific effects, and a detailed study, which partitions soil respiration in salt marshes, is needed. Regardless of its cause, increased soil respiration can potentially offset the increased C storage caused by higher plant productivity under elevated [CO2] and potentially act as a positive feedback to atmospheric [CO2]. Unlike the uncertain microbial biomass and community responses to elevated [CO2], total soil respiration generally increases (Pendall et al. 2004). In a review of forty-seven studies, there was only a single instance of declining soil respiration rates, although many increases were not statistically significant (Zak et al. 2000). After eight years of CO2 enrichment in the SERC marsh, soil respiration rates measured over six weeks during the growing season were enhanced by 15 percent relative to the respiration in the ambient [CO2] treatment (Ball and Drake 1998). This significant increase was observed in all three plant communities (S. olneyi,

S. patens, and mixed) and sampling dates. Due to the difficulty and expense of partitioning respiration sources (Trumbore 2006), it remains unclear whether the increase was due to enhanced rhizosphere or microbial activity. During that time, root biomass in S. olneyi plots had increased by 83 percent under elevated [CO2], but there was no such increase in S. patens plots (Ball and Drake 1998), suggesting the increase in SERC soil respiration was not a root response. A later study (begun after eleven years of CO2 enrichment and carried out for twenty-one months) in the SERC marsh measured microbial respiration during the winter when plants were presumed dormant. Winter respiration rates were generally higher in the S. olneyi chambers subjected to elevated [CO2], but this result was only significant during the winter of 1998–1999 (Marsh et al. 2005). In contrast, there were no increases in the winter respiration rates of the S. patens chambers. Because the winter respiration rates were only significantly greater during one year and because the soil respiration study only lasted six weeks, we regard these results as tentative and suggest that long-term soil respiration measurements are needed at the SERC marsh in order to understand whether increases in this important variable are a result of interannual variability or are part of a long term trend with implications for ecosystem C storage. Under highly reducing conditions, microbes respire methane (CH4) instead of CO2. Methane is a greenhouse gas of particular concern because its capacity to absorb and reradiate infrared energy in the atmosphere is far greater than that of CO2 (IPCC 2001). A week-long study in the SERC marsh found that methane emission was 80 percent higher in S. olneyi plots subjected to elevated [CO2] than in plots under ambient [CO2] (Dacey, Drake, and Klug 1994). Methane flux rates were 3.54 mmol m⫺2 day⫺1 under elevated CO2 and 1.97 mmol m⫺2 day⫺1 in the ambient [CO2] plots. However, the most recent SERC study found no [CO2] effect on CH4 emissions measured over two growing seasons in either S. olneyi– or S. patens–dominated

chambers. although there were sporadic increases in methane emissions in S. olneyi chambers (Marsh et al. 2005). This temporal discrepancy illustrates the need for long-term monitoring of soil gas fluxes. For example, in a P. australis–dominated wetland, radiative forcing effects by the wetland could be regarded as a source of greenhouse gases (i.e., CH4) if evaluated on short time scales or as a sink if evaluated over longer time scales (about one hundred years; Brix, Sorrell, and Lorenzen 2001). In the SERC marsh, it was estimated that increased [CO2] in the S. olneyi chambers stimulated CH4 emissions by three orders of magnitude less than it stimulated CO2 respiration (Marsh et al. 2005). Consequently, the contribution of CH4 flux to total C loss was judged insignificant. In other wetland soils, CH4 emissions consistently increased when subjected to elevated [CO2]. These included a study subjecting bald cypress (Taxodium distichum) and an emergent macrophyte (Orontium aquaticum) microcosm (Vann and Megonigal 2003). Relative to ambient [CO2] treatments, CH4 emissions increased 136 percent in tidal freshwater swamp soils planted with O. aquaticum in growth chambers under elevated [CO2] (Megonigal and Schlesinger 1997). In contrast, elevated [CO2] increased CH4 production but not emission in a study of submerged soil microcosms without plants (Cheng, Chander, and Inubushi 2000). Where increased CH4 emissions occur, they are likely a result of greater plant production under elevated [CO2]. Increased organic matter inputs from the more productive plants enhance microbial respiration (i.e., Marsh et al. 2005), which can include methanogenesis. Also, larger root systems have more surface area for CH4 to flux from soils, through plants, and into the atmosphere (Dacey et al. 1994). The discrepancies between the freshwater wetland studies and the SERC study are most likely due to the abundance of sulfate in salt marshes, a more thermodynamically favorable electron acceptor than H2, acetate, or CO2, the acceptors involved in methanogenesis. However, studies need to be carried out over an entire marsh with its

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range of salinities (and therefore sulfate concentrations) to investigate whether significant CH4 emissions occur under elevated [CO2] in low-salinity areas. LITTER DECOMPOSITION

Increased [CO2] can also affect belowground C cycling in salt marshes by altering the lability and availability of microbial substrates derived from plant litter. Two of the main factors controlling litter decomposition rates are the C:N ratio and lignin content of the litter (Aber and Melillo 2001), collectively referred to as litter quality (Chapin et al. 2002). The C:N ratios and lignin content increased in many plants grown under elevated [CO2] (Ball 1997). These increases could lead to a decrease in plant litter quality, a slower decomposition rate, and an increased C storage, assuming no changes in the amount of plant litter being exported from the salt marsh. A meta-analysis of studies testing litter quality under CO2 enrichment found significantly decreased N content and significantly increased lignin contents (Norby et al. 2001), but these effects were rarely significant for individual studies (e.g., Hirschel, Korner, and Arnone 1997) and did not show a consistent decrease in decomposition rates as would be expected based solely on litter quality (Norby et al. 2001). In the SERC marsh, the C:N ratio of S. olneyi litter increased from 86:1 under ambient [CO2] to 105:1 under elevated [CO2], while there were no significant changes in S. patens litter (77:1 ambient vs. 81:1 elevated; Ball and Drake 1997). The lignin content of both species was unaffected by greater [CO2]. The different physiologies (C3 vs. C4) of S. olneyi and S. patens, and the differing productivity responses to elevated [CO2], may explain why S. olneyi C:N ratios were affected more than those in S. patens (Ball 1997). The increased C:N ratio of S. olneyi could be due to multiple responses in S. olneyi leaves under elevated [CO2] concentrations. Rubisco enzymes contain the majority of leaf N (Lambers et al. 1998), and a reduced leaf Rubisco concentration in S. olneyi has been observed as an acclimatory response to elevated 218

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[CO2] (Jacob, Greitner, and Drake 1995). Therefore, the reduced Rubisco concentration in S. olneyi exposed to elevated [CO2] was likely responsible for the reduced leaf N concentration (Rasse et al. 2003). In a short-term (six-week) decomposition study using litters from each of the three SERC plant communities, there was a significant decrease in soil respiration in only the S. olneyi litter grown under elevated [CO2]—corresponding to predictions based solely on C:N ratios. This indicates that in salt marshes the effect of elevated [CO2] on short-term litter decomposition is species- or perhaps C3/C4-dependent, although the long-term decomposition trends remain unknown. Long-term decomposition studies are needed because seemingly insignificant differences in litter quality like the fourpoint increase in S. patens C:N under elevated [CO2] may amplify over time and because larger differences in C:N like those found in S. olneyi after six weeks may be dwarfed by annual variability (Norby et al. 2001). SOIL CARBON POOLS

The combined effect of increased soil respiration, changes in methanogenesis, and differing litter lability in soils is a change in the size and turnover rates of soil C pools. If the C fixed due to increased plant productivity ends up in slowturnover, recalcitrant pools, then salt marshes will have a net negative feedback on atmospheric [CO2] (Jastrow et al. 2005). If, however, the fixed C ends up in fast-turnover, labile pools or if the newly fixed C “primes” the decomposition of preexisting recalcitrant pools (Kuzyakov, Friedel, and Stahr 2000), then salt marshes will likely have a net positive feedback on atmospheric [CO2]. Soil C pools with relatively rapid turnover include soil microbial biomass, fine root biomass, and dissolved organic C (DOC), while pools with slow turnover include lignin, soil organic matter associated with clay aggregates, and soil humus (Giardina et al. 2005). Unfortunately, the longer the turnover time of a C pool, the longer a study must be carried out to detect changes. Most elevated [CO2] experiments

are not of significant duration to quantify changes in long-term soil C pools with subsequent feedbacks to atmospheric [CO2]. In the SERC marsh, fast-turnover porewater C pools, including dissolved inorganic C (DIC), DOC, and dissolved CH4, were measured after eleven growing seasons of CO2 enrichment (Marsh et al. 2005). Elevated [CO2] increased porewater DIC, which was positively correlated with root biomass. Porewater DOC and dissolved CH4 were greater in the S. olneyi stand under elevated CO2, but these effects were not significant. Increases in DIC, DOC, and dissolved CH4 indicate that a portion of the C fixed by increased primary productivity ends up in C pools that turn over quickly, essentially speeding up the C cycle. Similarly, rapidly cycling soil C pools (fine roots and microbial biomass) significantly increased in Californian grasslands after four years of CO2 enrichment (Hungate et al. 1997). If not exported to coastal oceans, the increased C in fastcycling pools in the SERC marsh will be returned to the atmosphere, a positive feedback on atmospheric [CO2] (Marsh et al. 2005). Because the SERC marsh is the world’s longest-running enriched atmospheric [CO2] experiment, it is an ideal system to study changes in long-term soil C pools and total net ecosystem C balance (Chapin et al. 2006). However, as of this writing, there are not yet published results on changes in long-term or even total (which includes both short- and longterm) soil C pools in the SERC marsh. The fate of enhanced C fixation by plants in the SERC marsh can only be inferred from the results of studies in other ecosystems, which are of shorter duration. A meta-analysis demonstrated that total soil C pools significantly increased in ecosystems subjected to elevated [CO2] (Luo et al. 2006), but the results were unclear as to how the increase was partitioned among fast and slowly decomposing pools. Increases in labile C pools would only lead to a short-term increase in C sequestration and may lead to a decrease in recalcitrant C pools due to increased microbial activity. This “priming” concept was demonstrated in an C3/C4 grassland where the total C

pool did not change under elevated [CO2] because as C increased in labile pools, C decreased in recalcitrant pools, possibly as a result of microbial N limitation (Gill et al. 2006). Similarly, “new” (likely labile) C inputs to soil increased, and “old” (likely recalcitrant) soil C pools decreased under elevated [CO2] in a controlled experiment using grassland plants, a result modulated by N availability (Louseau and Soussana 1999). In addition, elevated CO2 was recently shown to accelerate the decomposition of older soil C in a Florida scrub community as fungal abundance increased (Carney et al. 2007). These studies demonstrate the importance of partitioning soil C pools when investigating C sequestration under elevated [CO2] and also demonstrate how other variables such as nutrients and microbial communities may ultimately determine the C storage potential of soils under elevated [CO2].

PROGRESSIVE NITROGEN LIMITATION AND NITROGEN LOADING Whereas CO2 fertilization increases S. olneyi growth in salt marshes, other elements are also required for plant growth. Barring a change in nutrient use efficiency of a plant, increased C fixation and biomass production requires increasing supply of mineral nutrients. Without an increase in mineral nutrient supplies, the trend of greater C fixation in C3 plants in response to increasing atmospheric [CO2] may cease. This phenomenon has become formalized as the progressive nitrogen limitation (PNL) hypothesis (Luo et al. 2004). The PNL hypothesis states that as more C and N are sequestered in living biomass and organic matter in response to increased [CO2], plant-available N becomes increasingly limiting to plant growth, so long as there are no additional N inputs or reductions in N outputs. As the N requirements of plants released from C limitation goes up, plant-available N supplies in soils and sediments may be decreasing (Luo et al. 2004). Available N pools can decrease under increased C fixation because (1) more N is held up in plant

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biomass, and (2) when plants senesce, a portion of their biomass N becomes sequestered in soil organic matter over the long term. In salt marshes, even less of this organic N may remain available because of slow mineralization under anoxic conditions and continued loss of plant material through tidal action (Raymond and Bauer 2001). Thus, PNL is a negative feedback to continued plant C fixation under increased atmospheric [CO2]. While PNL is generally attributed to N limitation, limitations of other mineral nutrients are also possible. Progressive N limitation is constrained by biological stoichiometry. Plants need certain ratios of C:N, C:P, or C to any other required nutrient in their biomass to maintain their physiological functioning (Sterner and Elser 2002). The most well-known example of biological stoichiometry is the Redfield ratio, 106:16:1, which is the ratio of C:N:P found in marine phytoplankton (Valiela 1995). Stoichiometric ratios vary across different species and different parts of a plant, but they are not very fluid within species or individual plant components. For instance, S. patens shoots have a C:N ratio of 48 (Matamala and Drake 1999), while their roots have a C:N ratio of 62 (Curtis et al. 1990). Therefore, for every additional 10 grams of C in shoots caused by increased productivity, S. patens requires 0.208 grams of N that must be supplied from the soil. Under increased atmospheric [CO2], C3 salt marsh plants have a high potential to experience PNL because N is already limiting primary production in many coastal systems (Vitousek et al. 1997). The addition of N fertilizer has been shown to increase the growth of temperate salt marsh plants both in the United States (Sundareshwar et al. 2003) and in Europe (Wijnen and Bakker 1999; Kiehl, Esselink, and Bakker 1997). While it seems counterintuitive that elevated [CO2] would increase C fixation if salt marsh plants are already N limited, there are ways that plants can prevent limitation in the short term. Such mechanisms include increasing their N use efficiency (Drake et al. 1997) by recovering greater amounts of N prior 220

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to senescing leaves or by increasing the C:N ratios of their tissues (as much as physiologically possible), and increasing C investment in roots to gain greater access to soil N (Luo et al. 2004). A meta-analysis of systems exposed to elevated [CO2] showed that C:N ratios of roots and shoots increased by 11 percent relative to ambient [CO2] (Luo et al. 2006). Root:shoot ratios also increased, implying that the plants were expanding their root systems to gain greater N (Luo et al. 2006). Throughout the 2004 SERC marsh growing season, S. olneyi and S. patens both had significantly reduced root and shoot tissue N concentrations under elevated [CO2], thereby increasing their C:N ratios (Erickson et al. 2006). These mechanisms are only shortterm fixes, and if, as PNL predicts, N becomes sequestered in slow-turnover organic matter pools, N limitation will eventually be more severe than at original [CO2]. Other organisms within salt marsh ecosystems may also experience progressive limitation of mineral nutrients. Soil bacteria were primarily P limited in a pristine South Carolina salt marsh indicated by a positive response of bacterial numbers, production, and activity in response to P fertilization in the field and lab (Sundareshwar et al. 2003). If this P limitation increased under elevated [CO2], owing to more P tied up in plant organic matter, bacterial mineralization of N and P would slow down, making nutrient limitations more severe. In addition, a positive effect of microbial nutrient limitation could be a reduced decomposition of plant residues and an increased C residency time. As of now, it is unclear how P limitation might affect N inputs and outputs such as biological N2 fixation and denitrification in salt marshes. In a laboratory experiment, potential denitrification rates were higher under low-P conditions (Sundareshwar et al. 2003). Progressive nitrogen limitation can be delayed or prevented if N inputs to a system are increased or if N losses to a system are decreased. In salt marshes N inputs are from atmospheric deposition, N2 fixation (Kaplan, Valiela, and Teal 1979), runoff from land, groundwater flow

(Valiela et al. 1978), and tidal flooding (Valiela et al. 1978). Increased C fixation, as a result of elevated CO2, can directly affect N2 fixation rates because N2 fixation consumes large amounts of C and energy. Increased fixation of C means that the plants have more energy and C to give to associated N2-fixing microbes, and there is more organic substrate for heterotrophic, freeliving N2 fixers. Increases in N2 fixation with increased CO2 have been demonstrated in many non–salt marsh studies, especially when non-N nutrients are supplied (Reich, Hungate, and Luo 2006). Nitrogenase (the main enzyme involved in N2 fixation) activity and 15N incorporation into the plants and soils were measured in monotypic stands of S. olneyi and S. patens in the SERC marsh after four months of growth in elevated CO2 (Dakora and Drake 2000). Nitrogen fixation in both plant-associated and free-living diazotrophs (N fixers) was stimulated by elevated [CO2], and the effects were greater in C3 S. olneyi stands than in C4 S. patens stands. If this increase in N2 fixation were sustained at a sufficient magnitude, then this mechanism could prevent PNL, a possibility likely dependent on the dominant plant species. In salt marshes, the major N losses from the system are through tidal export (Valiela et al. 1978) and denitrification (Kaplan et al. 1979). After eight years in S. olneyi–dominated SERC marsh, it was found that growing season potential denitrification rates were lower in elevated [CO2] treatments than in ambient [CO2] treatments (Matamala and Drake 1999); however, this effect was only significant in October. Denitrification may have been limited by nitrate (NO3–) availability because soils under elevated [CO2] had lower amounts of ammonium (NH4⫹), which meant that less NH4⫹ was transformed into NO3⫹ via nitrification (Matamala and Drake 1999). If denitrification rates are significantly lower in marshes subjected to [CO2], PNL may be prevented or delayed. Because N fixation and denitrification can impact PNL, more research is needed on how these microbial processes respond to elevated [CO2] and the associated changes in plant production.

Nitrogen pollution in coastal areas may also ameliorate PNL. Humans have altered the global N cycle significantly, predominantly through industrial N fixation for agriculture (Galloway and Cowling 2002). The import of excess N via atmospheric deposition or runoff (N loading) has led to eutrophication of many estuaries (Vitousek et al. 1997; Parsons et al. 2006), including the Chesapeake Bay (Kemp et al. 2005). This excess N ends up in the tissues of primary producers. For example, sources of estuarine macrophyte-N were found to be derived from terrestrial sources at high N loading rates (Martinetto et al. 2006). Under elevated [CO2], additional anthropogenic N loading may sustain increased productivity over the long term until another nutrient becomes limiting. Nitrogen loading, however, can also lead to accelerated C losses from sediments (Morris and Bradley 1999), further complicating predictions of total C sequestration potential in a given marsh. Eutrophication can also affect whether PNL occurs by changing the dominant plant species. Invasion of many Atlantic coastal marshes by Phragmites australis has been partially attributed to coastal eutrophication (Bertness et al. 2002). Phragmites australis has both a higher N mobilization and uptake capacity than S. patens (Windham and Ehrenfeld 2003). If PNL does occur under increased [CO2], it is unclear whether the high N needs of P. australis will put the species at a competitive disadvantage counteracting eutrophication effects or if the high soil N mineralization rates associated with P. australis will delay PNL for that species, providing another competitive advantage. If the competitive superiority of P. australis is sustained and their growth stimulated in response to elevated [CO2] due to their C3-like physiology, the nitrogen and carbon cycles within a salt marsh could change greatly. These changes could accelerate as CO2 levels climb, shoreline development continues, and natural eutrophication buffers are removed (Silliman and Bertness 2004). As of this writing, there have not been any published direct tests of PNL in salt marsh ecosystems; however, recent SERC experiments

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FIGURE 11.5 A new generation of experiments in the SERC marsh. The larger chambers are designed for measuring not only elevated atmospheric [CO2] responses but also the interaction with elevated N deposition. Factorial treatments using both N and CO2 will enable researchers to tease apart the influences of each factor individually and in concert. Photo courtesy of Adam Langley and Patrick Megonigal, Smithsonian Environmental Research Center.

are aimed at examining the interaction of elevated [CO2] and N deposition (fig. 11.5). In order for the PNL hypothesis to be substantiated in an ecosystem, three things must occur: (1) there must be an initial stimulation of primary production when the system is subjected to increased CO2 (release of limitation), (2) there must be a decline of that initial stimulation over time (limitation renewal), and (3) the availability of N must decline because the ecosystem can no longer redistribute internal N to available pools and is not receiving enough new N inputs (N reduction; Hungate et al. 2006). In the only longterm study of salt marsh response to [CO2], only two of these prerequisites have been shown: (1) the initial increase in primary production, especially in the S. olneyi plants (Rasse et al. 2005) due to their release from C limitation, and (2) the reduction in available soil N after eight years (as measured by exchangeable NH4⫹) in both S. olneyi and S. patens communities (Matamala and Drake 1999). In S. olneyi dominated plots, exchangeable N was 28 to 55 percent less than in ambient [CO2] treatments, and 10 to 30 percent less in S. patens–dominated plots. This decrease was not monitored for long, however. Although it was also documented the following year in the S. olneyi dominated plots, it 222

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was not measured in the S. patens plots. Despite the observed decrease in N availability, increased primary production in S. olneyi marshes, as measured by daytime NEE, plant biomass, and shoot density, has been sustained for over eighteen years (this chapter; Rasse et al. 2005, Erickson et al. 2006), suggesting that this particular salt marsh has not yet been subject to PNL, though the mechanisms remain unclear. Additions to the salt marsh soil N pool could come from the documented increases in N fixation (Dakora and Drake 2000), decreases in denitrification (Matamala and Drake 1999), or N loading. However, these studies of specific microbial N processes were only short term, so there is still much research to be done on microbial-mediated N cycling within the SERC marsh. The sustained primary productivity increase may also be due to the relatively open N cycle of tidal marshes where rivers and tides can deliver available N (Erickson et al. 2006) or seasonal relief of PNL, such as when P is more limiting to primary production in the spring (Conley 2000). Therefore, it may be a combination of factors, such as increased in situ biological N fixation and greater allochthonous N inputs, that contribute to relieving PNL in S. olneyi communities. More studies in a variety of salt marshes are needed to see how PNL affects C storage potential of coastal marsh ecosystems. The questions that remain regarding whether PNL occurs under increased atmospheric [CO2] are not unique to salt marshes. Progressive nitrogen limitation has not been proven yet in a number of different ecosystems. A meta-analysis of 104 studies from forest, grassland, desert, and wetland ecosystems assessed whether PNL reliably occurs with CO2 enrichment (Luo et al. 2006). Wetlands were found to have lower N accumulations in plant and soil pools than forests or grasslands, but it remains unclear if this was a result of plant functional differences or greater wetland plant nutrient use efficiency (NUE) under elevated [CO2]. The overall results were inconclusive due to extreme variability among methods, organisms, and time scales but hinted that many ecosystems, especially those

undergoing a more realistic gradual, rather than abrupt stepwise, [CO2] increase may be able to slowly adjust to PNL by accumulating enough N in plant and soil pools to at least temporarily support increased primary production. However, sequestration of N into organic pools may cause N limitation if those pools have longer turnover times than the pools from which the N was originally immobilized.

CONCLUSION The sustained growth stimulation of S. olneyi over eighteen years, even with salinity fluctuations and increased sea level, suggest that S. olneyi can better adapt to increasing atmospheric [CO2] than other co-occurring salt marsh grasses such as S. patens and D. spicata and will thus come to dominate the SERC marsh site with time. This biomass response is in agreement with what is expected based solely on the differential responses of C3 versus C4 plants, yet it seems to circumvent PNL owing perhaps to greater N inputs from upland ecosystems and potentially greater in situ biological N fixation

under elevated [CO2]. Ultimately, however, it remains difficult to predict which species will eventually come to dominate coastal salt marsh ecosystems (Penning et al. 1992; Bertness and Ewanchuk 2002) without a reliable prediction of (1) how soil accretion will keep pace with rising sea level (Morris et al. 2003), (2) how drought periodicity will fluctuate in response to global climate change, (3) how temperature increases will affect species differentially, (4) how trophic interactions will exhibit controls on plant growth, and (5) how future land use changes will alter coastal ecosystems. Salt marshes may generate a negative feedback to rising atmospheric [CO2] by fixing, and subsequently storing, more C in the future, but changes to soil C storage generally function over long time scales. Using the SERC marsh results to estimate the C sink capacity of marshes under rising [CO2] indicates that C storage of C3 S. olneyi communities will increase, but C storage of C4 S. patens communities will not (table 11.2). However, the general applicability of these results remains uncertain. To further constrain the possibility of coastal

TABLE 11.2 Summary of elevated CO2 effects on ecosystem carbon processes in the SERC marsh and their effect on atmospheric CO2 concentrations

Effect

Parameter

Direction of Responsea

C fixation

C3 NEEdaytime C3 shoot biomass C4 shoot biomass C3 root biomass C4 root biomass C3 litter quality C4 litter quality C3 litter decomposition CO2 respiration CH4 respiration DIC pool DOC pool SOC pool

c* c* 4 c* 4 T* ↔ T* c* c c c Unknown

C mineralization and storage

Feedback on Atmospheric CO2 Concentrationsb ⫺ ⫺ None ⫺ None ⫺ None ⫺ ⫹ None ⫹ ⫹ Unknown

a Strength of response was determined by its significance. A c indicates a positive response, T indicates a negative response, *indicates the response is statistically significant, and 4 indicates a neutral response. b A negative sign indicates that the feedback is negative, thereby decreasing CO2 concentrations, while a positive sign indicates that the feedback is positive.

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ecosystems to act as long-term C sinks, we provide the following recommendations to guide future climate change research in salt marsh ecosystems, which are particularly relevant to current and future studies subjecting salt marshes to elevated atmospheric [CO2]: • Continue elevated [CO2] treatments, and expand to other salt marsh ecosystems differing in dominant C3 and C4 plant species compositions, including the broadly distributed S. alterniflora and the invading P. australis. • Include interactive effects in future elevated [CO2] studies such as grazing by insects, snails, and geese to create more realistic scenarios of future environments. • Expand measurements of NEE to include all respiratory losses, which will constrain the net C balance in these ecosystems. • Study long-term decomposition and total soil C pools in salt marsh ecosystems so that net ecosystem C balance can be determined. • Study all potential C losses from salt marsh ecosystems including DOC, DIC, and tidal transport of particulate C to the open ocean. • Begin measurement of soil accrual rates, which are vital to determining the sequestering potential of salt marsh ecosystems and to predicting competitive outcomes of salt marsh plant species. • Map changes in the distribution of C3 versus C4 plant communities among coastal salt marsh ecosystems so that competitive displacement can be monitored and coupled with other environmental variables and plant physiologies. • Monitor salinity changes and nutrient loading sources in order to assess modifiers of competitive interactions and preemptors of PNL. • Investigate changes in in situ nitrification, denitrification, and N mineralization of coastal salt marshes in more detail so that mechanisms preempting PNL can be assessed. 224

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M. Hay (eds.), Marine Community Ecology. Sunderland, MA: Sinauer. Pennings, S. C., and R. M. Callaway. 1992. Salt marsh plant zonation: The relative importance of competition and physical factors. Ecology 73: 681–690. Pennings, S. C., M. B. Grant, and M. D. Bertness. 2005. Plant zonation in low-latitude salt marshes: Disentangling the roles of flooding, salinity and competition. Journal of Ecology 93: 159–167. Pennings, S. C., and B. R. Silliman. 2005. Linking biogeography and community ecology: Latitudinal variation in plant–herbivore interaction strength. Ecology 86: 2310–2319. Poorter, H., and M.-L. Navas. 2003. Plant growth and competition at elevated CO2: On winners, losers, and functional groups. New Phytologist 157: 175–198. Rabenhorst, M. C. 1995. Carbon storage in tidal marsh soils. In R. Lal et al. (eds.), Soils and Global Change. Boca Raton, FL: CRC Lewis. Rasse, D. P., J.-H. Li, and B. G. Drake. 2003. Carbon dioxide assimilation by a wetland sedge canopy exposed to ambient and elevated CO2: Measurements and model analysis. Functional Ecology 17: 222–230. Rasse, D. P., G. Peresta, and B. G. Drake. 2005. Seventeen years of elevated CO2 exposure in a Chesapeake Bay wetland: Sustained but contrasting responses of plant growth and CO2 uptake. Global Change Biology 11: 369–377. Raymond, P. A., and J. E. Bauer. 2001. DOC cycling in a temperate estuary: A mass balance approach using natural 14C and 13C isotopes. Limnology and Oceanography 46: 655–667. Reich, P. B., B. A. Hungate, and Y. Q. Luo. 2006. Carbon–nitrogen interactions in terrestrial ecosystems in response to rising atmospheric carbon dioxide. Annual Review of Ecology, Evolution, and Systematics 37: 611–636. Sage, R. F. 2002. Variation in the kcat of Rubisco in C3 and C4 plants and some implications for photosynthetic performance at high and low temperature. Journal of Experimental Botany 53: 609–620. Saltonstall, K. 2002. Cryptic invasion by a non-native genotype of the common reed, Phragmites australis, into North America. Proceedings of the National Academy of Sciences of the USA 99: 2445–2449. Schlesinger, W. H., and J. Lichter. 2001. Limited carbon storage in soil and litter of experimental forest plots under increased atmospheric CO2. Nature 411: 466–469. Schuur, E. A. G., and S. E. Trumbore. 2006. Partitioning sources of soil respiration in boreal

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black spruce forest using radiocarbon. Global Change Biology 11: 165–176. Silliman, B. R., and M. D. Bertness. 2004. Shoreline development drives invasion of Phragmites australis and the loss of plant diversity on New England salt marshes. Conservation Biology 18: 1424–1434. Silliman, B. R., J. van de Koppel, M. D. Bertness, L. E. Stanton, and I. A. Mendelssohn. 2005. Drought, snails, and large-scale die-off of southern U.S. salt marshes. Science 310: 1803–1806. Sterner, R. W., and J. J. Elser. 2002. Ecological Stoichiometry. Princeton, NJ: Princeton University Press. Sundareshwar, P. V., J. T. Morris, E. K. Koepfler, and B. Fornwalt. 2003. Phosphorus limitation of coastal ecosystem processes. Science 299: 563–565. Trumbore, S. 2006. Carbon respired by terrestrial ecosystems: Recent progress and challenges. Global Change Biology 12: 141–153. UNESCO-SCOPE. 2006. The global carbon cycle. UNESCO-SCOPE Policy Briefs. October 2006, No. 2. Paris: Author. Valiela, I. 1995. Marine Ecological Processes. New York: Springer. Valiela, I., J. M. Teal, S. Volkmann, D. Shafer, and E. J. Carpenter. 1978. Nutrient and particulate fluxes in a salt marsh ecosystem: Tidal exchanges and inputs by precipitation and groundwater. Limnology and Oceanography 23: 798–812. Vann, C., and J. P. Megonigal. 2003. Elevated CO2 and water depth regulation of methane emissions:

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PART IV

Die-off, Loss, and Conservation

12

From Climate Change to Snails potential causes of salt marsh dieback along the u.s. eastern seaboard and gulf coasts David T. Osgood and Brian R. Silliman Over the last decade, salt marsh grasses, primarily Spartina alterniflora, have experienced sudden mortality events throughout the eastern United States. The large extent of affected areas and mortality acuteness is unprecedented in U.S. coastal marshes. Coincident with dieback in most areas was severe and prolonged drought, a pattern suggesting a strong, climatic trigger factor. These drought events are thought to have induced extreme stress in marsh soils, including desiccation-induced acidity and accompanying metal toxicity and/or elevated salinities. Evidence for elevated soil salinities has been provided by field data collected at marsh dieback sites both before and during dieback events, while evidence for metal toxicity is supported by laboratory experiments that were conducted after the dieback events occurred. These droughtinduced soil stressors are then thought to have either killed marsh grass outright or suppressed grass growth, making S. alterniflora more susceptible primarily to three plant-killing consumers: fungal pathogens, which occur along the entire eastern seaboard; fungal farming snails, found primarily in southeastern marshes; and sesarmid herbivorous crabs, found in high densities in New England dieback areas. In southeastern marshes, field experimentation has demonstrated that grazing by periwinkle snails contributes significantly to marsh dieback. The magnitude of this topdown impact, however, is strongly dependent on background snail densities (snails can increase dieback areas by 15 to 180 percent), and, at some dieback sites, snails were not present at all. Salt addition experiments in healthy marshes have shown that sublethal but elevated soil salinities can stimulate overgrazing by snails and subsequent fungal infection, resulting in dieback generation in small areas. In New England marshes, plant tethering and caging experiments have shown that crab grazing both prevents recovery and expands marsh dieback areas, implicating increased crab densities as a primary factor contributing to the initiation of marsh dieback in this region. Combined, results from these laboratory and field experiments indicate that multiple factors— from severe soil stress to intense consumer pressure, working synergistically and couched within an interaction cascade initiated by severe drought—were likely the cause of

231

massive marsh dieback. Importantly, the relative importance of these factors in causing marsh dieback likely varied greatly from site to site, depending on site-specific characteristics such as size and location of the dieback event (which varied immensely across the United States); presence and density of snails, crabs, soil salinities, soil acidity, and redox; tidal regime; wind-extended low tide durations; plant species present; and freshwater inflows that could mitigate drought impacts. This recent and widespread salt marsh dieback event may represent another example of background anthropogenic impacts working synergistically with natural stressors to culminate in ecosystem-level degradation (e.g., seagrass, kelp, and coral reef declines; overfishing). Future investigations of marsh dieback must include field experiments designed to isolate the relative role of top-down and bottom-up factors and test for interacting factors, such as soil metal toxicity, predator control of snail and herbivorous crab densities, and the role of fungi in killing off edaphically stressed plants. Management and scientists will need to adapt to the complex reality of cascading, interacting causal factors set off by climatic triggers and not rely on the hope of revealing one “silver bullet” as the sole cause of marsh dieback. Management must also begin to monitor snail densities and those of their predators along with soil salinities, redox, and acidity levels in healthy marshes so that the magnitude of marsh grass stressors can be measured before, during, and after dieback events. In addition to reexamining current regulations in tidal marshes, managers should consider actions such as freshwater releases on managed rivers to coincide with drought and tighter regulation of populations of snail and sesarmid crab predators (e.g., blue crabs).

Globally, coastal ecosystems are experiencing an unprecedented culmination of anthropogenic stressors (Kennish 2001; Harley et al. 2006; United Nations Environment Programme 2006 [UNEP]; Vitousek et al. 1997). These stressors are coincident with a threefold higher human population density along the coastline compared to the rest of the world land area (Small and Nichols 2003). Impacts of an increasing coastal population include pollution, eutrophication, habitat alteration, fisheries exploitation, and freshwater diversions in estuarine systems (Kennish 2001, 2002; Pauly et al. 2003). Additionally, coastal systems are likely to be impacted by the more diffuse but equally strong effects associated with climate change, including increased storm frequency, enhanced sealevel rise, and altered ocean circulation patterns (Harley et al. 2006). Coastal ecosystems provide vital services to human populations. Economies within one hundred kilometers of a coastline account for 61 percent of the global $44 trillion gross national 232

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product (Millennium Ecosystem Assessment 2005), and services from estuaries and associated tidal wetlands are estimated at $32,822 ha⫺1 yr⫺1 (Costanza et al. 1997). When importance to human well-being is coupled with rising threats, the vulnerability to coastal ecosystems and the critical services they provide becomes clear. For these reasons, there is now a worldwide call to allocate greater research and management attention to the conservation of coastal marine systems (Lindholm and Barr 2001; Roberts et al. 2002). In the past century, coastal ecosystems have been exhibiting symptoms of systemwide degradation or have experienced large-scale, speciesspecific mortality events as a result of increasing anthropogenic and nonanthropogenic stressors. In the 1930s, the eelgrass wasting disease resulted in a more than 90 percent loss of eelgrass, Zostera marina, beds throughout the U.S. Atlantic Coast (Ralph and Short 2002). Eelgrass beds have never rebounded to historic levels and still experience dieback (Short, Ibelings, and Hartog 1988). Similarly, a widespread loss of Thalassia

testudinum (turtlegrass) throughout Florida Bay in the late 1980s has been attributed to interactions between increased salinity, sulfides, and potential pathogens (Zieman, Fourqurean, and Frankovich 1999). In 1999, Long Island Sound populations of Homarus americanus (American lobster) experienced mortality rates approaching 99 percent (Valente and Cuomo 2005). Global fisheries landings are declining by hundreds of thousand of metric tons annually from a peak in the 1980s (Pauly et al. 2003). Coral reefs have experienced long-term losses from a variety of stressors that result in reductions of coral cover as high as 80 percent in many regions throughout the Caribbean (Gardner et al. 2003; Pandolfi et al. 1997). Productivity declines in kelp forests have been linked to the impact of human-modified trophic cascade, but they also are influenced over the long term by periodic shifts in oceanographic climate (nutrient availability) and localized anthropogenic stressors such as pollution (Dayton et al. 1998, 1999; Estes et al. 1998). Combined, these studies warn that marine ecosystem degradation brought on by human impacts is occurring at an unprecedented scale and that it is likely to increase further in the near future (Vitousek et al. 1997). U.S. coastal salt marshes have also been subjected to a variety of anthropogenic stressors, including dredging/fill impacts, invasive species introductions, ditch and levee construction, and hydrologic alteration such as drainage or water diversions (Kennish 2001; see also other chapters in this volume). These impacts have been spread out over a relatively long time frame and have led to severe degradation and loss of marsh habitat. Coastal salt marsh losses were at their highest rates historically between the 1950s and 1980s when a total of nearly 249,000 hectares were lost. Loss rates continued, but more slowly (about a twenty-thousand-hectare total loss from the 1980s to 2004; Dahl 2006) due in large part to protective legislation for this specific wetland classification. To date, approximately 50 percent of the original acreage of U.S. coastal salt marshes has been lost (Dahl 2006; also see chap. 13 in this volume).

The surviving 1.57 million hectares of salt marsh are often characterized as being safe from these former threats (e.g., reclamation) and are now generally thought to be among the most resilient and productive of the world’s ecosystems (Mitsch and Gosselink 2000; Dahl 2006). The high levels of primary production in this system are attributed primarily to grasses belonging to the genus Spartina and are facilitated in large part by a regular tidal energy subsidy (Odum, Odum, and Odum 1995) and a trophic cascade, where predators suppress densities of potent grass-grazing snails (Silliman and Bertness 2002). Recent events in Spartinadominated marshes throughout the eastern United States, however, have questioned the limits of these normally robust and productive ecological systems. In spring 2000, patches (of ten to ten thousand square meters each) of dead S. alterniflora appeared suddenly along the Louisiana Gulf Coast (Stewart, Proffit, and Charron 2001). The rapid rate of mortality and the large area impacted suggested that this particular “dieback” event was unlike any previously observed in U.S. coastal marshes. The emergence of prior sudden dieback observations in coastal Connecticut (1999) (R. Zajac and D. Osgood, personal observation) and the Florida panhandle (1990–1995) (Carlson et al. 2001) was an indication that this may not be an isolated event. The dieback episode took on even more urgency when similar patterns and extent of S. alterniflora mortality appeared on additional coastlines across eleven states, affecting at least two hundred thousand hectares (McKee, Mendelssohn, and Materne 2004; Silliman et al. 2005; Ogburn and Alber 2006). Although the phenomenon has mostly been isolated to S. alterniflora, there are also reports of mortality in some, but not all locations, for Spartina patens, Distichlis spicata, and Juncus roemerianus. The ensuing exploration into the recent dieback phenomenon (fueled in part by emergency legislation and multiple interagency efforts) has resulted in a large volume of information mostly conveyed through thematic

causes of dieback in the eastern u.s.

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meetings, Web sites, and gray literature (but see McKee et al. 2004; Silliman et al. 2005; Ogburn and Alber 2006). The information indicates that the patterns of the dieback phenology (total area impacted, spatial patterns within marshes) vary significantly by geographic location, as do the relative importance of the proposed causative and interacting agents, including severe drought, soil toxicity, and runaway consumer impacts. The objective of this chapter is to provide an up-to-date summary on the U.S. salt marsh dieback event starting with a descriptive overview of the current and historic dieback, and a summary of postulated causes. We will also assess the dieback in the context of disturbance ecology, synergistic stressors, and global climate change and the implications these have on remediation efforts and management of the crisis.

BACKGROUND: DESCRIPTIVE OVERVIEW OF CURRENT AND HISTORIC MARSH DIEBACK CURRENT DIEBACK

The current marsh dieback first gained real attention in spring 2000 when reports of sudden mortality of S. alterniflora emerged during flyover surveys in coastal Louisiana (Stewart et al. 2001). The incident quickly became known as “brown marsh,” a term that describes the premature conversion of live biomass to standing dead shoots that came to characterize most of the current dieback events. In the current incidences, mortality of the whole plant (aboveand belowground) within a short time frame (less than a year in many cases) over large areas of marsh indicated that this event was unlike any encountered for the region in recent history (McKee et al. 2004). Shortly after the discovery of the Louisiana dieback, it became obvious that two previously reported dieback events (Florida panhandle 1990–1995 and Branford, Connecticut, 1999) (Carlson et al. 2001; R. Zajac and D. Osgood, personal observation) may be related to the 234

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more recent event. Dieback was subsequently reported in Georgia and South Carolina starting in 2001 (Ogburn and Alber 2006; Silliman et al. 2005) and throughout New England (Rozsa 2006), where the widespread nature of the mortality earned it the phrase “sudden salt marsh dieback” (Adamowitz and Wagner, 2005). Areas of dieback have also been documented in Texas (Edwards et al. 2005), Virginia (Marsh 2007), and Delaware (Bason and Jacobs 2007). The total area of dieback appears to be greatest on the Louisiana Gulf and Georgia coasts. Total marsh area affected by the dieback appears to be lower at other locations, but the mid-Atlantic and Northeast seem to report a greater number of sites experiencing dieback (fig. 12.1). In the initial incidences of current dieback, only one species (S. alterniflora) was experiencing mortality, despite the proximity of other species (S. patens, Avicennia germinans) within or adjacent to dieback areas (McKee et al. 2004) (fig. 12.3A, later). Subsequent dieback, however, included mortality of S. patens, Distichlis spicata, and Juncus roemerianus. These species did not suffer dieback at all reported locations, and the total extent of dieback in these species was dwarfed by dieback in S. alterniflora (table 12.1). Dieback is often extremely patchy within a marsh, with healthy areas being starkly juxtaposed to affected ones. Dieback has occurred in both high and low marsh, along creek banks, and adjacent to the upland (table 12.1; fig. 12.3B). Location of dieback within individual marshes did, however, vary by site (table 12.1). Recovery of the dieback zones also varies by site. Several studies report relatively quick recovery on the order of one to three years (McKee et al. 2004; Ogburn and Alber 2006; Edwards et al. 2005), while some sites report a mix of recovery and continued dieback (Rozsa 2006; Smith 2006). Potential, but noninvestigated, factors controlling rate of recovery likely include timing of cessation of drought, intensity of grazing by snails and crabs along borders, marsh location, marsh zones affected, the extent of positive feedbacks in plant recovery dynamics, and size of initial dieback. In addition, dieback has emerged in

FIGURE 12.1 Locations of major dieback sites along the Gulf of Mexico and U.S. Atlantic coastlines. Size of the point indicated the relative amount of total area impacted in hectares. The year of first reported dieback is indicated, as is the number of reported sites at each locale (in parentheses) where information on the number of sites was available. Sources of information: Adamowicz and Wagner 2005; Bason and Jacobs 2007; Marsh 2007; Michot et al. 2004; Silliman et al. 2005; Smith 2006.

Delaware (Bason and Jacobs 2007) and Massachusetts (Adamowitz and Wagner 2005) in recent years, concomitant with low-rain summers, indicating a continuance of the event. Three primary Web sites are dedicated to the dieback events and are hosted by the U.S. Geological Survey (www.brownmarsh.net), Georgia Coastal Research Council (http://www. gcrc.uga.edu/FocusAreas/marsh_dieback.htm), and the New England Estuarine Research Society (wetland.neers.org). The sites feature meeting summaries, posting locations for reporting dieback, and data repositories including remote sensing information and site-specific mortality information. The Breaux Act Web site (www.LACoast.gov) also has been updated to include maps and remote imagery on the brown marsh phenomenon. These Web sites and accompanying information sharing have no doubt been important in dealing with the marsh dieback, identifying causal factors, and coordinating management decisions in different regions.

The current dieback phenomenon has also resulted in a rapid political response that builds upon the existing Breaux Act (the Coastal Wetlands Planning, Protection, and Restoration Act) enacted in 1990 to mobilize federal funds for preservation and restoration of Louisiana wetlands (www.LACoast.gov). In October 2000, Louisiana governor Mike Foster issued a proclamation establishing a state of emergency in several Louisiana parishes. The proclamation was followed by an executive order (MJF 2000-41), the Saltwater Marsh Dieback Action Plan (http://brownmarsh.net/press/2000-11-07a/ SaltwaterMarshDieOff.pdf). The plan ordered that steps be taken to secure dieback locations, including the possible reintroduction of freshwater from the Mississippi and Atchafalaya rivers into the affected areas. HISTORIC DIEBACK IN THE LOUISIANA REGION

Most of the historical documentation of S. alterniflora dieback stems from the Mississippi causes of dieback in the eastern u.s.

235

n/a

n/a n/a

n/a

1999 (CT), 2002 (MA)

1975

2004

Louisiana

New England

North Carolina

Virginia

Low marsh; panne marsh Low marsh

⬎400 max. and several smaller areas 0.7e

Low marsh; interior marsh

Interior marsh

Levees; levee berms; midmarsh

Location within Marsh

None

None

Juncus roemerianus; Spartina patens; Distichlis spicata

Spartina patens (some areas)

Juncus roemerianus

Other Species Impacted

⬍1 yearf

⬍1 year

⬍1 year

15 monthsd

⬍1 year

Reported Recovery

b

This table summarizes only those areas with a reasonable amount of reported information; “n/a” denotes when specific categorical information was not available. 1 ⫽ excess salinity; 2 ⫽ acidity; 3 ⫽ sulfide; 4 ⫽ grazing; 5 ⫽ pathogen. c Of these, 43,000 were reported as severe. d Minimum recovery time. Overall recovery was variable. e One patch reported. f Some area persisted through 2006.

a

0.3–5,000

100,000c

2000

Georgia

n/a

⬎160 max. and several smaller areas

⬎800

2001

Region/Statea

Reported Patch Size (⫻1,000 m2)

Total Area Impacted (ha)

Year First Reported

TABLE 12.1 Summary of Spartina alterniflora dieback by region or state

n/a

1, 2, 3

1, 2, 3, 5

2, 4

1, 2, 4

Postulated Causesb

Marsh (2007)

Linthurst and Seneca (1980)

Adamowicz and Wagner (2005); Smith (2006)

McKee et al. (2004); Silliman et al. (2005)

Ogburn and Alber (2006)

Citation

Delta region in Louisiana. The reported dieback from this region has been the subject of much research and debate for almost forty years. Marsh loss rates associated with historical dieback range from 2,400 to 9,070 ha yr⫺1 depending in part on the specific classification of coastal marsh (Webb, Mendelssohn, and Wilsey 1995; Day et al. 2000; Dahl 2006). The highest loss rate occurred between 1955 and 1978 (12,700 ha yr⫺1; Turner 1997), contributing to a total of 390,000 hectares of marsh loss as of 1994 (Webb et al. 1995; Day et al. 2000). This type of historical dieback affects areas of intertidal salt marsh and oligohaline marshes, the latter of which comprise about 70 percent of the region’s 1.2 million hectares of tidal wetland (Mitsch and Gosselink 2000). Louisiana marsh dieback areas were first observed in 1968 and described by Smith (1970). The author borrowed the term dieback from the description of mortality events involving S. townsendii in UK marshes. The early patterns of dieback were similar to that documented for the UK marshes—namely, the occurrence of channel dieback adjacent to tidal channels and panne dieback areas in the interior marsh. Smith (1970) summarized several potential causes, including soil waterlogging, lack of available iron, hydrogen sulfide toxicity, and altered tidal flooding. The author concluded that it was likely to be a synergism among more than one factor. Later it became clear that excess tidal submergence and accumulation of the phytotoxin hydrogen sulfide were likely primary mechanisms behind plant mortality associated with the historic dieback as well as an important regulator of macrophyte productivity for both tidal salt and oligohaline systems (Mendelssohn and McKee 1988; Webb et al. 1995; Webb and Mendelssohn 1996; Mendelssohn and Morris 2000). The exact nature and underlying causes behind the excess anaerobosis are still disputed, but there is general agreement that the dieback is set in motion through a complex combination of large- and small-scale hydrologic manipulation (canals, dredging, diversion of freshwater flows), coastal subsidence, and

nutria grazing (Turner 1997; Day et al. 2000; Mitsch and Gosselink 2000). OTHER HISTORICAL DIEBACK

Linthurst and Seneca (1980) use the term dieback to expressly describe areas of S. alterniflora mortality along tidal channels and within interior pannes in marshes of the Cape Fear region of North Carolina. Causal factors were not elucidated in this study, but lower-thanaverage tidal range and low precipitation coinciding with observed desiccation cracks and salt encrustations on the soil surface were cited for at least one of the sites. Additionally, the dry sites exhibited low pH (about 3 to 6) that is consistent with some observations from the recent dieback phenomenon in Louisiana (McKee et al. 2004; see also the later discussion of postulated causes). The relative impacts of grazers were never experimentally tested, even though snails and geese are common in these areas (Silliman and Bortolus 2003). Another similarity between the documented dieback in North Carolina and the current dieback on the eastern seaboard is elevated nitrogen availability in resultant exposed mudflat soils. Soil nitrogen was elevated in dieback areas relative to healthy areas in a Georgia marsh (Ogburn and Alber 2006) and, in the past, has also been shown to increase in areas denuded by intense snail grazing (Silliman and Zieman 2001). High nitrogen availability is likely attributed to a rapid recovery at the prior North Carolina dieback site and likely occurs because the plants that died off are no longer taking up soil nitrogen, leading to a buildup in the soils (Silliman and Zieman 2001). S. alterniflora recovered within two to three years of the dieback onset, and Linthurst and Seneca (1980) concluded that the condition(s) resulting in the mortality were a temporary occurrence for these marshes. S. alterniflora dieback has also been implicated as a contributing factor to marsh loss in Jamaica Bay salt marshes near New York City (Hartig et al. 2002). Jamaica Bay marshes are experiencing widespread degradation resulting from a sediment deficiency exacerbated by boat

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traffic, waterfowl grazing, and dredging (Hartig et al. 2002). This marsh system, and others like it, experience marsh loss or conversion because urban development creates static boundaries (e.g., seawalls and other hardened shoreline features) that are incompatible with landward migration of marsh as sea level rises (Hartig et al. 2002; Shirley and Battaglia 2006). S. alterniflora dieback in particular interior locations and associated loss of sediment-trapping capacity is suspected to act as a positive feedback for overall marsh loss at Jamaica Bay (Hartig et al. 2002). Causal factors of the Jamaica Bay dieback were not reported by Hartig et al. (2002), but sediment deficits as well as intense overgrazing by increasing populations of Canada geese are suspected as interacting, contributing factors (B. R. Silliman and R. Burke, unpublished data). COMMONLY OCCURRING MARSH GRASS MORTALITY EVENTS, NOT TO BE CONFUSED WITH MARSH DIEBACK

Small-scale incidences of plant mortality are common in tidal salt marshes. The most frequent form of plant mortality is the formation of isolated dead patches typically ranging in size from one square meter to tens of square meters. These patches are often generated from wrack deposition in the high marsh; ice scouring in the low marsh; localized subsidence; and invertebrate, avian, or mammalian herbivory (Reidenbaugh and Banta 1980; Bertness and Ellison 1987; Wijte and Gallagher 1991). Reference to these categories of marsh grass mortality date back to some of the earliest marsh literature (Chapman 1960; Miller and Egler 1959; Teal 1962; Adams 1962; Redfield 1972). Often, these mortality patches are recolonized by the same vegetation or replaced by species adapted to the edaphic changes that accompany vegetation loss (Bertness and Ellison 1987). Patches of mortality can also convert to permanent open water pannes (Niering and Warren 1980; Bertness and Pennings 2000). Tidal marsh systems seem to accommodate these recurring events that are important

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factors in shaping the vegetation community structure and the ecology of the marsh as a whole (Niering and Warren 1980; Bertness and Ellison 1987; Bertness 1991). Importantly, once exposed mudflats are generated and dead grass is gone, it is nearly impossible to distinguish these sites from those caused by massive marsh browning and dieback, other than obviously large areas of dieback. As of now, dieback areas are typically identified as areas of extensive brown, dead grass that is upright and still intact and areas that have died relatively quickly, within one to two months. To date, most research has failed to produce conclusive evidence that systematically eliminates these alternative, naturally occurring explanations for generation of exposed mudflats. A more rigorous discussion and experimental investigation of alternative explanations is needed to make an accurate diagnosis of dieback and avoid potential mix-ups with naturally occurring mudflat generation in areas normally inhabited by marsh grass.

EVIDENCE FOR POSTULATED AND DEMONSTRATED CAUSES OF MARSH DIEBACK The literature is replete with studies investigating sublethal and lethal abiotic stresses that impact S. alterniflora growth and historic dieback events (see overview in and Morris and Mendelssohn 2000). The physiological stresses underlying the current dieback events may be similar to those operating during historic dieback. There are, however, very important differences characterizing the present dieback events, including the “spread” of the event throughout the U.S. coastline, the rapidity of the mortality, the potential role of large-scale climatic factors (McKee et al. 2004; Silliman et al. 2005), and the more recently discovered biological stressors (i.e., the interactive, destructive role of consumer fronts—Silliman et al. 2005; see also Silliman and Zieman 2001; Silliman and Bortolus 2003). These differences suggest

(A)

(C)

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FIGURE 12.2 Typical marsh dieback area on Sapelo Island, Georgia. (A) Edge of dieback area with massive snail front and (B) and (C) sequentially close pictures moving into the same spot. In (D), snails are finishing off a small patch of live grass remaining in the middle of the dieback area.

that stressor identity and the manner in which those stressors operate and interact in the present dieback are in many ways unique compared to past events. POLLUTION AND PATHOGENS

In contrast to the Phragmites dieback in the United Kingdom and Europe (summary by Ostendorp 1989), the current marsh dieback in the United States (figs. 12.1–12.4) does not at present appear to directly arise from eutrophication or any acute pollution event (McKee et al. 2004; Silliman et al. 2005). Many coastal systems in the United States are under the influence of cultural eutrophication and experience a variety of other pollutant exposure (Howarth et al. 2000; Kennish 2001). These background

anthropogenic stressors may be playing a role in the overall stress experienced by tidal marsh vegetation communities, but the current consensus holds that eutrophication is not likely to be the solitary or primary precipitating cause of the sudden dieback events in most areas. Past nutrient enrichment studies have shown that increasing nitrogen availability to and in S. alterniflora leads to overgrazing by grazers and, at high nutrient levels, the potential for denuding of the substrate (Silliman and Zieman 2001). There is no current experimental evidence that eutrophication is stimulating overgrazing on a large-scale basis, but recent studies in New England suggest that nutrient-triggered overgrazing by herbivorous crabs and insects has the potential to be an important contributing

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(A)

(A)

(B) FIGURE 12.3 Aerial view of marsh dieback in Chatham County, Georgia, depicting dieback along creek banks and within interior sections of the marsh (A) and marsh dieback along Burrell Creek in Camden County, Georgia, depicting creek bank dieback (B). From http://www.gcrc.uga.edu/ MarshDieback/marsh_photos.htm.

factor to marsh dieback in those regions (M. Bertness et al., unpublished data). To accurately assess the potential role of eutrophication in generating and contributing to marsh dieback, scientists should utilize experiments in the field and subvert traditional and current reliance on correlation studies in marsh ecosystem science to an investigative and explanatory role. To date, laboratory and correlative research suggests that a pathogen is not likely the primary underlying causal factor of marsh dieback, but it is a secondary factor in at least some of the dieback events (Silliman et al. 2005) and likely many more. Although several species of fungus have been isolated from plants collected at dieback sites, two fungal species in particular

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(B) FIGURE 12.4 Aerial view of marsh dieback near Wellfleet, Massachusetts, where dieback appears isolated to interior sections of the marsh (light shade) (A; Smith 2006) and within a marsh on Delaware’s Inland Bays (B) where dieback occurs along creek bank and interior sections (Bason and Jacobs 2007).

appear suspect. Both Fusarium veritcillioides and Rhizoctionia solani have been isolated from soil and plant collections made in Louisiana, Georgia, Florida, Delaware, and Connecticut (Carlson et al. 2001; Ogburn and Alber 2006; Elmer and O’Donnell 2006; Schneider and Usmer 2006; Bason and Jacobs 2007). These occurrences appear to be secondary infections, and it is more likely that pathogens are having a strong but secondary effect on plant growth, with infection being stimulated by external stressors that weaken plant immune defenses, such as snail grazing (Silliman and Newell 2003), drought-induced stress, or a combination of the two (Silliman et al. 2005; McKee et al. 2004).

The negative impacts of fungal infection on S. alterniflora growth as stimulated by external stress have been experimentally demonstrated in the field. In marshes on Sapelo Island Georgia, Silliman and Newell (2003) showed that tissue damage via snail grazing facilitates fungal growth on leaf tissue. The snails subsequently consume the facilitated fungal crop and graze primarily on fungal tissue as opposed to live grass. Nevertheless, the end result of this unique form of plant grazing, as compared to typical, direct consumption of plant tissue by herbivores, is the same for S. alterniflora: severe reductions in plant biomass (about 30 to 80 percent; Silliman and Newell 2003, fig. 2). Current research into the role of pathogenic causes in causing or contributing to dieback events (acting either independently or in concert with other stressors such as grazing) is ongoing. In particular, these investigations are seeking to elucidate the potential role of new pathogens and/or fungi that normally exhibit a nonlethal relationship with plants (e.g., decomposers and mutualists), but under the presence of compounding or novel environmental stressors, they switch to pathogenic lifestyles and attack S. alterniflora (Elmer and O’Donnell 2006; Rosza 2006; Silliman and Newell 2003). Critical for the success of these studies is the use of field studies that experimentally manipulate both fungi and stressor (e.g., soil salinity, metals, redox, snails) presence and examine impacts on grass growth in areas that are both healthy and already affected by marsh dieback stress. SEDIMENT ANAEROBOSIS AND PHYTOTOXINS

Soil anaerobosis and accumulation of phytotoxins associated with waterlogged conditions in the intertidal zone are normally strong regulators of vegetation community composition and primary production of marsh grasses (Morris and Mendelssohn 2000). Prolonged waterlogging and accumulation of hydrogen sulfides are thought to be the primary mechanisms involved in most of the historic dieback along the U.S. Gulf Coast (Mendelssohn and McKee 1988;

Koch, Mendelssohn, and McKee 1990; Webb et al. 1995). In Louisiana marshes, McKee et al. (2004) measured marsh grass vigor (stem density, tissue element concentrations), plant pathogens, and soil physicochemistry within dieback and healthy areas in a variety of Louisiana marshes immediately following first documentation of the recent dieback in 2000. The sampled dieback occurred in interior areas of the marsh, and healthy control locations were situated near the shoreline. The affected areas exhibited 50 to 100 percent mortality and very low viability of rhizome tissue (less than 5 percent in the dieback areas). The presence of standing dead shoots in June 2000 in the dieback areas testified to the acuteness of the mortality event. The primary species affected was S. alterniflora, with some mortality of S. patens. Porewater sulfide levels, although exceeding one-millimole lethal concentrations (Koch et al. 1990), were similar between dieback and healthy zones. McKee et al. (2004) concluded that the present dieback phenomenon in the Louisiana marshes was not related solely to elevated sulfide levels. Similar information from other reported dieback sites also does not seem to support that soil oxygen status and associated soil phytotoxins as primary drivers of the current dieback phenomenon (Carlson et al. 2001; Silliman et al. 2005; Rozsa 2006). DROUGHT-INDUCED SOIL ACIDITY, ION TOXICITY, AND SALINITY

Record drought conditions and low river flow combined to result in record-low water levels in Louisiana marshes just prior to the Louisiana dieback in 2000 (McKee et al. 2004; Silliman et al. 2005). These circumstances led to speculation of a primary role for soil desiccation in dieback areas in Louisiana (McKee et al. 2004). McKee et al. (2004) documented elevated ratios of leaf tissue sodium–potassium in dieback areas and also found post hoc evidence that porewater salinity reached sublethal but not lethal stress levels for S. alterniflora. In Georgia,

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sublethal salinities were documented both during and before dieback events (Silliman et al. 2005). Besides these observed sublethal salinity stresses, evidence for soil drying was accompanied with a drop in pH to about 4 at one site in Louisiana, which likely contributed to pyrite oxidation, a stress that may reduce grass growth or cause mortality (McKee et al. 2004). To test for the potential lethal and sublethal impacts of drought stress on soil acidity, Mendelssohn et al. (2006) asked whether soils could acidify in response to simulated drought stress. Soils from dieback areas were found to acidify upon experimental oxidation (under laboratory conditions) in contrast to treated soils from healthy marsh areas, supporting the idea that drought-induced soil acidity could have initiated marsh dieback (McKee et al. 2004). In addition, Mendelssohn et al. (2006) described lab work that shows the potential for water stress stemming from the preceding drought to interact with soil acidification and metal toxicity to cause mortality in S. alterniflora. Low pH (down to 2.5) was tolerated by culms of S. alterniflora in experimental pots. However, when metal toxicity (aluminum, iron, zinc, and manganese) and water stress were combined with the low pH, mortality in S. alterniflora increased by 80 to 90 percent over the pH stress acting alone. In addition, the same treatment for Avicennia germinans and S. patens did not yield higher mortality. These same species survived within and adjacent to dieback areas in the field (Mendelssohn et al. 2006; McKee et al. 2004), suggesting these interacting factors could have been primary stress forces acting to initiate Louisiana dieback events. We can assume that soil desiccation and physiological water stress are relatively uncommon events (e.g., Anisfield and Benoit 1997) in tidal marshes given that most marshes are regularly flooded. Therefore, the novel incidence of desiccation associated with climatic extremes could fit the acuity of the recent dieback event. The relatively rapid recovery rates for some sites (McKee et al. 2004; Ogburn and Alber 2006; Edwards et al. 2005) are also consistent with the 242

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acute nature of a soil acidification episode. Soil desiccation and a concurrent drop in soil pH have been reported for former dieback episodes as well where recovery was also relatively rapid (Linthurst and Seneca 1980). Caution should be exercised, however, if applying an umbrella, causal role of soil acidification and metal toxicity across all dieback sites, as some dieback sites did not coincide with drought episodes, and many sites did not experience prolonged exposure of soils to desiccation stress (Carlson et al. 2001; Rozsa 2006; Smith 2006). Caution should also be taken in accepting these results as a “silver bullet” for dieback causation in Louisiana and other sites since both Mendelssohn et al. (2004) and McKee et al. (2004) did not employ field experiments. Instead, field correlations and lab experiments were used to test hypotheses. In addition, given the strong evidence these studies provide that soil acidity could have occurred during Louisiana dieback, but also the lack of strong results indicating that levels such as these were reached in the field, it is likely that sublethal soil acidity levels interacted with other droughtinduced stresses such as sublethal salinity levels and increased fungal infection and snail grazing (see later discussion) to initiate marsh dieback. In the end, to truly assess the relative importance and impacts (e.g., percent of negative impact and order of contribution) of these factors, researchers should (1) collect descriptive soil and biological data from affected and unaffected areas before, during, and after dieback events; (2) manipulate proposed causal factors in the field as dieback is occurring; and (3) employ large-scale experiments in the field in healthy marshes that manipulate all or some of these factors in fully factorial designs.

TOP-DOWN FORCES

Top-down control of plant community structure is now a foundational principle in ecology and has been shown to be important in a variety of ecosystems, including savannahs, kelp beds, seagrasses, mangroves, grasslands, forests, and

algal beds on coral reefs (Krebs 2006; see chap. 6, this volume, for an expanded discussion of top-down forces). In these systems, grazers reduce plant growth and biomass and, when left unchecked, can overgraze plant communities, leading to wholesale reduction in plant cover. In contrast to ecological theory and experimental evidence in these systems, marsh plant communities have long been held to be the quintessential example of a plant ecosystem controlled overwhelmingly by bottom-up forces (e.g., nutrients, salinity). Early studies that examined top-down control in marshes found them to be inconsequential to plant community dynamics and concluded that grazers played no significant role in controlling plant community structure and primary production (Teal 1962; Odum and De La Cruz 1967). These early studies, however, lacked experimentation testing the hypothesis that plant consumers were unimportant by caging out marsh grazers. After forty years of domination of this bottom-up paradigm in the scientific and estuarine education fields, ecologists began to exclude grazers and found that rather than having minimal impacts in salt marshes as dogma espoused, grazers were exerting strong top-down control that, at times, could lead to complete denuding of marsh substrate. Exclusion experiments have repeatedly demonstrated that marsh grazers, such as snails, crabs, and geese, commonly exerted strong top-down control of marsh plant growth that occurred both locally and across large scales at a variety of geographic locales (Bazely and Jefferies 1986; Jefferies and Rockwell 2002; Siiliman and Zieman 2001; Silliman and Bortolus 2003; Silliman et al. 2005; Jefferies et al. 2006). Indeed, the once largest, most continuous salt marsh in the world in Hudson Bay has now been completely destroyed by overgrazing by snow geese (Jefferies et al. 2006). Despite this recently emerged, abundant evidence for grazer control in salt marshes, all initial studies investigating marsh dieback in the United States did not employ cages to test for the relative impacts of grazers (Carlson et al. 2001; McKee et al. 2004; Ogbern and Alber

2006). Once consumer removal cages were deployed in New England (Holdredge and Bertness 2008) and in the southeastern United States (Silliman et al. 2005), however, it became quickly clear that these correlation-only approaches to examining causes of marsh dieback had succumbed to one of the oldest fallacies in science—mixing up cause and effect. Exclusion of snails in the Southeast and crabs in the Northeast showed that these invertebrate consumers were contributing in significant ways to marsh dieback (fig. 12.2) and that their impact in some cases expanded initial dieback areas by more than 200 percent. Extensive surveys across regional scales showed that crab and snail grazing occurred at almost all marsh dieback sites. Clearly, the relative importance of marsh grazers and top-down forces needs to be experimentally tested in all instances of putative, marsh dieback, and their potential impacts can no longer be ignored. The role of marsh consumers as primary contributing factors (along with physical stressors) to marsh dieback is now firmly established with experimental science. This science and how its has changed the old marsh paradigm must now be referenced to inform conservationists and managers in the design of experiments aimed at elucidating drivers of marsh dieback. In addition, studies must ask when and where top-down forces are important, and how top-down impacts vary in their relative importance across gradients in physical and bottom-up forces.

CLIMATE-INDUCED CHANGES

The majority of investigators studying the dieback phenomenon have linked large-scale climate change events with the present dieback phenomenon. Links between initiation of dieback and antecedent drought conditions exist for the Louisiana, Georgia, Virginia, Delaware, and New England dieback events (McKee et al. 2004; Silliman et al. 2005; Bason and Jacobs 2007; Ogburn and Alber 2006; Rosza 2006). Notable exceptions to this pattern

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are the Florida panhandle dieback (Carlson et al. 2001) and a dieback event in Wellfleet, Cape Cod, Massachusetts, in 2005–2006 (Smith 2006). The specific mechanisms that have been connected to antecedent drought conditions with the present dieback are drought-induced grazing, salinity stress, and oxidation and ultimate acidification of marsh soils (McKee et al. 2004; Silliman et al. 2005; Ogburn and Alber 2006; Rosza 2006; see also the discussion on postulated causes). Given the virtual inevitability of commonplace episodic climatic shifts in the upcoming decades (International Panel on Climate Change 2007), we are challenged with predicting possible responses to future climate scenarios and formulating a proper reaction. The impacts of small-scale climatic shifts on vegetation community structure in tidal marshes has been documented previously. Dunton et al. (2001) demonstrated significant shifts in coverage by various marsh perennials on an interannual time frame in response to local precipitation patterns in an arid marsh in Texas. In addition, local manifestations of sealevel rise operating over longer time frames are impacting tidal marshes (Hartig et al. 2002; Kennish 2001). Is the current dieback scenario indicative that long-term episodic events such as increased drought frequency and severity may impact the vegetation community across a wide variety of tidal systems? Changes in tidal marsh vegetation composition have been previously linked to episodic events operating over long time scales. Chambers et al. (2003) explained a trend between timing of Phragmites invasions of tidal marshes throughout North America and the metonic lunar tidal cycle. Significant expansion by Phragmites occurs at about twenty-year intervals and coincides with times of astronomically low tides operating on the metonic cycle (18.6-year cycle). The metonic cycle is controlled by variation in rotation of the nodal points of the lunar orbit and the Earth’s solar orbit and results in shifts in ocean tides as much as five to ten centimeters. Chambers et al. (2003) suggested that Phragmites invasion could be facilitated by 244

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episodic reductions in tidal excursions when these excursions result in low porewater sulfide, which has been shown to decrease vigor and expansion of Phragmites (Chambers et al. 2002). In this manner, episodic sea-level fluctuations open up windows of opportunity for large-scale ecological change that occur against a backdrop of increasing anthropogenic stressors. A similar shift of dominant vegetation may have occurred in European salt marshes in response to eutrophication (Rozema et al. 2000). If regular episodic drought becomes instrumental in triggering mortality events for S. alterniflora, a dominant foundation species, it may create the opportunity for significant plant community shifts throughout tidal marshes. These episodic events may be particularly important in the coastal zone and estuaries where climate shifts have multiple simultaneous results such as change in river discharge, change in tidal regime (sea-level rise), and chemical shifts associated with drying of saline soils (McKee et al. 2004). The acute nature of the recent dieback events would seem to indicate that coastal salt marshes are experiencing an unprecedented condition such as would be expected when local stressors operate synergistically with a large-scale climatic event.

SYNERGISM AND MULTIPLE STRESSORS One of the most significant challenges in ecosystem management and conservation biology is recognizing and dealing with multiple stressors in anthropogenically dominated systems (Breitburg et al. 1998; Folt et al. 1999; Groom et al. 2006). Inherent in this challenge is recognition of interaction among individual stressors (Breitburg et al. 1997) and the influence of antecedent conditions or history of exposure to either multiple or single stressors (Hughes and Connell 1999). One consistent conclusion emerging from studies of past or present marsh dieback is that no single causal factor is evident but that multiple factors are involved and they vary in their importance and interactions from site to site (Smith 1970;

Linthurst and Seneca 1980; Armstrong et al. 1996; Day et al. 2000; Silliman et al. 2005; Rosza 2006). Multiple stressors and interactive effects have been expressly noted for prior large-scale mortality episodes in ecological systems. A common theme among these large-scale events is the presence of a new level of threat emerging out of a history of long-term, multiple stressors. Among these is the long-term, widespread degradation of coral reef communities that are subject to overfishing, species introductions, tourism-related disturbances, rising sea surface temperature, pollution, storm impacts, and turbidity events (Pandolfi et al. 1997; Hughes and Connell 1999). The global decline of amphibians is described as a potentially complex interaction between relatively well-known historical stressors (exploitation, habitat change, species introductions) and less understood newer stressors (climate change, ultraviolet radiation, emerging pathogens) (Collins and Storfer 2003). The mortality of Long Island Sound lobster populations in 1999 has been attributed to the interaction of pathogen, sediment contaminants, and an episode of particularly low dissolved oxygen in the bottom waters coinciding with record-high water temperatures (Valente and Cuomo 2005). A large part of the complications associated with factor interaction in complex ecosystems is the variety of possible mechanisms and manifestations. Models describing the outcome of factor interaction can account for either additive or multiplicative effects (Folt et al. 1999). These combined effects can also act synergistically or antagonistically (Folt et al. 1999). For obvious logistical reasons and dominance of the bottom-up marsh paradigm, the effects of individual stressors are more commonly tested than a combination of impacts (Folt et al. 1999). This is true of study into sublethal stressors in S. alterniflora and other tidal marsh macrophytes, although important experimental work into factor interaction has been conducted (Silliman and Zieman 2001; Silliman et al. 2005). However, these studies commonly manipulate

two factors at once, salt or nitrogen and snail presence. The outcomes of these and more complex experiments will become increasingly important to elucidating interactive, causation chains if dieback events continue. One substantial question emerging from the study of interactive factors is how a normally resilient ecosystem such as coastal marshes will respond under continued exposure to an increasing magnitude of multiple, interacting factors. Breitburg et al. (1997) argue that when a dominant species is the target of stressors (as for S. alterniflora), the ecosystem response is similar whether multiple or single stressors are operating. For tidal marshes, edaphic stress is higher compared to nontidal systems, and the system normally experiences regular perturbation in the form of tidal exchange. The vegetation community of tidal marshes successfully accommodates these stressors when they fall within a normal range of variation (Odum et al. 1995). When multiple stressors are compounded or operate under a set of circumstances unique for that system, regardless of its resilience, the likelihood of a system-level response should increase. In this case, one could argue that multiple, sublethal stressors that are compounded increase the likelihood of a system-level response initiated through the dominant species. Furthermore, it would seem more likely that interactive factors would act synergistically to initiate a systemwide degradation. There is good reason to believe this what is happening with the current salt marsh dieback on eastern U.S. coastlines. The exact nature of this interactive effect, however, is impossible to ascertain without well-designed, targeted experimentation in the field before, during, and after dieback events. There is strong evidence that record droughts at the regional level act as a trigger factor with the current marsh dieback episodes (Stewart et al. 2001; McKee et al. 2004; Silliman et al. 2005; Rozsa 2006). The obvious lack of response to prior drought episodes may simply be because they were not severe enough in intensity and/or duration to trigger a dieback event. More likely,

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this dieback is symptomatic of multiplicative stressors working synergistically with drought conditions for the first time. In other words, background stressors (increasing snail densities, new fungal pathogens, increased eutrophication) may have been accumulating through time to a threshold where synergism now acts in a unique way. Unfortunately, this means there is also a strong likelihood that the nature of a dieback phenomenon, such as the one with S. alterniflora, could differ depending on the future trajectory of the stressors and their outcomes (Hughes and Connell 1999) and become more frequent in the near future. Ecological stress can start with locally induced disturbances that become amplified through a nonlinear ecosystem response (Jeffries 1997). Alternatively, localized disturbances can be manifested when they are exacerbated by a larger-scale event (e.g., drought conditions). If the present dieback events are linked by the occurrence of drought, then the exact interactive effect operating at the local level could be different from site to site. The upside in this case is the ability to accept competing ideas of the postulated causes at the local level and employ appropriately targeted management approaches. The downside is that this would indicate a larger problem of system response to climate-induced changes, and the system could change further or experience unpredictable responses if these large-scale forcing factors continue into the future.

CONCLUSION AND RECOMMENDATIONS Globally, ecosystems are displaying symptoms of cumulative stressors (Vitousek et al. 1997; Millennium Ecosystem Assessment 2005). Recent widespread coral bleaching events and a worldwide decline in abundance and diversity among amphibian populations are just two examples where multiple complex factors are manifesting themselves at unprecedented levels (Pandolfi et al. 2003; Collins and Storfer 2003). Tidal marshes have also historically experienced shifts in vegetation community 246

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structure resulting from oftentimes complex biotic and abiotic interactions, but these typically operate over long time scales or as relatively small-scale events within individual marshes (Redfield 1972; Warren and Niering 1980; Rozema et al. 2000; Donnelly and Bertness 2001). With the onset of the present dieback events, tidal marshes would appear to be experiencing a new level of perturbation symptomatic of increasing threat from climate change (Harley et al. 2006) and human population pressure (Kennish 2002). Synergism among the postulated causes is a much more likely explanation for the recent dieback phenomenon than the operation of any single stressor, with the simple additive combination of stressors having been already demonstrated in the field (i.e., grazers, acidity, and salt stress). The causes are most likely resulting from a unique set of conditions set into motion by episodes of climatic stress (regional drought). Based on the evidence to date, edaphic stressors (acidity, metal toxicity, elevated salinity) and biotic stressors (runaway grazing, grazer facilitated diseases, trophic cascades, pathogens) are each likely to play a role in the dieback events. These factors seem to be tolerated independently by the normally resilient plant species in these tidal marshes, but in combination the stressors represent a unique threat. In addition, the exact combination of factors may vary by site depending on the local biological and meteorological setting (intertidal herbivore and predator population levels, propensity for hypersaline conditions, etc.). Synergism among stressors was suspected early in marsh dieback episodes (Smith 1970; Ostendorp 1989) and is clearly suspect in the current dieback (McKee et al. 2004; Silliman et al. 2005). Even though our understanding of factor interaction has improved we are still challenged about how it operates or how best to diagnose it (Folt et al. 1999; Breitburg et al. 1998; Collins and Storfer 2003). A full understanding of the dieback events will necessitate a resolution of multiple operating hypotheses. In

particular, it will be important to understand how background levels of historic stressors are operating with the added impacts of climate change and possibly novel biotic interactions (i.e., anthropogenic exploitation of important predatory species and resultant trophic cascades, sensu Collins and Storfer 2003). The present marsh dieback is impacting a normally resilient, robust collection of species (but mostly S. alterniflora). S. alterniflora is considered a foundation species in tidal marshes by virtue of its dominance, high level of production, and its role as habitat (Mitsch and Gosselink 2000). This one species impacts the ecology of virtually all tidal marshes in the U.S. Gulf and Atlantic coasts at a fundamental level (Mendelssohn and Morris 2000; Mitsch and Gosselink 2000). If the dieback events continue and are exacerbated by future episodic events of climate change, there is significant potential for a major change in the coastal ecology of these regions, as we could lose the vital ecosystem services marsh plants generate (e.g., fisheries, wave protection, pollution filters). Future research on the recent dieback events should be directed toward a few key questions. What are the exact interactions resulting in the dieback events? Are the causes and the system response similar among various locations experiencing the dieback? Have anthropogenic stressors operating over a long time scale combined with the more immediate causal factors into a new scenario where dieback will become a common manifestation? What will be the nature of responses to future episodic perturbations (regional climate events, top-down ecosystem-level impacts)? In addition, management of tidal marshes will need to accommodate synergistic stressors and identify controls/abatements for episodic trigger events such as climate change and runaway biological forces (grazing). We recommend that, where appropriate, managed hydrologic inflows and outflows are designed to respond to water excess or deficiency within a short time frame. Furthermore, wildlife conservation planning should be reassessed to

account for necessary biological controls (i.e., predation on potentially prolific herbivore species). Replanting of dieback areas should consider genetic resistance and variation and positive interactions (Halpern et al. 2007) to promote sustainable plant populations and regrowth. On one hand, the relatively quick recolonization in several of the dieback areas is indicative that the resiliency of the coastal salt marsh system has, for the time being, persevered. On the other hand, the acuity, magnitude, and the ubiquitous nature of this recent dieback by a normally dominant and resistant plant species has few analogs. The sudden dieback phenomenon should be a warning sign that even the most resilient systems may be prone to impact when stressors operate at a complex, interactive level against a history of antecedent anthropogenic impacts. The current marsh dieback and similar examples from other coastal marine systems (coral reefs, mangroves, seagrass beds) underscore the need to understand the ecology of multiple, interactive stressors not only at the local level but overlapping with larger changes associated with resource exploitation (reduction in toplevel predators) and global climate change (droughts). Management practices must also adapt to these conditions if they are to effectively deploy ecologically relevant remediation, restoration, and conservation plans to address the salt marsh dieback. Acknowledgments. The authors appreciate the helpful conversations with L. Blum, B. Christian, W. Elmer, and D. Burdick. We thank two anonymous reviewers for their comments. A. Howard of the Delaware Department of Natural Resources and Environmental Control kindly provided useful information on the Delaware vegetation dieback. D. Yozzo and R. Chambers provided comments on an earlier version of the chapter. We acknowledge funding from Georgia Sea Grant, David H. Smith Conservation Research Fellowship, The Nature Conservancy, the National Science Foundation,

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the Mellon Foundation, the National Atmospheric and Oceanic Association National Estuarine Research Reserve System, the Georgia Coastal Ecosystems-LTER, the University of Georgia Marine Institute, and the Environmental Protection Agency. This work could not have been completed without the help of Tracy Buck, Sarah Lee, and Jane Garbisch.

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13

Patterns of Salt Marsh Loss within Coastal Regions of North America presettlement to present Keryn Bromberg Gedan and Brian R. Silliman Habitat degradation and loss is a severe threat to nearshore environments. Coral reefs, seagrass beds, mangroves, and salt marshes have all been negatively affected by human actions, and ecosystem services provided by these habitats have been diminished. Here, we review the extent of tidal marsh loss in the coastal regions of North America over a centennial to bicentennial time scale. Aerial photos and satellite imagery are often used to set baselines of habitat extent, but we find pre–satellite era historical losses to be significant, with the United States’ Pacific Coast having the most dramatic reduction of 93 percent in coastal marsh area. Although there is serious and rapid marsh dieback occurring across the eastern United States due to drought, physical stress, and overgrazing, these areas were not included in our loss statistics due to their seemingly rapid recovery. The greatest losses historically were due to direct habitat conversion to agricultural or urban lands. However, in the last few decades, the primary anthropogenic impacts have been indirect, due to hydrologic alteration in the Gulf Coast and runaway herbivory in the Hudson Bay region. Over the last two centuries, coastal marsh losses in North America have been permanent, despite many restoration and wetland creation efforts. Due to the absence of recovery once these ecosystems are converted to other land uses, we recommend more stringent protection of remaining coastal marsh habitat and proactive protection of marsh habitat in areas of continued development. More difficult to address are anthropogenic factors causing indirect salt marsh loss, as these factors are difficult to pinpoint mechanistically and often cause considerable negative impacts before they are identified. In the face of sea-level rise, continued hardening of the North American shoreline by human development, drought stress, and overgrazing will be some of the most glaring future long-term threats to the persistence of extensive salt marsh ecosystems in North America.

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Habitat degradation and loss in coastal interface systems is cause for serious alarm due to the associated loss of valuable ecosystem services including critical nursery, fishery, and waterfowl habitat; shoreline buffering from storm erosion; nutrient recycling and absorption; and carbon dioxide uptake (Lotze et al. 2006). To understand how critical coastal habitats have been compromised and to establish key benchmarks for conservation and restoration goals, we will need to know the extent and causes of coastal habitat loss. Shifting or resetting historical baselines of ecosystem health and/or extent in other habitats has revealed dramatic changes in the size and structure of fish, coral reef, and benthic invertebrate communities (see Jackson et al. 2001 for examples). Loss of key coastal interface ecosystems critical to the success of nearby human populations has been severe, with coral reefs experiencing the highest degree of degradation. Pandolfi and colleagues (2003) estimate that there is essentially no remaining pristine reef habitat, and Gardner and colleagues (2003) found over 80 percent loss of coral cover in Caribbean reefs. Similarly, loss of oyster reefs and seagrass beds has been extreme, with oyster reefs having been commercially depleted well over a century ago in most locales (Romero 2003; Kirby 2004) and more than ninety thousand hectares of seagrass beds destroyed by dredging, eutrophication, and disease in the mid– to late twentieth century (Short and Wyllie-Echeverria 1996). The status of kelp habitats has varied wildly over the last several decades, in part due to kelp forests’ sensitivity to climate fluctuations, making it difficult to draw conclusions about the level of anthropogenic impact (Dayton et al. 1998). The largest scientific effort is presently being directed toward understanding coral reef habitat status. Although these habitats are important and at risk, reef degradation has been studied in finer and finer detail, while baselines of habitat loss for salt marsh and mangrove environments are, in many places, rare to nonexistent.

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Dahl (1990) estimates that 50 percent of all wetlands (both freshwater and marine) in the coterminous United States were lost between 1780 and 1980. One global approximation of coastal marsh loss, based on data from twelve large estuaries, estimated loss of a similar magnitude, 65 percent (Lotze et al. 2006), which would leave very little remaining habitat to perform vital services, such as storm-surge buffering and fishery provisioning, to evergrowing coastal populations. Studies of estuarine habitat degradation, such as the one done by Lotze and colleagues (2006), have been restricted largely to the United States and Europe, where detailed data about wetlands are widely available. In North America, Canada is in the process of developing a national wetland inventory (Milton and Hélie 2003). Mexico has no wetland inventory, and little data exist on Mexican wetland status except for the sixty-five sites included in the Ramsar List of Wetlands of International Importance (Ramsar Convention on Wetlands 2007). In the continental United States, there is currently an estimated 1.6 million hectares of vegetated coastal marsh (this term is synonymous with estuarine emergent wetlands and includes tidally influenced brackish and salt marshes) (Dahl 2006). In recent decades, the loss rate of coastal marsh has dropped to less than 1 percent per decade in the United States, due to national wetland legislation and a great deal of public attention (Dahl 2000, 2006). Currently, the largest losses of coastal marsh are in Louisiana, where a combination of natural and anthropogenic effects has caused sustained coastal land losses (Dahl 2006). Although contemporary rates of marsh loss are low, today’s coastal wetlands represent only a small fraction of their former size, and conversion has been largely permanent (figs. 13.1 and 13.2). In the early settlement of North America, salt marsh lands were highly valued as natural pasture. Salt marsh grasses were used for a variety of commercial products, and some marshes

FIGURE 13.1 What remains: A tidal marsh in Quincy, Massachusetts, bisected by a highway interchange and surrounded by development. A large portion of the marsh has been converted to industrial land. From Office of Geographic and Environmental Information (MassGIS).

FIGURE 13.2 Coastal land use pressures have surrounded a small marsh in Somerset, Massachusetts (circled in black). To the left of the marsh is Brayton Point Power Plant, a major source of air pollution in New England. Although the marsh is severely degraded, it has been left intact and is not accounted for in marsh loss statistics. From Rhode Island Geographic Information System (RIGIS).

were reclaimed for agricultural fields in the European tradition (Sebold 1998; Hatvany 2003). But with the westward expansion, the agricultural value of salt marshes declined, and large areas of coastal marsh were converted to urban lands. In the United States and Canada, losses due to direct conversion have slowed, whereas in many developing nations around the world, reclamation for agricultural and urban land continues to pose a major threat to coastal marsh areas.

In this chapter, we estimate historical losses of coastal marsh by region, dividing U.S. coasts into four major regions (fig. 13.3): North Atlantic, South Atlantic, Gulf Coast, and Pacific Coast. Two coastal regions in Canada will be discussed: the Canadian Maritimes and Hudson Bay (fig. 13.3). No information could be found on salt marsh loss from the Canadian Pacific region or from the Mexican coasts, and we were unable to make a quantitative study of these regions.

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Hudson Bay

Canadian North Atlantic

U.S. Pacific Coast U.S. North Atlantic

South Atlantic Gulf Coast

FIGURE 13.3 North American coastal regions discussed in this chapter.

MEASURING COASTAL WETLAND LOSS Although disregarded in many studies of wetland loss trends, historical patterns of wetland exploitation and loss have been explored in a number of published studies. Losses are measured by comparing one or multiple historical record(s) of estuarine marsh area (e.g., predevelopment hydric soil surveys, early navigational charts) with more modern data (e.g., GIS land use layers, National Wetlands Inventory data). These analyses measure total conversion; heavily impacted or degraded marsh areas are not taken into account. Analyses vary greatly in temporal and spatial scale, each with biases. Local scales often focus on a single harbor and likely overestimate regional losses. In addition, regional analyses tend to be less accurate, because they draw from multiple sources and use estimation techniques. To compare between time periods and different spatial scales, percentages of areal loss are often used. We will supplement percentage data with estimates of areal loss to facilitate comparisons among regions. In this chapter, we have summarized published long-term estimates of marsh loss and compiled them into regional loss estimates (fig. 13.4). Here, we include any studies with 256

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data from before 1900. Although our data would ideally come from much earlier, high-quality map data are rare prior to the founding of the U.S. Coast Survey in 1834 and the U.S. Geological Survey in 1879. Using estimates covering different time periods may seem to give an inaccurate estimate, but we do not see this as an intractable issue because data were available earlier in regions with greater human development early on, and marsh loss rates were coupled to development patterns. As we compare regional loss rates, this pattern will become clear.

U.S. NORTH ATLANTIC Along the North Atlantic coast of the United States, human use of salt marshes generally follows a pattern of drainage for agricultural and pastoral exploitation soon after European settlement, followed by spoil deposition for use as a landfill or for the direct conversion to urban land uses. Here, we discuss a few examples. A weighted average of these studies suggests a 38 percent loss of coastal marsh in the North Atlantic region since European settlement (baseline dates ranged from 1609 to 1893; fig. 13.4). Coastal development has been heavy along the Washington–Boston coastal corridor and

100% *high uncertainty

N=5

Marsh area lost

80%

N=2

N = 1*

60%

40%

N=6

20%

N=4 N = 1*

0%

93%

18%

12%

38%

64%

63%

Pacific Coast

Gulf of Mexico

US South Atlantic

US North Atlantic

Canadian Maritimes

Hudson Bay

FIGURE 13.4 Coastal marsh loss in United States regions, as quantified by a weighted average of published estimates of subregional areas. Number of subregional areas varied by region: Pacific Coast N ⫽ 5, Gulf Coast N ⫽ 4, U.S. South Atlantic N ⫽ 1, U.S. North Atlantic N ⫽ 6, Canadian Maritimes N ⫽ 2; and Hudson Bay N ⫽ 1.

has adversely affected salt marshes (fig. 13.5). Bromberg and Bertness (2005) estimated a 37 percent, or 16,253 hectare, loss of salt marsh across the New England region over the last two centuries. Losses were correlated with the extent of urbanization of coastal watersheds. Similarly, a study by Carlisle and colleagues (2005) found a rate of 30 percent loss in Massachusetts Bay, Cape Cod, and the islands’ salt marshes from 1893 to 1995, with the great majority of losses taking place before 1970, when most of the urban expansion and coastal development occurred in New England. The salt marshes of New York and New Jersey were used as sewage and trash dumps during the late 1800s and early 1900s before

finally being filled in to create more residential and industrial space. A scant 3 of the 149 hectares of presettlement salt marsh remain today in Manhattan, a 98 percent loss (Sanderson and Brown 2007). An estimated 73 percent of the Hackensack Meadowlands (Tiner, Swords, and McClain 2002) and 68 percent of the Staten Island coastal marshes (Tiner 2000; fig. 13.6) have been filled since the late 1800s. These losses in highly urbanized areas are virtually irrevocable. In the southern North Atlantic states of Delaware and Maryland, many large marsh complexes have been spared from development pressure, but they have still experienced rather large loss rates. Tiner (2005) estimates that

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FIGURE 13.5 Urban expansion in Boston between 1852 (top) and 1903 (bottom) resulted in a dramatic loss of tidal marsh (indicated in both maps with a plant symbol over light shading). Graphic reprinted with permission from Carlisle et al. 2005.

28 percent, or 2,720 hectares, of tidal marsh has been lost since presettlement days from the Nanticoke River watershed, a subbasin of Chesapeake Bay. Contrary to the northern examples, much of the salt marsh loss in the Nanticoke watershed has occurred in the mid–twentieth century and has been attributed to shoreline erosion, creek expansion, and interior ponding caused by sediment deficits and sea-level rise (Kearney, Grace, and Stevenson 1988). Within the last two years, small vegetation dieback areas have been noticed in the Spartina alterniflora zone of some New England salt marshes (Salzman 2006). The cause of vegetation death is consumption by herbivorous crabs, Sesarma reticulatum (Holdredge, Bertness, and Altieri in press). These areas may recover as 258

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dieback areas in southern U.S. marshes have. Therefore, these areas were not included here as lost habitat.

SOUTH ATLANTIC The loss of coastal marsh in the South Atlantic region of the United States is difficult to assess. One study of the North Carolina coastal plain estimated 12 percent of tidal marshes had been significantly altered since presettlement times (Cashin, Dorney, and Richardson 1992). Clearly, several coastal cities may have adversely affected large areas of salt marsh during their development. Savannah, Georgia, and Jacksonville, Florida, abut marshlands; and Fort Lauderdale and Miami, Florida, vicinities probably once

Remaining Coastal Wetlands in 1994/5

Coastal Wetlands in 1870

Remnants Now Freshwater Wetlands in 1994/5

FIGURE 13.6 Wetland loss in Staten Island, New York, between 1870 and 1994 has caused severe fragmentation of remaining salt marshes. Graphic reprinted with permission from Tiner 2000.

contained large areas of marsh and/or mangrove. The South Atlantic region still contains vast areas of salt marsh, approximately 362,000 hectares, or 21 percent of the U.S. total (Field et al. 1991). Although we list 12 percent as the average coastal marsh loss experienced by the South Atlantic region (fig. 13.4), this number should be treated with high uncertainty because of the lack of published data from the area. Within the past decade, considerable areas of S. alterniflora in South Atlantic and Gulf Coast salt marshes have experienced rapid browning and death. The definitive cause of the dieback phenomenon is unknown. This disturbance has affected upward of one hundred thousand hectares across Louisiana, Georgia, Texas, Florida, South Carolina, and North Carolina, but bare areas seem to recover over several growing seasons (McKee, Mendelssohn, and Materne 2004). An interaction between drought-induced physical stress and grazing by the snail Littoraria irrorata can create die-off areas, and fronts of grazing snails routinely advance dieback borders into otherwise healthy marsh habitat (Silliman et al. 2005), but explicit identification of the exact physical mechanism initiating sudden marsh dieback is lacking. Due

to their rapid recovery and cyclical nature, dieback areas were not included in our loss estimates. If, however, the principal cause of the dieback became severe enough to inhibit vegetation recovery, dieback would become a major source of coastal marsh habitat loss in both the South Atlantic and Gulf Coast regions.

GULF COAST Little information exists on marsh reclamation in the coastal Gulf of Mexico prior to the 1930s. As in the South Atlantic region, we can presume that the expansion of New Orleans, Corpus Christi, and possibly Houston areas reclaimed some marshland. Harrison and Kollmorgen (1947) report an astonishing 55,632 hectares of coastal marsh reclamation projects completed between 1907 and 1916 in the Mississippi Delta. Reclamation projects in Louisiana were frequently problematic, and many of these reported projects likely reflooded and were abandoned (Harrison and Kollmorgen 1947). Still, it is likely historical losses in the Gulf Coast region have been underestimated. Additionally, considerable spatial losses may have gone unnoticed due to the ample amount

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of naturally occurring coastal marsh habitat— over eight hundred thousand hectares, or 58 percent of the U.S. total coastal marsh area (Field et al. 1991). In a weighted average of six sites, we estimate that the Gulf Coast has lost 18 percent of its coastal wetlands (fig. 13.4). Existing data from the 1930s until today suggest high modern rates of marsh loss. Whereas coastal marsh loss rates have bottomed out at nearly zero for the last decade in many regions, loss rates in the Gulf Coast remain extremely high, at between two thousand and ten thousand hectares ha per year (Dahl 2006). The overwhelming majority of these losses are occurring in coastal Louisiana, where over 395,000 hectares of marsh have been lost between 1932 and 1990 (Britsch and Dunbar 1993). Twelve percent of these losses are due to direct conversion to navigation, trapping, and petroleum canals (Davis 1973; Britsch and Dunbar 1993). The remainder of losses is due to indirect effects of hydrologic alterations, but the relative contributions of sediment starvation, belowground extraction, erosion, and natural subsidence are debatable. These high rates of coastal land loss in the Mississippi Delta area, exacerbated in recent years by severe hurricane strikes, are currently a subject of great concern.

PACIFIC COAST Due to geomorphological differences, the Pacific Coast of the United States has never supported development of large areas of coastal marsh as seen on the Atlantic and Gulf coasts. The Pacific Coast has a steeper, shorter continental shelf that abuts a subduction zone and fewer large estuaries. With a smaller area of marsh to begin with, even minor losses can seem proportionally large. In part for this reason and in part due to heavy coastal development, the historical loss statistics are dramatic: 57 percent, or 400 hectares, of Elkhorn Slough (Van Dyke and Wasson 2005); 87 percent, or 1,671 hectares, of San Diego’s coastal marsh

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(Zedler 1996); and 94 percent, or 207,500 hectares, of San Francisco Bay’s wetlands (Atwater et al. 1979) have disappeared. The Pacific Northwest has fared slightly better. The small estuary of Coos Bay, Oregon, has lost 28 percent, or 134 hectares, of marsh; and Willapa Bay, Washington, has lost a mere 3 percent, or 71 hectares, since 1855 (Borde et al. 2003). Before being converted to urban land uses, many of the Pacific Coast marshes were first filled for railroad passages or converted to evaporative ponds for salt harvesting, as opposed to agriculture as in the North Atlantic. Here, we combined five subregional estimates for a loss statistic of 93 percent loss (fig. 13.4). The formerly vast marshes of San Francisco Bay Estuary dominate these statistics. These net losses are even more extreme considering that marsh formation was accelerated in the mid–nineteenth century by an increased sediment flux from hydraulic gold mining techniques (Nichols et al. 1986).

CANADIAN MARITIMES The Canadian Maritimes region, with large macrotidal areas and the sizeable estuaries of Bay of Fundy and the St. Lawrence River, has a significant amount of coastal marsh. Similar to the U.S. North Atlantic region, the salt marshes of the Canadian Maritimes were diked and farmed by early colonists (fig. 13.7). Eastern Canadian colonists were French Acadians with a long and proud tradition of agricultural reclamation (Hatvany 2003). Within the Côte-de-Sud region of the St. Lawrence Estuary, these reclamation activities resulted in a 32 percent, or 1,009-hectare, conversion of tidal marsh prior to 1971 (Reed and Moisan 1971). In a review of land use change in Canadian wetlands, Lynch-Stewart (1983) cites even more drastic agricultural marsh conversions—a loss of 66 percent, or 23,235 hectares, of Nova Scotian salt marsh since presettlement times. In a weighted average of these two cases, we found a 64 percent loss overall in the

FIGURE 13.7 An agricultural dike in New Brunswick, Canada, separates a coastal fringe of salt marsh (left of dike) from reclaimed farmland (right of dike). Photo courtesy of K. Bromberg Gedan.

Canadian Maritimes (fig. 13.4). In the twentieth century, many of these dikes were breached, and reclamation efforts subsided due to rising dike maintenance costs and interest in bird conservation and ecotourism in the Maritimes region (Summerby-Murray 1999).

HUDSON BAY The Hudson Bay region salt marshes have experienced a high degree of loss over the last thirty years due indirectly to human actions. Agricultural subsidy in the U.S. Midwest has caused a massive increase in the lesser snow and Ross’s geese populations, which have subsequently grubbed away the vegetation of entire marshes in their winter feeding ground (Abraham et al. 2005). In their denuded state, marsh soils become hypersaline, and plant recolonization is inhibited, leading to an alternate stable state of bare peat. From 1976 to 2000, over thirty thousand hectares of marsh were destroyed by geese along just a portion of the Hudson Bay coast, a decline of 63 percent at one site (Abraham et al. 2005; Jefferies, Jano, and Abraham 2006) (fig. 13.4). This represents a novel case in that the region has experienced severe habitat loss in a high-latitude, sparsely populated area, generally considered to be outside the sphere of human influence.

CONCLUSION Establishing historical baselines for comparison to current habitat status is critical to understanding the degree of change and thereby setting appropriate conservation goals. Here, we have summarized baselines for the coastal United States and portions of coastal Canada. In these two countries, nationally sponsored wetland inventories are quantifying wetland area status and trends, but in both cases baselines are being set with late twentieth-century data. Establishing appropriate historical baselines should be given equal footing to developing modern inventories. Without correct baseline data, the extent of habitat loss and degradation will be greatly underestimated (fig. 13.8). Other nations, including Mexico, may have fewer resources to put into quantifying habitat health and establishing environmental baselines. However, the process need not be all-consuming. The Ramsar Convention has chosen internationally important wetland sites in many countries and has collected extensive data on the current status of and anthropogenic threats to each site to aid in management efforts (Ramsar Convention on Wetlands 2007). For resource-poor countries, establishing baselines for these sites where data are already available is a good starting point.

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12,000 Boston Hackensack

Salt marsh remaining (ha)

10,000

San Diego

8000

6000

4000

2000

0 1850

1870

1890

1910

1930

1950

1970

1990

Year FIGURE 13.8 Reconstruction of salt marsh area over time for three areas: Massachusetts Bay, Hackensack, and San Diego (see source citations in the text). Use of the 1950 salt marsh area as a baseline (dotted lines) severely distorts the degree of habitat loss.

The data summarized in this chapter show that North American coastal marshes have experienced a high degree of loss, particularly in highly populated coastal areas. If we apply these results more generally to a model of the way coastal habitats are utilized over cultural periods of development (e.g., Frontier, Agricultural, Industrial, Postindustrial; see Mustard et al. 2004 for greater detail), a pattern of continual habitat loss without significant recovery develops over a cultural time scale (fig. 13.9). This pattern stands in stark contrast to a model of habitat exploitation developed for terrestrial landscapes that predicts a concentration of resource use and habitat recovery in the industrial and postindustrial periods (Mustard et al. 2004). The pattern of forestation in the eastern United States, for example, fits this model rather closely (Mustard et al. 2004). Due to this pattern of ongoing and irreversible loss, we recommend continued stringent protection of remaining coastal marsh. In countries currently in a process of cultural development and indus262

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trialization, there is a need for proactive prevention of habitat loss in coastal areas, before permanent habitat conversion. Differences in historical patterns of land use over the past three centuries and variation in the natural distribution of coastal marsh have resulted in very different patterns of coastal marsh loss across North America. The timing of wetland loss differs among regions (fig. 13.10), as does the cause. Direct human impacts of urban expansion and land filling are responsible for the majority of marsh loss in the North Atlantic and Pacific coasts, with the heaviest losses occurring there in the nineteenth and early twentieth centuries. In contrast, the coastal marsh losses in the Gulf Coast and the Canadian Hudson Bay region have been much more recent, at their peak in the late twentieth century and still occurring today. In both regions, large areas of marsh are being lost over a relatively short time period. Moreover, recent losses have been caused indirectly by human action, by hydrologic alteration in the Gulf Coast,

100%

Habitat area remaining

80%

60%

40%

20%

0%

Frontier

Agricultural

Early industrial

Late industrial

Post industrial

FIGURE 13.9 Coastal marsh loss over cultural development periods (solid line) shows a declining pattern without recovery in remaining marsh habitat. The land use model developed for terrestrial habitats (Mustard et al. 2004) (dashed line) predicts major habitat loss during the agricultural period and resource recovery in the industrial and postindustrial periods. Circles are unweighted means ⫾2 SE for twelve sites; Gulf Coast sites except Louisiana, were excluded (Britsch and Dunbar 1993) due to inadequate historical coverage.

FIGURE 13.10 A generalized model of coastal marsh loss over time by region.

and by crop-fed goose populations in the Hudson Bay area. These indirect effects are much more difficult to predict and, therefore, harder to prevent than direct habitat conversion. Emerging evidence of marsh dieback in the U.S. Northeast (Holdredge, Bertness, and Altieri in press; Salzman 2006) and salt marsh erosion due to decreased ice cover and rising sea levels in eastern Canada (Daigle 2006) suggests that indirect

anthropogenic effects will continue to impact coastal marshes in the near future. It is clear that this topic needs further exploration by marsh ecologists and conservation biologists. That a large percentage of coastal marsh has been lost on nearly all coasts of North America is troubling. The U.S. South Atlantic is the only region that may have maintained much of its natural coastal wetland area since European

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settlement. The potential services offered by coastal wetlands in maintaining healthy waterfowl and fish populations, shoreline buffering, nitrogen removal, and carbon sequestration have been compromised by these losses. In many places, fishery declines, artificially hardened shorelines, and eutrophied coastal waters are the consequence. We must make establishing historical baselines and exploring management of indirect anthropogenic impacts in coastal wetland habitats urgent conservation priorities.

Acknowledgments. We thank Mark Bertness for helpful comments on this chapter. This chapter was written with funding from an Environmental Protection Agency (EPA) STAR graduate fellowship to K. Bromberg Gedan. The EPA has not officially endorsed this publication, and the views expressed herein may not reflect the views of the EPA. We acknowledge funding to B. R. Silliman from Georgia Sea Grant, the National Science Foundation, The Nature Conservancy, the National Atmospheric and Oceanic Association National Estuarine Research Reserve System, the Georgia Coastal Ecosystems-LTER, the University of Georgia Marine Institute, a David H. Smith Conservation Research Fellowship, and the EPA.

REFERENCES Abraham, K. F., R, L. Jefferies, and R. T. Alisauskas. 2005. The dynamics of landscape change and snow geese in mid-continent North America. Global Change Biology 11: 841–855. Atwater, B. F., S. G. Conard, J. N. Dowden, C. W. Hedel, R. L. MacDonald, and W. Savage. 1979. History, landforms, and vegetation of the estuary’s tidal marshes. Pages 247–286 in T. J. Conomos, A. E. Leviton, and M. Berson (eds.), San Francisco Bay: The Urbanized Estuary, Investigations into the Natural History of San Francisco Bay and Delta with Reference to the Influence of Man. San Francisco: Pacific Division of the American Association for the Advancement of Science. Borde, A. B., R. M. Thom, S. Rumrill, and L. M. Miller. 2003. Geospatial habitat change analysis in Pacific Northwest coastal estuaries. Estuaries 26: 1104–1116. 264

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Britsch, L. D., and J. B. Dunbar. 1993. Land loss rates: Louisiana coastal plain. Journal of Coastal Research 9: 324–338. Bromberg, K., and M. D. Bertness. 2005. Reconstructing New England salt marsh losses using historical maps. Estuaries 28: 823–832. Carlisle, B. K., R. W. Tiner, M. Carullo, I. K. Huber, T. Nuerminger, C. Polzen, and M. Shaffer. 2005. 100 Years of Estuarine Marsh Trends in Massachusetts (1893 to 1995): Boston Harbor, Cape Cod, Nantucket, Martha’s Vineyard, and the Elizabeth Islands. Massachusetts Office of Coastal Zone Management, Boston; U.S. Fish and Wildlife Service, Hadley, MA; and University of Massachusetts, Amherst, MA, Cooperative Report. Hadley, MA: U.S. Fish and Wildlife Service. Cashin, G. E., J. R. Dorney, and C. J. Richardson. 1992. Wetland alteration trends on the North Carolina coastal plain. Wetlands 12: 63–71. Dahl, T. E. 1990. Wetlands Losses in the United States 1780’s to 1980’s. Washington, DC: Department of the Interior, Fish and Wildlife Service. ———. 2000. Status and Trends of Wetlands in the Conterminous United States 1986 to 1997. Washington, DC: Department of the Interior, Fish and Wildlife Service. ———. 2006. Status and Trends of Wetlands in the Conterminous United States 1998 to 2004. Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. Daigle, R. 2006. Impacts of Sea-Level Rise and Climate Change on the Coastal Zone of Southeastern New Brunswick. Fredericton: New Brunswick: Environment Canada. Davis, D. W. 1973. Louisiana canals and their influence on wetland development. Unpublished PhD diss., Louisiana State University, Baton Rouge. Dayton, P. K., M. J. Tegner, P. B. Edwards, and K. L. Riser. 1998. Sliding baselines, ghosts, and reduced expectations in kelp forest communities. Ecological Applications 8: 309–322. Field, D. W., A. J. Reyer, P. V. Genovese, and B. D. Shearer. 1991. Coastal Wetlands of the United States. Washington, DC: National Oceanic and Atmospheric Administration. Gardner, T. A., I. M. Cote, J. A. Gill, A. Grant, and A. R. Watkinson. 2003. Long-term region-wide declines in Caribbean corals. Science 301: 958–960. Harrison, R. W., and W. M. Kollmorgen. 1947. Drainage reclamation in the coastal marshlands of the Mississippi River Delta. Louisiana Historical Quarterly 30: 654–709. Hatvany, M. G. 2003. Marshlands: Four Centuries of Environmental Change on the Shores of the St. Lawrence. Laval, Quebec: Les Presses de l’Université Laval.

Holdredge, C., Bertness, M. D., and A. H. Altieri, in press. Role of crab herbivory in die-off of New England salt marshes. Conservation Biology. Jackson, J. B. C., M. X. Kirby, W. H. Berger, K. A. Bjorndal, L. W. Botsford, B. J. Bourque, R. H. Bradbury, R. Cooke, J. Erlandson, J. A. Estes, T. P. Hughes, S. Kidwell, C. B. Lange, H. S. Lenihan, J. M. Pandolfi, C. H. Peterson, R. S. Steneck, M. J. Tegner, and R. R. Warner. 2001. Historical overfishing and the recent collapse of coastal ecosystems. Science 293: 629–638. Jefferies, R. L., A. P. Jano, and K. F. Abraham. 2006. A biotic agent promotes large-scale catastrophic change in the coastal marshes of Hudson Bay. Journal of Ecology 94: 234–242. Kearney, M. S., R. E. Grace, and J. C. Stevenson. 1988. Marsh loss in Nanticoke Estuary, Chesapeake Bay. Geographical Review 78: 205–220. Kirby, M. X. 2004. Fishing down the coast: Historical expansion and collapse of oyster fisheries along continental margins. Proceedings of the National Academy of Sciences of the USA 101: 13096–13099. Lotze, H. K., H. S. Lenihan, B. J. Bourque, R. H. Bradbury, R. G. Cooke, M. C. Kay, S. M. Kidwell, M. X. Kirby, C. H. Peterson, and J. B. C. Jackson. 2006. Depletion, degradation, and recovery potential of estuaries and coastal seas. Science 312: 1806–1809. McKee, K. L., I. A. Mendelssohn, and M. D. Materne, 2004. Acute salt marsh dieback in the Mississippi River Deltaic Plain: A drought-induced phenomenon? Global Ecology and Biogeography 13: 65–73. Milton, G. R., and R. Hélie. 2003. Wetlands Inventory and Monitoring: Partnering to Provide A National Coverage, Conference Proceedings, Conference on Canadian Wetlands Stewardship. North American Wetlands. Mustard, J. F., R. S. Defries, T. Fisher, and E. Moran. 2004. Land-use and land-cover change pathways and impacts. Pages 411–430 in G. Gutman, A. C. Janetos, C. O. Justice, E. F. Moran, J. F. Mustard, R. R. Rindfuss, D. Skole, B. L. I. Turner, and M. A. Cochrane (eds.), Land Change Science: Observing, Monitoring and Understanding Trajectories of Change on the Earth’s Surface. New York: Springer. Nichols, F. H., J. E. Cloern, S. N. Luoma, and D. H. Peterson. 1986. The modification of an estuary. Science 231: 567–573. Pandolfi, J. M., R. H. Bradbury, E. Sala, T. P. Hughes, K. A. Bjorndal, R. G. Cooke, D. McArdle, L. McClenachan, M. J. H. Newman, G. Paredes, R. R. Warner, and J. B. C. Jackson. 2003. Global trajectories of the long-term decline of coral reef ecosystems. Science 301: 955–958.

Ramsar Convention on Wetlands. 2007. Gland, Switzerland: Ramsar Convention Secretariat. www.ramsar.org. Reed, A., and G. Moisan. 1971. The Spartina tidal marshes of the St. Lawrence Estuary and their importance to aquatic birds. Le Naturaliste Canadien 98: 905–921. Romero, A. 2003. Death and taxes: The case of the depletion of pearl oyster beds in sixteenth-century Venezuela. Conservation Biology 17: 1013–1023. Salzman, A. 2006. Something is killing the marsh grass, but no one is sure what it is. New York Times, July 2. Sanderson, E. W., and M. Brown. 2007. Mannahatta: An ecological first look at the Manhattan landscape prior to Henry Hudson. Northeastern Naturalist 14, no. 4: 545–570. Sebold, K. R. 1998. The low green prairies of the sea: Economic usage and cultural construction of the Gulf of Maine salt marshes. Unpublished PhD diss., University of Maine, Orono. Short, F. T., and S. Wyllie-Echeverria. 1996. A review of natural and human-induced disturbance of seagrasses. Environmental Conservation 23: 17–27. Silliman, B. R., J. van de Koppel, M. D. Bertness, L. E. Stanton, and I. A. Mendelssohn. 2005. Drought, snails, and large-scale die-off of southern U.S. salt marshes. Science 310: 1803–1806. Summerby-Murray, R. 1999. Interpreting cultural landscapes: A historical geography of human settlement on the Tantramar marshes, New Brunswick. Salzburger Geographische Arbeiten 34: 157–174. Tiner, R. W. 2000. Wetlands of Staten Island, New York: Valuable Vanishing Urban Wildlands. Prepared for U.S. Environmental Protection Agency, Region II, New York, NY. Cooperative National Wetlands Inventory Publication Hadley, MA: U.S. Fish and Wildlife Service, Ecological Services. ———. 2005. Assessing cumulative loss of wetland functions in the Nanticoke River Watershed using enhanced National Wetlands Inventory data. Wetlands 25: 405–419. Tiner, R. W., J. Q. Swords, and B. J. McClain. 2002. Wetland Status and Trends for the Hackensack Meadowlands. An Assessment Report from the U.S. Fish and Wildlife Service’s National Wetlands Inventory Program. Hadley, MA: U.S. Fish and Wildlife Service, Northeast Region. Van Dyke, E., and K. Wasson. 2005. Historical ecology of a central California estuary: 150 years of habitat change. Estuaries 28: 173–189. Zedler, J. B., 1996. Coastal mitigation in southern California: The need for a regional restoration strategy. Ecological Applications 6: 84–93.

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14

The Use of Science in the Restoration of Northeastern U.S. Salt Marshes John Teal and Susan Peterson The long-term threats to East Coast U.S. salt marshes are from sea-level rise, invasive species, nutrient additions, top predator removal, increasingly ferocious storms as a result of climate change, and sudden salt marsh dieback. Salt marsh conservation provides ecological functions and societal values, including nursery areas for fish and shellfish; feeding and refuge areas for fish, birds, turtles, and mammals; storm protection for onshore areas; nutrient reduction; and aesthetic values. Scientists contribute to preservation and restoration by elucidating the functions, describing the changes likely from the long-term threats, and suggesting and investigating methods by which they may be overcome. Since humans must make decisions with imperfect knowledge, we recommend restoration and protection programs be planned as ecosystem experiments, using ecological engineering and adaptive management to guide our actions. All of this must be done in a broad political context that recognizes that humans are dependent on coastal resources, that coastal management will require hard work and imagination, and that sustainability requires transdisciplinary approaches.

Our understanding of salt marshes has changed since Europeans settled in North America. Marshes were valued as hay fields in New England and as farmland, once diked, in the region of Delaware Bay. As uplands were settled and cleared for farming, and as cities grew in size, salt marshes came to be thought of as useless, mosquito-ridden areas to be filled for more profitable uses (Teal and Teal 1969). In the mid–twentieth century, scientists began to understand the other values provided by marshes: nourishing fish and shellfish and offering storm protection for adjacent uplands. Accordingly, we sought to

protect them for these values (Teal 1986). Now they are threatened by other changes we describe here. In the United States, local, state, and federal laws protecting tidal marshes need to be strengthened and enforced and policies need to be developed that ensure the long-term survival of these coastal features. While current laws reduce the likelihood that marshes may be directly altered, as by dredging or filling (Massachusetts 1972), they are not sufficient for long-term preservation of the ecological functions and societal values provided by coastal wetland systems. Sustained conservation of these resources requires that society 267

address threats less obvious than bulldozers, draglines, and backhoes. Using historic records and maps to compare with current conditions, Bromberg and Bertness (2005) estimated that the New England coast has lost about 37 percent of its original salt marsh in the last two hundred years. Rhode Island showed the highest loss (53 percent), followed by Massachusetts with 41 percent loss. Boston’s loss amounted to 81 percent from filling wetlands that are now occupied by urban buildings. Most of Maine’s coast is not hospitable to salt marsh formation and therefore lost very little to human activity. Maryland lost about 10 percent of its salt marshes in the first eighty years of the twentieth century (Spaur et al. 2001). Georgia and South Carolina, with large amounts of salt marsh, have lost much less than the New England states, although current development along their coasts threatens increased losses in the future (Sanger and Holland 2002).

LONG-TERM THREATS TO SALT MARSHES Over the past eighty-plus years, botanists, biologists, hydrologists, geomorphologists, and artists have studied the structure and function of northeastern salt marshes. The result is a body of work that allows us to predict the effects of certain types of changes to both structure and function. We discuss in some detail the following changes: sea-level rise and its effect on vertical accretion, impediments to landward transgression, invasive species, and mosquito ditching. To preserve and restore coastal wetlands and maintain their relationship to the coastal ocean, we need to be familiar with geologic processes (sedimentation, erosion, glaciations, rebound, etc.) as well as with the plant and animal community structure and function. Scientific research is critical to understanding the causes and consequences of marsh loss and the degradation of ecological functions that marshes perform. With this information, we can (1) investigate methods to overcome marsh loss and functional degradation and (2) design policies to try to 268

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ensure the survival of these ecological functions into the next century and beyond. SEA-LEVEL RISE AND VERTICAL ACCRETION

Marshes accrete vertically in two ways: (1) by trapping sediment delivered by tidal waters and storms and (2) through preservation of belowground production (roots and rhizomes) in the form of peat (Redfield 1972; Turner, Swenson, and Milan 2000; Turner et al. 2006). There is evidence from the fifteenth and seventeenth centuries that, when hurricanes caused significant erosion in Connecticut marshes, the marshes recovered rapidly as mud and peat accumulated (van de Plassche et al. 2006). Human activities and natural hydrologic modifications (e.g., barrier beach excavation and/or erosion) and nutrient enrichment influence these processes both directly and indirectly. Increase in nutrients reduces the fraction of production marsh plants allocate to belowground biomass (Valiela et al. 1976; Morris 1982) and therefore the amount of peat that is produced. Long-term monitoring of marsh plain elevations relative to local sea level has been implemented in various marshes along the East Coast (Roman et al. 2006). This information is critical to track whether marshes are keeping up with sea-level rise. The oldest tidal marshes in the Northeast began forming four to five thousand years ago, as sea-level rise slowed from 3 to 4 mm yr⫺1 to about 1 mm yr⫺1 (Redfield 1972; Donnelly et al. 2004). The presence of Spartina peat in over fifty meters of water offshore of Nantucket demonstrates that salt marshes existed prior to four thousand years ago but did not sustain themselves in the face of rapid postglacial sea-level rise. Since the mid-1800s, sea level rise in Connecticut has accelerated to 2.0 to 2.5 mm yr⫺1 (Donnelly et al. 2004). Teal and Howes (1996) found a 2.5 mm yr⫺1 rise in sea level between 1971 and 1991 in Massachusetts and a weak, negative correlation between sea level and Spartina production. The southeastern Massachusetts and Maine coasts are subject to the current eustatic sea-level rise of about

2 mm yr⫺1. The rates for the mid-Atlantic coasts are higher—about 3 mm yr1, due to local subsidence (Larsen, Williams, and Anonymous 2004). With global warming, the eustatic rate will increase over the next several decades. For the past fifteen years, the tide gauge at New London, Connecticut, has recorded a rate of 4.1 mm yr⫺1 (Donnelly and Cleary 2002). The Intergovernmental Panel on Climate Change (Alley et al. 2007) projects a global rate of 3 mm yr⫺1. They exclude how this rate could be affected by changing rates of glacial melt in Greenland and Antarctica that could greatly affect sea-level rise (Vaughn and Arthern 2007). If the rise rate accelerates to over 10 mm yr⫺1, marshes will have difficulty in keeping up, and greater losses can be expected (Orson, Panageotou, and Leatherman 1985). SEA-LEVEL RISE AND LANDWARD MARSH DEVELOPMENT

As sea level rises on an unmodified coast, upland vegetation on the landward side of the marshes is drowned, and the marsh vegetation replaces it. This process may result in a loss or gain of marsh area, depending on the slope of the land, the ability of the marsh to maintain its elevation in relation to sea level, and the extent to which the seaward edge of the marsh is protected from erosion. Historically, as sea level has risen and tidal influence moves landward, tidal marsh plants replace upland species, eventually converting fields and forests to salt marsh; in older systems, two or more meters of marsh peat may cover old upland soils (Redfield 1972). In the Delaware Estuary, what was a cedar swamp in 1850 is now a salt marsh where standing dead cedar trees are still visible (Philipp 2005; fig. 14.1). Salt marsh transgression allows marsh formation to compensate for marsh loss from erosion on the bay front side of these systems. Without room for landward movement, many marshes will eventually be lost, even if vertical accretion does keep pace with sea-level rise. This is the case along most of the Northeast coast, where shoreline development (buildings, roads, railroads, houses) has left little or no room for

FIGURE 14.1 A dead white cedar still standing in what is now a fully developed salt marsh, Cedar Creek, New Jersey. Photo courtesy of J. Teal.

landward marsh movement in the face of rising sea levels. And as sea levels rise, there will be pressure to protect the human-developed environment with levees or dikes. Hardened coastlines absolutely prevent marshes from moving landward and thus accelerate erosion of the seaward edge and overall loss of marshes. INVASIVE SPECIES

Invasive species of plants and animals, some introduced deliberately by humans, have changed tidal marsh and estuarine systems throughout the world. By far the most important invasive plant in Northeast tidelands is Phragmites australis (common reed grass). Phragmites has been a constant, usually minor, component of brackish tidal marsh vegetation as far back as peat records exist, four to five thousand years (Orson 1999). But about a hundred years ago, the European haplotype of Phragmites began to appear in herbarium collections. The European strain is highly invasive in the northeastern United States, thriving in disturbed upland borders of salt marshes and undisturbed brackish (mesohaline to oligohaline) tidelands. Once established, it can spread into more saline soils and into soils with higher sulfides (Chambers et al. 2003). It outcompetes native species, including native Phragmites, rapidly forming dense monocultures and erasing the natural vegetation mosaic (Burdick, Buchsbaum, and Holt 2001; Philipp and Field 2005; Silliman

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FIGURE 14.2 This tidal creek was covered over by Phragmites rhizomes; it is reopening two years after the Phragmites was killed. Photo courtesy of J. Teal.

and Bertness 2004). Nutrient runoff from shoreline developments in the Northeast has caused replacement of normal high marsh plant species (e.g., Spartina patens) by S. alterniflora, which has moved up from lower marsh, and by Phragmites, which has moved down (Bertness, Ewanchuk, and Silliman 2002; Silliman and Bertness 2005). This Phragmites invasion reduces native vegetation and flattens the marsh plain by Phragmites rhizome growth (Weinstein and Balletto 1999). Epifaunal communities are less diverse and abundant in Phragmites compared with S. alterniflora marshes (Robertson and Weis 2005). Once Phragmites is the dominant vegetation in a marsh, reproduction of the common marsh fish Fundulus heteroclitus is severely reduced (Able and Hagan 2003). The dense mat of rhizomes steepens banks of tidal creeks and reduces feeding areas for shorebirds and refuge areas for small fishes (Weishar and Teal 2007). Rhizome growth also results in the filling of small tidal channels by sediment (Windham and Lathrop 1999; Lathrop, Windham, and Montesano 2003; Teal and Peterson 2005a). Even microbial populations are affected. For example, the microbial populations of 270

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Phragmites marshes are less effective in removing bromine from a common pollutant, the halogenated flame retardant TBBPA, than are S. alterniflora marshes (Ravit, Ehrenfeld, and Haggblom 2005). All of these effects result in a diminution of marsh function (fig. 14.2). Although usually described as a liability to coastal systems in the Northeast, there are ways in which Phragmites may be considered to be beneficial. For example, S. alterniflora moves more metal pollutants into its leaves and therefore into the food web than does Phragmites (Windham, Weis, and Weis 2003). Some believe that Phragmites growth slows the erosion of marsh edge by the binding of sediments by its roots and rhizomes, while others believe it speeds erosion up by making banks steep and concentrating wave force immediately under the root mass resulting in undercutting and caving in (Weishar and Teal 2007). Animal invasives have had and are having important effects on salt marshes. We can probably never accurately sort out all the effects of early invasives such as European periwinkles (Littorina littorea) and the green crab (Carcinas maenas) since we have no data on the marshes

before their introductions. Bertness (1984), in a Littorina exclusion experiment, showed that Littorina inhibited S. alterniflora growth and expansion both by reducing sediment accumulation on hard shores and by eating shoots and rhizomes. We know that the green crab is an active predator and feeds on snails and fiddler crabs and thus assume its introduction had an effect on marsh structure (Leonard et al. 1998). We should be able to elucidate the effects of recent invasives such as Hemigrapsis sanguineus, using well-replicated exclusion experiments and comparing sites where it is present with those where it is absent. We can also evaluate impacts in locations where it is increasing in abundance by using data from before it was introduced. We know Hemigrapsis competes with native crabs. It reduces the abundance of green crabs on its rocky shore habitat (Lohrer and Whitlatch 2002) and should therefore affect adjacent salt marshes. It is omnivorous, so its presence can also affect the distribution and abundance of other marsh species (Ledesma and O’Connor 2001; O’Connor 2001). MOSQUITO DITCHING

Mosquito control has altered the surface of most marshes in the northeastern United States, with resulting changes in vegetation distribution, salinity gradients, and fish use. Mosquitoes (as well as other insects undesirable to humans such as various species of tabanids) are common on salt marshes. Salt marsh mosquitoes generally belong to the genus Ochlerotatus, with O. sollicitans (formerly Aedes sollicitans) the most common in the Northeast. The females lay their eggs singly on wet mud just above the water line. The eggs develop into adults within a few days in hot weather once the mud is flooded. Marsh pools were thought to be mosquito-breeding areas, and this presumption led to aggressive programs to drain the pools by digging ditches on the marsh plain to intersect with main drainage channels (Miller and Egler 1950; Ranwell 1972). Most aerial photos of New England marshes show active and/or historic ditches. Since this practice became common in

the 1930s, we have learned that, if marsh fishes such as mummichogs (Fundulus heteroclitus) can get into the pools and survive there, most mosquito larvae enter the food chain rather than hatch out as pests to nearby human populations (Teal and Teal 1969). Pools occupy about 9 percent of the surface area of New England marshes, with fewer pools on marshes that have been ditched for mosquito control (Adamowicz and Roman 2005). Similar pool areas (7.9 percent) in unmanaged marshes were found in New Jersey (Lathrop, Cole, and Showalter 2000), again with fewer pools in ditched marshes. Marsh pools are a separate environment on the marsh plain and support a different assemblage of fishes than the creeks and main channels and so they are included in some restoration plans (Able, Smith, and Hagan 2005). Marsh pools are less common in southeastern marshes (Wiegert and Freeman 1990). In a Delaware Bay salt marsh restoration (Teal and Peterson 2005b), we observed sites where spring plowing (an effort at Phragmites control) left many small puddles on the marsh surface. We found many larval mummichogs in these puddles, but never a mosquito larva (J. Teal, personal observation, 2007). Mosquito ditching has also modified marsh sediments and marsh plant communities so that the poorly drained panes in the high marshes, which had a more diverse flora, have become rare. As the drained sediments developed higher redox potentials, they were invaded by Spartina patens that exclude panne species (Ewanchuk and Bertness 2004). As ditches have become abandoned or allowed to become shallow, panne forbs are beginning to return in some areas (J. Teal, personal observation, 2006). The largest impact of extensive ditching on northeastern marsh plant communities was most ostensibly a switch from spatial dominance by the low marsh species S. alterniflora to dominance by high marsh species, S. patens and Juncus gerardi, that outcompete S. alterniflora for nutrients under well-drained conditions (Ewanchuk and Bertness 2004).

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PHRAGMITES MARSHES AND MOSQUITO CONTROL

The Cape May County Mosquito Control Commission reported that three thousand pounds of organophosphate larvacide was used on a 233-hectare New Jersey salt hay farm seven to nine times a year. They found that, if the larvacide was not used to coincide with hatching cycles, the adults could spread over a radius of twenty-five miles, requiring even a greater amount of insecticide. They noted that the larvacide sprayed from the air did not penetrate Phragmites stands very well. Once this marsh was restored with normal tidal cycles flooding the marsh plain, a range of creeks sizes with appropriate sinuosity, and pools where mummichugs could thrive, no mosquito spraying has been needed because natural predators now control the mosquitoes. SUDDEN SALT MARSH DIEBACK

The dieback, sudden and otherwise, currently being studied on the East and Gulf coasts (Mackinnon and Huntington 2005; Silliman et al. 2005; Adamowicz 2006; Ogburn and Alber 2006; Smith 2006) may be caused, but is certainly exacerbated, by drought. Drought may become more common with climate change (McKee, Mendelssohn, and Materne 2004). While drought is a causal agent in some cases, there may be other causes that are unequally distributed across the landscape such as nutrient enrichment, top predator removal and subsequent overgrazing (Silliman et al. 2005), pathogen additions, tidal restrictions, and others. A recently suggested possible cause is infection by a nematode, not previously found in New England, which may facilitate infections by the pathogenic fungus Fusarium (LaMondia 2007).

TIDAL MARSH RESTORATION AND PRESERVATION The long-term threats described here no longer include most of those that were common from the time of European settlement of the Northeast: building railroad and road bridges, 272

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filling for buildings, dredging for marinas or shipyards, and diking for farming. One way to protect marsh values and functions is to look at those impaired by past actions and find ways to restore them. We use restoration to mean “returning an ecosystem to a close approximation of its condition prior to disturbance. Accomplishing restoration means ensuring that ecosystem structure and function are recreated or repaired, and that natural dynamic ecosystem processes are operating effectively again” (National Research Council [NRC] 1992). A standard for tidal marsh restoration is the return of ecological functions. These functions include diversity and abundance of plant and animal communities; biogeochemical processes; and biotic, chemical, and physical links with estuarine and near-shore waters. For regulators and managers, restoration is usually measured as the return of “natural” or “native” vegetation because plant communities are relatively easy to assess and provide a convenient standard for permitting any other regulatory activities. It is important to know what organisms are key to the restoration of an ecosystem so that the process may proceed as effectively as possible (Suding, Gross, and Houseman 2004). Spartina species are the foundation plants in salt marshes that determine colonization by other plants (Proffitt et al. 2005) and eventually by animals. So getting the vegetation right is key. There are or may be positive interactions among organisms in a system being restored that are important to success. In some salt marsh restorations, the plans specify a vegetation-planting program. In those cases, it may be necessary to ensure that sediments have appropriate nutrients or that microrhizal fungi are present or are introduced (Halpern et al. 2005). Where the landscape in which the restoration is to take place has abundant salt marshes with healthy vegetation, allowing the system to self-design is sufficient (see later discussion). Vegetation coverage can be measured using aerial photography, which makes it relatively inexpensive for largescale restorations. It can also be measured from

satellite images, although this is expensive at the resolution generally needed. Remote sensing is especially useful in early stages of restoration when vegetation mosaic is highly broken up and difficult to measure by ground transects. Calibration of remote data by fieldwork is always needed at some level to confirm the photographic interpretations. Vegetation measurements integrate a suite of environmental factors and are often, but not always, a reasonable analog for ecological functions. For example, on examination of eight pairs of restored/control marshes in North Carolina, algal growth was found to be variable and associated with nutrients in the marsh sediments rather than with marsh status as control or restoration (Zheng, Stevenson, and Craft 2004). RESTORING TIDALLY RESTRICTED MARSHES

Returning appropriate tidal hydrology (most commonly by increasing culvert size and/or removing tide gates) increases soil salinity and reduces peat Eh, stressing or killing Phragmites and allowing reestablishment of native salt marsh grasses either naturally, from seeds from adjacent marsh, or by seeding using commercial sources of seed. Spartina seeds are viable for less than a year. The rate and success of restoration depend in large part on conditions behind the barrier. Typically, barriers have reduced the volume of sediment that would normally have been distributed on the marsh plain and allowed marsh peats to compact and oxidize. The resulting elevation may be too low to support marsh vegetation because most of the plants will not survive below about mean sea level. Barriers also are likely to have reduced the salinity of the system, allowing other vegetation to invade. Both the lack of sediment and change in sediment decay processes reduce the rate at which the elevation in a partially isolated marsh keeps up with sea-level rise. Barriers may also trap water on the marsh and drown some of the plants. Marsh system response to restoration of tidal hydrology has been modeled (Boumans, Burdick, and Dionne 2002). In one case in New Hampshire, restoration of tidal exchange

resulted in good restoration in two years. At the same time, an inadvertent restoration in southern Maine required eight years for a partial restoration (Burdick et al. 1997). In a Massachusetts marsh, it took more than four years for complete restoration (Buchsbaum et al. 2006). A twenty-hectare Connecticut marsh was tidally restricted for thirty-two years, during which time Typha covered 74 percent of the marsh plain. Ten years after tidal exchange was restored, 45 percent of the marsh plain was covered with S. alterniflora and 20 percent with S. patens (Sinicrope et al. 1990). Other restored Connecticut marshes where changes in fauna were tracked for twenty years had considerable variation in faunal recovery. Swamy et al. (2002) note that faunal recovery at these sites required at least two decades and was not necessarily in synchrony with vegetation recovery (Warren et al. 2002). PLANNING LARGE RESTORATION PROJECTS

Large restoration projects are just as or more challenging as small ones, with each requiring planning, design, and modeling to have a reasonable chance for success, as outlined by Simenstad, Reed, and Ford (2006). The Delaware Bay salt marsh restoration, which we describe here as a model for large-scale restoration, was done by Public Service Enterprise Group (PSEG). The program started in 1993 and involved the restoration of about 4,550 hectares of marshes: 1,780 hectares of diked salt hay farms in the mesohaline part of the estuary and 2,770 hectares of Phragmites-dominated sites in the brackish portion of the estuary. The restoration was done as mitigation for oncethrough cooling at the Salem Generating Station (Balletto et al. 2005). In planning for the PSEG program, ten restoration principles were developed (Weinstein, Phillipp, and Goodwin 2001). These were (1) state project goals clearly, as agreed to by the stakeholders (e.g., citizens, property owners, etc., and make the goals site-specific and realistic; (2) restore degraded sites rather than create new wetlands; (3) select sites in a landscape ecology framework;

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FIGURE 14.3 The Maurice River salt hay farm. This diked marsh looks like a farm field rather than a salt marsh flooded by bay waters twice a day. Photo courtesy of PSEG.

(4) ecological engineering practices should apply; (5) restored sites should be self-sustaining and managed using adaptive management to reach stated goals; (6) site monitoring should be planned and implemented, and last until success is assured; (7) success criteria should include functional as well as structural components (framed by a “bound of expectation”); (8) consider people and property—a management plan should be developed that protects offsite elements (e.g., upland flooding, salt intrusion into wells, septic systems); (9) where possible, sites should be developed with conservation restrictions to ensure their protection in perpetuity and to protect adjacent property; and (10) site plans should encourage public access for sustainable uses. The three salt hay farms, of approximately 150 (Dennis), 460 (Maurice River), and 1,170 hectares (Commercial), had been diked and isolated from Delaware Bay for at least fifty years except for occasional flooding during storms. There was no significant tidal exchange with the estuary. Most of the tidal creeks in the farms were filled to facilitate hay harvest, resulting in relatively firm fields crossed by drainage ditches and roads. Because the farms were isolated from the bay, the salinity in the sediments declined. Rain also freshened the soil. Areas of higher elevation, once dominated by S. patens, were invaded by Phragmites. The sediment lev274

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els decreased due to oxidation of the peat, lack of new sediment from the bay, and compaction (fig. 14.3). Swedish and Dutch settlers started the practice of diking the Delaware Bay marshes in New Jersey in the seventeenth century. The marshes in this restoration had various histories (Philipp 2005), but Commercial had been diked for so long that a major part of it was below the level at which S. alterniflora would grow, with some areas nearly three feet below that level (Weinstein et al. 2000). Teal and Weinstein (2002) described the steps needed for restoration planning: • Establish goals for the restored marsh that include structural and functional characteristics required for success and an acceptable time line. • Elucidate the methods for choosing the goals and time line. • Plan how propagules of the various living ecosystem components will become established. • Plan, design, and model how proper tidal circulation will be achieved. • Outline construction and management techniques. • Establish criteria for and choose reference sites; develop data collection protocols, data

review processes, and adaptive management plans. • Plan for and manage oversight—advisory committees of independent experts, regulators, and stakeholders. After about 1950, many of the diked marshes of Delaware Bay that were no longer actively farmed and maintained had been opened to the bay by storms and restored by natural processes. Aerial photographs taken over that period show that it took about ten years for these systems to become vegetated and that there were considerable differences in the ways they “self-restored.” These data were the basis for hypothetical restoration trajectories for marsh plain vegetation coverage. Successful restoration criteria included not only the percent of the marsh plain vegetated but also vegetation composition and tidal channel density, length, bifurcation, and sinuosity (Weinstein et al. 1997). The restoration trajectory and success criteria were reviewed and approved by an oversight committee composed of federal and state regulators; outside experts in marsh ecology, fish ecology, and hydrodynamics; and local stakeholders and their local government representatives. The PSEG restoration sites were situated in a landscape of tidal marshes on both sides of the bay, including ones immediately adjacent to the restoration sites. Tides and currents transport seeds of marsh plants, so it was decided not to plant the sites with S. alterniflora but to rely on natural processes for revegetation (Hinkle and Mitsch 2005; Teal and Weishar 2005). The regulators required some experimental planting, but except on high marsh areas, planted areas were overwhelmed by the natural revegetation (J. Teal, personal observation). Comparisons with neighboring marshes provided a model for the numbers and sizes of tidal channels needed for the restored sites. Using these data, hydrodynamic models guided final construction designs (Weinstein et al. 2000; Weishar, Teal, and Hinkle 2005a). Stakeholders were invited to review these

designs and offered comments and contributions of local knowledge (Weishar et al. 2005a). The major primary and secondary channels at the restoration sites were dredged because the farmers had filled in nearly all of the original channels. With marsh plain elevations lower than where salt marsh grasses could survive, the design relied on restored tidal flow through the new primary and secondary channels to deposit adequate amounts of sediment on the restoration sites. Smaller channels were allowed to develop by themselves in a process of ecological engineering or self-design (Teal and Weishar 2005). Reference marshes near the restoration sites that had undisturbed appearances and had undergone natural restoration were chosen (Weinstein et al. 1997). An adaptive management process was agreed on that involved regular field visits to the sites to observe progress and decide whether thresholds that would require intervention had been crossed. The field visits were augmented with annual aerial mapping of vegetation and drainage patterns combined with ground calibration transects (Weinstein et al. 1997; Teal and Weishar 2005). The formal adaptive management process defined “triggers” for intervention and required consultation with regulators and advisory committees before deciding on and taking action. Another process called restoration management addressed potential problems before a trigger was reached (Teal and Weishar 2005). The final restoration plans approved by regulators specified oversight and advisory committees to review progress of the restorations and the biological monitoring program (Balletto et al. 2005). Construction planning, design, engineering, modeling, and permitting were completed in 1995, and construction began. The major channels were dredged while the sites were flooded; the dredged material was then strategically placed in areas designated to become high marsh. Where transitional areas could not be purchased, upland dikes were constructed to prevent flooding or other damage to

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FIGURE 14.4 The Maurice River Salt Hay Farm seven years after tidal action, which allowed natural seeding of Spartina alterniflora from adjacent marshes, was restored. In the foreground are boat launches and a bird observation platform built during restoration. Photo courtesy of PSEG.

neighboring inhabited areas, such as saline intrusion to wells or flooding of septic systems. Dredging was overseen to ensure that operators could achieve desired sinuosities of new channels or, where this was impossible, to modify plans (Teal and Weishar 2005). Once dredging was finished, bayside dikes were opened, and natural tidal exchanges resulted. The entire construction process took two to three years, and, during this period, viewing platforms and walkways were installed for the public to observe restoration progress and visit the sites. As a final act of the restoration process, all sites were permanently protected by conservation easements (Balletto et al. 2005). LONG-TERM MONITORING AND ADAPTIVE MANAGEMENT

In the PSEG program, experts in marsh ecology and hydrodynamics monitored each site quarterly. As the sites were large, this required about a week each quarter, with access to the marshes gained by boat, tracked vehicles, and airboats. As the areas became more vegetated, traveling on marshes could not be done without damage to plants, and thus later in the restoration process, there was more dependence on aerial photographs for progress monitoring. Twice a year the oversight committees visited selected sites, but there were more frequent visits to the sites by project personnel, especially to the land276

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ward edges, to inspect dikes, upland drains, and access points. After restorations were well established (five to eight years), monitoring by expert groups was reduced to twice per year—in the early spring when winter damage could be examined, and in the fall at the end of the growing season (fig. 14.4). The monitoring team recommended a number of interventions as adaptive management. The few natural creeks remaining at the sites had elevated “natural levees” on the creek banks that had been invaded by Phragmites. The rhizome mats produced by these plants inhibited development of drainage creeks. Cuts through these levees were made at some sites to enhance drainage (Teal and Weishar 2005). In the euryhaline conditions of the salt hay farms (fifteen to twenty parts per thousand), Phragmites is controlled by salinity. It is not clear whether osmotic pressure, sulfide, or a combination of both reduce Phragmites vigor (Chambers, Mozder, and Ambrose 1998; Howes et al. 2005). In such marshes, Phragmites gradually declines and disappears in a process that may take a decade or more. Adaptive management combined with ecological engineering may allow unexpected changes at a site to be viewed as natural experiments—ones created by storms or other unplanned events. The PSEG restoration was

on a scale too large to be replicated and for which several control sites were impossible because of size and cost. Ideally, of course, an experiment should be replicated and have controls, but it would be tragic not to observe and carefully interpret the “natural experiments” that occur. An example at Commercial occurred when an internal berm, constructed to control tidal circulation as indicated by the hydrodynamic models, was breached by a storm. Careful monitoring with tide gauges and observations showed the circulation was improved by the breach, and it was allowed to develop (Teal and Weishar 2005). Annual aerial color infrared photography processed to digital orthophotographs monitored vegetation coverage and gross composition on both restoration and reference sites. Quantitative ground transects were used to obtain more detailed plant species data and net production. Adequate sediment deposition to support S. alterniflora took a few years at some sites. The sediment had first to settle and then dewater before it was stable enough to support Spartina seedlings. Once the seedlings were established, dewatering by evapotranspiration sped up the process, and vegetation spread rapidly (Hinkle and Mitsch 2005). Algal production and benthic invertebrate populations were also measured (PSEG 1995–2000). Some invertebrates, the more opportunistic species such as the capitellids, can develop populations in newly created marshes even before, or without, return of salt marsh grasses, while other invertebrates may require two decades or more before reaching equivalence with nearby control marshlands (Levin et al. 1996). Although abundances may vary from year to year, dominance was found to stay constant in South Carolina marshes (Alphin and Posey 2000). In created marshes, development of benthic faunas can take even longer, more than twenty-five years for oligochaete populations, to develop (Craft and Sacco 2003). RESTORED SALT MARSH FUNCTION

The diked salt hay farms had very few fish until the dikes were breached. Fish moved into the

restored marshes with the first rush of water when the bayside dikes were opened. One year after the marsh was opened, Fundulus heteroclitus (the most abundant resident marsh fish) was resident and breeding in the restored marsh in numbers equivalent to those in the reference marshes (Teo and Able 2003a, 2003b). At the same time, juvenile sciaenid fishes were using the restored marsh in numbers similar to those in reference marshes and were feeding and growing similarly as well (Nemerson and Able 2005). Use of the restored marshes by older sciaenids and other predatory fishes was also similar to reference marshes (Able, Nemerson, and Grothues 2004). Blue crabs were as abundant and grew as fast or faster than in the reference marshes (Jivoff and Able 2003). DETERMINING WHEN RESTORATION IS COMPLETED

The purpose of the PSEG restoration program was to provide marsh habitat and food for Delaware Bay fish and offset potential impacts of the Salem Generating Station. The monitoring programs produced data that showed that the marshes were structurally and functionally equivalent to adjacent reference marshes. The restored marsh plains were vegetated with desirable marsh species; marsh creeks, channel, and rivulets had developed with the sinuosity typical of control marshes; and fish were resident in and breeding within the sites. It took two years at Dennis and Maurice River for the tidal creeks to develop to where analysis of the orthophotographs showed that drainage structure of these sites was comparable to reference marshes (Weishar et al. 2005b). The vegetation coverage at these two sites achieved success in four years postconstruction. Vegetation coverage at Commercial was slower, but after seven years, it is on the trajectory for success following a plan that allows twelve years for completion. This large site had the lowest elevation and the fewest tidal connections to the bay and therefore required the most time to accumulate sediment so that S. alterniflora could invade and spread (Teal and Weishar 2005).

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RESTORATION IN PHRAGMITES-DOMINATED SITES

Up in the brackish part of the estuary, Phragmites is not inhibited by salinity and, once it has invaded a site, excludes other vegetation. The Phragmites-dominated sites in the PSEG program had been diked, but the dikes had failed and had some tidal exchange with the bay. Phragmites coverage began on the remnant dikes and subsequently spread over most of the area of those sites (Philipp 2005). The restoration process was to restore the desirable native vegetation and marsh function. The original plan for Phragmites control involved helicopter spraying with glyphosate annually until control was achieved. The plan relied on natural seeding to restore desirable plants. One spray application killed aboveground and immediately adjacent rhizomes, but buds on rhizomes farther from the killed ramets were released from apical dominance and sprouted, so that within a year of the first spray, Phragmites again dominated. During these initial sprayings, environmental activist opposition to spraying an herbicide developed. Concern was not alleviated by scientific studies of the effects of glyphosate and/or the surfactants used with it (Williams, Kroes, and Munro 2000; Solomon and Thompson 2003). As a result, a test program was instituted to see if other methods, by themselves or in combination with herbicide, could be effective. These included mowing, rhizome ripping, surface scarification, and grazing by goats. The final conclusion from these studies and other experience was that herbicide application was the only effective technique for Phragmites control (Teal and Peterson 2005a). The compromise reached in New Jersey continued to rely on herbicide, but the application rate and area were reduced, resulting in a longer time to restore the marshes. Data on mummichog feeding in former Phragmites marshes showed populations were similar to those in unmodified Spartinadominated marshes within a year, long before all of the vegetation changes had taken place (James-Pirri, Raposa, and Catena 2001), 278

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although there was little change in the fishes living in the deeper tidal creeks (Grothues and Able 2003). Phragmites domination also changes the arthropod assemblages on the vegetation; for example, surface feeders disappeared and were replaced by insects feeding inside Phragmites stems. Stable isotope composition of the arthropod assemblage in the Phragmites stands were different from those in the Spartina areas and showed that those in the Phragmites stands fed on detritus and benthic algae. Once the restoration removed the Phragmites and Spartina was restored, the arthropod assemblage became that characteristic of a Spartina marsh, and isotope studies showed that they fed on Spartina (Gratton and Denno 2005, 2006).

CONCLUSION The long-term threats to northeastern U.S. salt marshes are from sea-level rise, invasive species, nutrient additions, increasingly ferocious storms as a result of climate change, and sudden salt marsh dieback. Sea-level changes over time are normal. Marshes would survive if it were not for the human modifications to the coastal environment. Residents of coastal communities have used myriad techniques to stop the effects of sea-level rise, most of which have had serious consequences for ecosystem functions. Revetments, jetties, groins, sea walls, levees, and so forth, may preserve adjacent upland features for a few years, but they generally result in loss of other nearby coastal features by changing sediment supply or transport, reducing wave attenuation and buffering capacity, and causing habitat loss (NRC 2006). With rapid sea-level rise, it would be important to reduce the amount of shoreline armoring to ensure adequate sediment supply to protected or restored marshes. Invasive species are probably normal if one considers geologic time. As the globe warmed and cooled, species shifted accordingly. What we have seen in the last fifty years is rapid change in species, loss of indigenous populations, and functional changes that are measured in decades rather than eons. Phragmites can be managed

and traditional salt marsh function restored, but success requires large-scale planning and implementation. For example, it has been commonplace to add fill to the land at the marsh/upland edge to build roads, bridges, malls, condominiums, and houses. These actions may increase freshwater table levels, which, in turn, reduce salinity at marsh edges, thereby favoring Phragmites (Burdick and Konisky 2003). Changes in state and local policies and permits would be needed to reduce inadvertent creation of Phragmites habitat. Other plant and animal species are more difficult to manage. Scientists contribute to preservation and restoration of salt marshes by elucidating the functions, describing the changes likely from the long-term threats, and suggesting and investigating methods by which they may be overcome. Sudden salt marsh dieback in the northeastern United States is a phenomenon that has many scientists working to determine the causes, time lines, extent, and persistence of the problem. Until we understand the cause (or causes), it will be difficult to recommend preservation techniques. The rate of marsh loss in Long Island Sound and in Jamaica Bay is so high that restoration may be the only option for preserving their ecosystem functions (Jamaica Bay Proceedings 2004). The Delaware Bay salt marsh restoration is a good example of restoration principles. Planning, construction, and management processes were documented in 2005 in Ecological Engineering, volume 25, issue 3, so that other practitioners could avoid mistakes and improve their likelihood for success. Careful planning, ecological engineering, and adaptive management were key to those successful restorations; it is hard to imagine any restoration succeeding without them. We have shown that restoration is one way to maintain salt marshes along the East Coast of the United States. Another way is to buy and protect the adjacent upland to allow marshes to retreat landward without losing area, but the high cost of acquiring adjacent upland is a deterrent to this method of marsh protection.

Having looked at the long-term threats to salt marshes, we recognize that it is inevitable that some marshes will be lost. Our responsibility is to describe how marshes work, provide data to help choose which marshes can or should be preserved or restored, and describe how that may be done. But we have not discussed how to pay for the protection or restoration. The PSEG restoration referred to here was a negotiated settlement in the process of acquiring a National Pollutant Discharge Elimination Service (NPDES) permit under the Clean Water Act. The agreement included provisions for the company to restore thousands of acres of salt marsh and to purchase and permanently protect uplands adjacent to the restorations to allow for movement of the marsh as sea level rises. Unfortunately, use of restoration measures to address cooling water intake issues has been found to be inconsistent with the Clean Water Act (United States Court of Appeals 2007). This decision, unless addressed in an amendment to the Clean Water Act or a successful appeal, removes a motive and source of significant funds for salt marsh restoration and preservation. Since humans must make decisions with imperfect knowledge, we recommend restoration and protection programs be planned as well-designed, ecosystem experiments, using replication and proper controls when possible and ecological engineering and adaptive management to guide our actions. All of this must be done in a broad political context that recognizes that humans are dependent on coastal resources, that coastal management will require hard work and imagination to find funding to carry out the work, and that sustainability will require transdisciplinary approaches (Weinstein et al. 2007). Acknowledgments. We thank all the people at Public Service Enterprise Group we worked with on the Delaware restoration project, especially John Balletto, Brenda Evans, Joe Klein, Jeff Pantazes, and Ken Strait. We also thank other colleagues: Ken Able, Maureen Vaskis Heimbuch, Ray Hinkle, Kurt Philipp, Michael Weinstein, and Lee Weishar.

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Conserving the Diverse Marshes of the Pacific Coast J. C. Callaway and J. B. Zedler Pacific Coast salt marshes tend to be discrete and unique entities because of their location on the leading edge of the continent and the specific features of each watershed. Compared to those on the Atlantic and Gulf Coasts, western salt marshes are smaller, more isolated, and influenced by a broad range of precipitation. Direct modification, due to urbanization, agriculture, and other development, has eliminated salt marsh habitat, particularly in southern California, San Francisco Bay, and Puget Sound, while substantial efforts have been undertaken recently to restore salt marshes and offset impacts from development. Hydrology and sediment dynamics strongly influence salt marshes, and many Pacific Coast marshes have highly modified hydrologic conditions. Tidal connections for some systems have been restricted by highways, railroads, and development, while others have increased tidal influence due to dredging for harbors and marinas. Land use changes in coastal watersheds have altered sediment dynamics and vegetation in downstream salt marshes. Sediments delivered to southern California salt marshes by catastrophic flooding have filled channels and elevated marsh plains. Other marshes have experienced conversion of salt marsh to mudflats, due to subsidence or scouring associated with increased tidal flows. Invasive plants are another common problem. Invasive Spartina species have colonized mudflats and marshes in San Francisco Bay and the Pacific Northwest; other invasive plants are found along the salt marsh/upland transition and are associated with increased freshwater inputs, especially in southern California. Several conservation lessons and recommendations are apparent from the history of modifications to Pacific Coast salt marshes. Lessons include the value of protecting remaining salt marsh ecosystems, including a diversity of habitats; the importance of maintaining linkages to other natural ecosystems; and the benefit of public support for marsh restoration and preservation. We recommend that future efforts include an inventory of existing Pacific Coast salt marshes, coordinated regional monitoring of Pacific Coast estuaries, largerscale and better-funded restoration projects, and the incorporation of adaptive restoration approaches.

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The coast of our continent’s leading edge is young and dynamic with a narrow continental shelf, rugged terrestrial terrain, and many small watersheds. Slow-moving tectonic plates and volcanic eruptions shape the landscape, sometimes catastrophically. Earthquakes and tsunamis have affected coastal salt marshes directly by shifting elevations and indirectly by changing sediment fluxes (Nelson, Asquith, and Grant 2004). The coastal plains are small in size and sandwiched between steep mountains and deep ocean waters. These features contrast sharply with the more gradually sloping landscapes and large river basins of Gulf of Mexico and Atlantic Coasts. As a result, most Pacific Coast salt marshes are small and isolated, at the distal end of small catchments. Only three major embayments (five hundred square kilometers) contain large areas of salt marsh: San Francisco Bay, the Columbia River, and Puget Sound (National Oceanic and Atmospheric Administration [NOAA] 1990). NOAA (1990) reports that half the wetland area of the Pacific Coast states (2,350 square kilometers) occurs in San Francisco Bay, even though this area is a small fraction (perhaps 5 percent) of its historical extent. Puget Sound is second-largest, with one thousand square kilometers of wetlands (mostly tidal flats), and the Columbia River includes four hundred square kilometers (mostly forested wetlands). The remaining twenty-five estuaries recognized by NOAA (1990) have much smaller drainage areas and less wetland area (see fig. 11 of NOAA 1990). Although some coastal wetlands have been described and measured, a systematic census and mapping effort, like that of the San Francisco Bay Goals Project (Goals Project 1999) is sorely needed for the entire region. NOAA (1990) included twentyeight Pacific Coast estuaries in its summary, but this report lacks detail and misses many smaller ocean inlets that support salt marsh. Seliskar and Gallagher (1983) mapped twenty coastal wetlands in Oregon alone. Washington has six more, not including the Puget Sound area. They report the total area of salt marsh in fourteen of 286

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the Oregon estuaries to be 2,889 to 3,731 hectares, depending on the information source. The largest area is in Coos Bay (581 and 906 hectares, again depending on the information source; Seliskar and Gallagher 1983). Macdonald and Barbour (1974) recognized thirty-one salt marshes for the entire region and listed salt marsh areas of 4,850 hectares for Washington, 2,830 hectares for Oregon, and 22,870 hectares for California. In terms of abiotic factors, annual precipitation along the Pacific Coast varies greatly from north (more than 350 centimeters) to south (25 centimeters in San Diego; San Diego County Water Authority, http://www.sdcwa.org/manage/ rainfall-lindbergh.phtml), with substantial effects on freshwater inputs to local estuaries and salt marshes. Even though precipitation is highly variable regionally, coastal areas along the Pacific Coast have little variation in annual temperature relative to the rest of the United States. The mild climate is a result of the Pacific Ocean, which moderates extremes and provides sea fog, especially in spring and summer. Also unusual is that winter is the wettest season, due to cyclonic depressions. The Mediterranean-type climate of California becomes warmer from northern to southern parts of the state, and only the latter has completely dry summers. In addition to climate effects, hydrological regimes are affected greatly by tidal regimes and coastal wetland size. Some small salt marshes occur adjacent to lagoons that are rarely tidal and have no riverine influence (e.g., McGrath Lake in southern California), while others occur adjacent to rivers within confined valleys (e.g., San Luis Rey River in southern California). Tidal flushing is not dependable in smaller coastal wetlands, because long-shore transport of sand tends to block river mouths. Such wetlands can remain closed for years until a flood event restores the ocean connection (Carpelan 1969) or managers physically open the connection. The larger wetlands are less variable, due to more dependable inflows of freshwater, lack of mouth closure, and the mixing effects of continual tidal influx and efflux.

Other abiotic factors that structure Pacific Coast salt marshes are similar to those affecting marshes worldwide and include elevation (a surrogate for inundation), tidal regime, salinity, substrate, sedimentation, erosion, wave energy, water circulation, and marsh age. However, the relative importance of these factors can be substantially different for Pacific Coast salt marshes. Wave force, for example, restricts salt marshes to riverine fringes (e.g., Salmon River Estuary) and areas inside bay mouth bars that enclose broad river valleys (e.g., Grays Harbor and Willapa Bay, Washington; and Tillamook Bay, Oregon) (Seliskar and Gallagher 1983). Seawater inundation regimes are typically semidiurnal mixed tides, with two unequal high tides per day and two unequal low tides. This mixed, semidiurnal tidal regime exposes the surface of the marsh plain more often than the semidiurnal or diurnal regimes that occur along the Atlantic Coast. In the warmer climates (e.g., in California), prolonged periods of exposure, coupled with long periods without rainfall or riverine influence, create hypersaline soil conditions. Soil salinity levels commonly reach or exceed forty to sixty parts per thousand in southern California marshes, especially during long rain-free periods (typically April through October). Even during the “wet” season, rainfall is confined to a few days, and marsh soils remain hypersaline unless river flooding inundates the marshes for successive days. Tidal amplitudes are up to three meters, and sea storms often raise the elevation of seawater influence by over thirty centimeters (Bromirski, Flick, and Cayan 2003). Pacific Northwest salt marshes are much lower in overall salinity and differ in type in response to varying substrate texture, elevation, and salinity regime (Seliskar and Gallagher 1983). Typical plant species are Salicornia virginica, Triglochin maritima, Jaumea carnosa, Plantago maritima, Distichlis spicata, Spergularia marina, Scirpus americanus, Carex lyngbei, Eleocharis palustris, and Glaux maritima at lower, sloping elevations. Large expanses of Carex lyngbei are found on the slightly higher

and flatter marsh plain, with Deschampsia cespitosa, Juncus balticus, Agrostis alba, Grindelia integrifolia, Potentilla pacifica, and Atriplex triangularis also found in the upper marshes (Seliskar and Gallagher 1983). In southern California, vigorous stands of Spartina foliosa (about one meter tall) are restricted to the channel edges of fully tidal salt marshes, although tall S. foliosa occurs broadly between mean tide and mean high water in San Francisco Bay (Goals Project 1999). Above mean high water, salt marsh plains are dominated by succulent halophytes, principally S. virginica (Macdonald and Barbour 1974; Baye, Faber, and Grewell 2000; Zedler 2001). The marsh plain of San Diego Bay and Tijuana Estuary is comprised of eight to ten halophytes, many of which reach their northern distribution limit at Point Conception (Macdonald and Barbour 1974). Like other salt marsh systems, species distributions vary with elevation and proximity to bay waters, as shown for a San Diego Bay salt marsh (fig. 15.1). The upper marsh grades into upland but retains its saline character and dominance by halophytes in California systems. The upper marsh is the most species-rich community (Traut 2005), likely because of salinity dilution and ability to support a wider variety of life forms. Even in the most saline marshes, relief from salt stress occurs following rainfall, and seeds of annuals and other species readily germinate in the high marsh (Noe and Zedler 2001a, 2001b). Shrubs are also a part of the high marsh community where they can tap less saline groundwater (James and Zedler 2000). Pacific Coast salt marshes are thus more variable in precipitation and salinity, and smaller, more isolated, and more diverse in vegetation than their Atlantic and Gulf Coast counterparts. Despite these differences, the nation’s western salt marshes support the same general ecological functions and societal values as those elsewhere. In addition to recreational, aesthetic, and heritage significance, they provide open space, bird habitat, and expanses of green vegetation for people who live in large coastal cities. Salt marsh

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FIGURE 15.1 Elevational distribution of plants in a San Diego Bay salt marsh. From Zedler et al. 1999.

vegetation stabilizes natural tidal creek banks (Collins, Collins, and Leopold 1987), as well as salt pond levees, flood control channel banks, and shorelines along marinas. That function, however, is challenged by dredging of channels and marinas to accommodate boating. Plant productivity fuels the salt marsh food web, but rates are probably not as great as those of other coasts, due to hypersalinity. However, algal productivity is likely to be more important on the Pacific Coast due to more open plant canopies (Zedler 1980). The lesser area of salt marshes, however, means that the overall contribution of organic matter is lower than on other coasts. Locally, organic 288

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matter production is critical to invertebrates, which in turn feed waterbirds and fish. Many waterbird species, including shorebirds, migrate along the Pacific Coast, refueling where there are broad mudflats and ample invertebrate foods. Fishery support services are highly valued in Pacific Northwest salt marshes, especially for juvenile salmon (Shreffler, Simenstad, and Thom 1990, 1992; Gray et al. 2002). Due to their isolation, small size, and history of human impacts, Pacific Coast salt marshes have high value as habitat for the many animal and plant species that have become rare, threatened, or endangered (table 15.1).

TABLE 15.1 Rare species associated with or dependent on salt marshes of San Francisco Bay Plants Aster lentus Cirsium hydrophilum var. hydrophilum Cordyhlanthus mollis subsp. mollis Lathyrus jepsonii var. jepsonii Lilaeopsis masonii Limosella subulata Fish Eucyclogobius newberryi Gillichthys mirabilis Oncorhynchus tshawytscha Insects Cicindela haemorrhagica Cicindela hirticollis* Cicindela oregona Cicindela senilis senilis Perizoma custodiata Trichocorixa reticulata

MAJOR HUMAN IMPACTS In a coastal region with mostly steep topography, cities developed along rivers and next to the few places that were flat enough to support salt marshes. Therefore, it is not surprising that a larger proportion of natural salt marsh has been lost along the Pacific Coast than in other regions of the United States. In addition to direct habitat modifications, salt marshes have been degraded by indirect impacts that reduce the quality of the remaining habitat. We have chosen to focus on three major factors causing loss of habitat and function: (1) direct habitat modifications, primarily through urbanization; (2) physiographic changes, including hydrologic and geomorphologic impacts, both locally and within the watershed; and (3) invasive plants. Other factors may affect salt marshes but are beyond the scope of our chapter, including climate change and sealevel rise, pollutants, and invasive animals. HABITAT MODIFICATIONS

Mammals Lutra canadensis Microtus californicus Reithrodontomys raviventris Sorex vagrans haliocoetes Sorex ornatus sinuosis Birds Anas acuta Anser albifrons gambelli Charadrius alexandrinus nivosus Laterallus jamaicensis coturniculus Melospiza melodia maxillaries Melospiza melodia pusillula Melospiza melodia samuelis Passerculus sandwichensis Rallus longirostris obsoletus *

Extirpated.

SOURCES: For plants, Baye et al. 2000; Holstein 2000. For fish, Hieb 2000; Maragni 2000. For Insects, Maffei 2000a, 2000b, 2000c. For mammals, Harding 2000; Johnson 2000; Lidicker 2000; Shelhammer 2000a, 2000b. For birds, Alberts and Evens 2000; Becker 2000; Casazza and Miller 2000; Cogswell 2000a, 2000b; Terill 2000; Trulio and Evens 2000.

Historical wetland area loss rates are around 90 percent in southern California (NOAA 1990) where human populations are largest (more than fifteen million people) and salt marshes most scarce. However, impacts are high across the Pacific Coast, with similar rates of wetland loss in San Francisco Bay (Goals Project 1999) and rates nearly as high in the Pacific Northwest. Coos Bay, Oregon, has lost 90 percent of its salt marsh area to agriculture, industry, and residential development, and the entire city of Coos Bay sits on former salt marsh (Seliskar and Gallagher 1983). Since the early 1800s, approximately 60 percent of the marsh area in Puget Sound has been dredged or filled (Seliskar and Gallagher 1983). In Oregon, the largest area of salt marsh loss was caused by diking to create farmland (Seliskar and Gallagher 1983). Diking was one of the few tools available to settlers in search of flat land in a region of steep topography. Unfortunately, the regulation of coastal development and infilling of wetlands has been too little, too late. Filling to raise ground elevations to buildable levels is regulated now, when

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only about 10 percent of coastal wetland area remains in California. Even when an entire salt marsh was not filled, the upper marsh most often was impacted, as the infrequent tidal inundation has given false security to developers who were unaware of the threats to real estate caused by infrequent but violent sea storms and ongoing sea-level rise. The complete extent of these direct losses is unknown. Still visible are the indirect impacts that follow the extension of streets, sidewalks, buildings, utilities, and human presence into coastal salt marshes. Runoff increases with increased “hardscape,” and this brings excess nutrients, sediments, and contaminants directly into salt marshes. In addition, lighting, noise, and airborne materials all disrupt native species’ behavior and health. Although the extent and magnitude of indirect effects are unquantified, it is safe to say that no coastal salt marsh is without some loss or modification. A second impact of urbanization is dredging. Large coastal cities all have ports, shipping channels, and flood control channels that need periodic dredging, which results in the direct loss of salt marsh habitat where dredged areas abut marshes. There is further indirect loss due to boat and ship wakes that destabilize salt marsh banks. Within southern California, Marina del Rey, just north of Los Angeles International Airport, displaced the main salt marsh area of Ballona Wetland. In San Diego Bay, the marina at Chula Vista was excavated around 1980; salt marsh losses were mitigated by moving sediments to a dredge spoil island intended to support the light-footed clapper rail (Rallus longirostris levipes). It is doubtful that this endangered bird ever used the vegetation planted for it, but another endangered bird, the California least tern (Sterna antillarum browni), has nested in the higher, sandier parts of the dredge spoil island (J. Zedler, personal observation). Dredge spoils disposal also can impact habitats, with disposal to over 1,500 hectares in Grays Harbor, 280 hectares in the Columbia River Estuary, and much of the Skagit River Delta in Puget Sound; less than 20 percent of marshes remain in these estuaries (Seliskar and Gallagher 1983). 290

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Third, urban residents demand recreational outlets, and those living next to the ocean make use of water. The nation’s largest aquatic park, San Diego’s Mission Bay, is a former salt marsh. Over eight hundred hectares were dredged and reformed into channels for boating, water skiing, sailing, swimming, and jet skiing and islands for picnicking and other recreational uses. However, beaches are occasionally closed due to high coliform counts. With increased runoff and decreased area of salt marsh to absorb inflowing materials, it is not surprising that Mission Bay has water quality problems. ALTERED PHYSIOGRAPHY

The hydrologic regimes of Pacific Coast salt marshes have been impacted in two primary ways: through modification to tidal connections and through changes in the watershed that delivers freshwater and sediment to the salt marsh. These alterations have occurred for a variety of reasons related to land use and development. In many cases, tidal inlets have been restricted by highway and railroad crossings, adjacent urban or residential development, or flood control channels. In fact, out of the seventeen lagoons and estuaries in San Diego County, it is possible to drive across parts of eleven of them (Marcus 1989). These restrictions are compounded by reductions in the tidal prism of the estuary or lagoon due to upstream filling or sediment inputs. The reduction of the tidal prism creates a feedback mechanism that further reduces the size of the tidal inlet; reduced tidal flows result in reduced scour and increased sediment accumulation in the inlet (Webb 1989). The reduced inlet further restricts flow, eventually leading to the closure of the tidal inlet. In other cases, inlets have been armored with riprap and dredged to maintain permanent tidal openings, leading to increased tidal flows into the estuary. For example, Agua Hediondo Lagoon has a stabilized inlet that maintains the intake of cooling water for a large power plant. Elkhorn Slough has an artificially large tidal opening that is maintained for boating; as a

result, the salt marshes have increased tidal flows, which are a likely cause of salt marsh erosion (Caffrey et al. 2002). Similar to problems with reduction of tidal inlets, enlarged inlets can be sustained by feedbacks once flows are substantially out of balance. The large inlet maintains artificially high tidal flows and reduces opportunities for sediment accumulation. It also changes the salinity regime within the estuary, with implications for both flora and fauna. On the larger scale of the watershed, impacts to salt marshes are related to changes in both freshwater and sediment inputs from the watershed. Along the Pacific Coast, and especially in California, freshwater has been redistributed at the large scale, including interbasin transfers. For example, over half of the annual freshwater inflow to the San Francisco Bay Estuary commonly is diverted (Fox, Mongan, and Miller 1990; Jassby et al. 1995), primarily for urban and agricultural uses, ranging from lawn watering and irrigating plantings along freeways to maintaining crops such as flowers, vegetables, and strawberries. Water importation and irrigation make year-round agriculture and horticulture possible, and runoff greatly exceeds natural inflows, especially in summer. The runoff, its sediment, and unknown contaminants make their way to salt marshes. At Tijuana Estuary, for example, the effect of year-round freshwater inflows is visible at every street drain: nonnative plants intrude into the salt marsh. Without the freshwater subsidy, they would be excluded by hypersaline soil. Many watersheds deliver more sediment to estuaries today due to land uses that increase soil disturbance, such as agriculture (Chambers et al. 2000) and urban development (Greer and Stow 2003). Catastrophic rates of sediment deposition in southern California salt marshes have been recorded following local storm events that suspended this disturbed sediment and moved it into local estuaries and salt marshes. Mugu Lagoon lost 40 percent of its low-tide volume as a consequence of storms in winter 1977–1978, when sediment from agricultural fields filled tidal channels (Onuf 1987). Tijuana

Estuary is highly vulnerable to sedimentation due to its large steep watershed with highly erodible and disturbed hillsides. Furthermore, the ability of the United States to control erosion rates is limited, since most of the watershed is in Mexico. Rainfall often comes in intense storms that saturate the substrate and cause landslides. The problem is exacerbated by the rapidly growing human population just across the border and development that occurs in steep canyons without erosion control measures. Cahoon, Lynch, and Powell (1996) and Ward, Callaway, and Zedler (2003) have documented accumulation of one to eight centimeters of sediment over wet winters in Tijuana Estuary. Callaway and Zedler (2004) found even higher rates (up to thirty centimeters of sediment in one year) in the south arm of the estuary, where sedimentation has resulted in the filling of substantial areas of tidal salt marsh and conversion of formerly tidal areas to elevations above the tides. The Pacific Northwest has similar concerns about sedimentation in coastal watersheds, especially where logging disturbs soil. In other coastal areas, sediment inputs have been reduced by upstream dams, which restrict sediment movement into estuaries. Hydrologic regimes and geomorphologic characteristics drive the development of coastal salt marshes. Shifts in either of these drivers affect the flora and fauna. Mouth closure changes patterns of inundation and stagnation of water, and anoxic conditions can develop in both the water column and soils. If periods of closure are long enough, plants become stressed and die, and so do benthic invertebrates and fish. In southern California, the most obvious indicator of nontidal conditions is monotypic S. virginica. Spartina foliosa and two short-lived succulents, Salicornia bigelovii and Suaeda esteroa, are eliminated by mouth closure in as few as eight months (Zedler, Callaway, and Sullivan 2001). Salicornia virginica, on the other hand, becomes more robust, presumably because it can root deeply and track the declining water table (Zedler, Winfield, and Williams 1980). In addition to lowering dissolved oxygen concentrations,

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nutrients can build up within the lagoon, allowing algal blooms that grow and decay and further reduce dissolved oxygen. Salinity is variable and can increase with evaporation of seawater or decrease with riverine inflows. In either case, prolonged periods of extreme salinity and rapid shifts in salt concentration can negatively affect both animals and plants. Salinity changes and sediment inputs have changed salt marsh vegetation at Los Peñasquitos Lagoon (Greer and Stow 2003). Finally, in the case of marine-dominated salt marshes (e.g., Crissy Field in San Francisco Bay), closure of the tidal inlet can reduce sediment input to the salt marsh leading to slow development of the marsh or decreases in marsh surface elevation. INVASIVE PLANTS

In general, invasive species result in loss of native species, changes in ecosystem functions, and substantial economic impacts. Salt marshes are not particularly vulnerable to invasive plants, because relatively few species are adapted to both high salinity and regular inundation. However, a number of Spartina species have invaded large areas of Pacific Coast salt marshes, and many opportunistic species are problematic in transitional high-marsh habitats. Four Spartina species have been introduced in Pacific Coast salt marshes: S. alterniflora, S. densiflora, S. anglica, and S. patens (Daehler and Strong 1996). Spartina alterniflora and its hybrids with S. foliosa have the most widespread distribution, with hybrid populations in San Francisco Bay and pure S. alterniflora in Willapa Bay (Callaway and Josselyn 1992; Daehler and Strong 1996; Feist and Simenstad 2000). Spartina densiflora has widespread distribution in Humboldt Bay and a smaller population in San Francisco Bay (Spicher and Josselyn 1985; Daehler and Strong 1996; Kittelson and Boyd 1997). Spartina anglica is most widespread in the northern part of Puget Sound, with seventy-three affected sites and over three thousand affected hectares (Hacker et al. 2001). High marsh invasives are abundant in areas where the high marsh and transitional uplands habitats have been subject to substantial distur292

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bance. In southern California, the primary disturbance promoting invasive species establishment and spread is increased input of freshwater to coastal estuaries, leading to a reduction in soil salinity in transitional high marsh locations. Common high marsh invasives include nonnative Mediterranean grasses, such as Polypogon monspeliensis and Parapholis incurva (Kuhn and Zedler 1997; Callaway and Zedler 1998; Noe and Zedler 2001a), as well as native freshwater and brackish marsh species, such as Typha and Scirpus spp. (Zedler and Beare 1986; Beare and Zedler 1987). One of these, P. incurva, interferes with the reproduction of an endangered hemiparasite, Cordylanthus maritimus subsp. maritimus, which attaches its haustoria onto the annual grass to draw moisture and nutrients but is left “high and dry” when the host grass dies before the hemiparasite can produce seed (Fellows and Zedler 2005). Alien species have invaded Pacific Coast salt marshes following deliberate introductions (e.g., S. alterniflora, Avicennia marina, Myoporum laetum, Carpobrotus edulis), hydrologic modifications (e.g., Cotula coronopifolia, Typha domingensis, Rumex crispus), and substrate disruptions (e.g., P. monspeliensis, Bassia hyssopifolia, Salsola tragus, Atriplex semibaccata) (Zedler 1992). Spartina alterniflora was intentionally introduced into San Francisco Bay in the 1970s as an experiment in salt marsh restoration, in part because of earlier difficulties establishing the native S. foliosa. A physiologist intentionally introduced A. marina from New Zealand to Mission Bay, and the population naturalized. When conservationists became concerned that raptors might roost in mangroves and prey on endangered clapper rail chicks, eradication efforts were undertaken. Mature trees were removed, and annual removal of seedlings eliminated the species (note: the closest natural mangroves occur on the tip of the Baja California peninsula). More commonly, invasives have come in accidentally. In the Pacific Northwest, S. alterniflora became established with the oyster industry. Plant fragments were used to pack oysters

and keep them cool during the long railroad travel from the east coast. The stems of S. alterniflora were dumped into Willapa Bay along with the oysters and eventually became widely established (Feist and Simenstad 2000). Similarly, S. densiflora was likely established in Humboldt Bay accidentally through shipping associated with the lumber industry and shipments to and from South America (Spicher and Josselyn 1985). Many exotic species also have intruded opportunistically into high marsh habitats. At Carpinteria Marsh (near Santa Barbara) and Sweetwater Marsh (in San Diego Bay), Myoporum laetum forms a conspicuous band of trees adjacent to the north-south railroad. It was first introduced for horticultural reasons and is probably not the only invasive to be transported by rail traffic. However invasive species become established in salt marshes, the outcome is a degraded ecosystem. Invasives compete with native species and reduce their distribution and abundance. Within San Francisco Bay, S. alterniflora hybridizes with S. foliosa, compounding impacts on the native (see chaps. 1 and 2, this volume). Salt marsh habitat has been extended onto mudflats by S. alterniflora and its hybrids, because these plants can grow lower in elevation than S. foliosa in San Francisco Bay and below low marsh natives in the Pacific Northwest, such as Carex lyngbyei (Callaway and Josselyn 1992; Feist and Simenstad 2000). Lost mudflat reduces available foraging areas for shorebirds during many low-tide events. Finally, some invasive plants cause economic impacts. Because of its ability to grow at lower elevations, S. alterniflora has negatively affected the oyster industry within Willapa Bay, and this has led to pressure to control its spread.

MANAGEMENT EFFORTS As illustrated, the Pacific Coast “conservation story” has been one of massive salt marsh loss and degradation, as the human population expanded and plant and animal populations

declined. Historically, nature and humans competed for scarce flat lands along the coast; today, nature is threatened, and people are finding ways to compensate for past losses. Although it seems obvious that the 10 to 20 percent of remaining salt marshes should be inviolate, developers continue to argue that small increments can still be filled or dredged without impact and that any damages to salt marsh can be compensated by creating or restoring habitat. Substantial efforts have been undertaken to deal with the three impacts to Pacific Coast salt marshes that we have highlighted (habitat loss, physiographic alterations, and invasive species). To compensate for habitat loss due to urbanization and agriculture, restoration efforts have been undertaken within the mitigation-permitting process. To offset hydrological impacts, managers have used a variety of approaches; and to control or eradicate invasives, public agencies have used chemical methods, manual pulling, and other efforts. Each of these responses is evaluated here.

HABITAT LOSS AND RESTORATION

Some loss of salt marsh habitat has been offset by restoration, either as mitigation for new impacts or by agency or nonprofit efforts to expand habitat. Early attempts at salt marsh restoration focused on establishing appropriate elevations for plant recruitment, assuming this would be adequate to establish a functioning salt marsh. Over time, it has become apparent that restoration is not so simple, especially in an urban context. One challenge is that salt marshes appear extremely flat, but small-scale heterogeneity likely contributes to overall biodiversity by creating variations in drainage, soil, and other conditions. Elevation shifts of ten centimeters or less are sufficient to shift plant composition (fig. 15.1; Zedler et al. 1999; Sanderson, Ustin, and Foin 2000). Similarly, invertebrates and fish are also likely affected by small-scale heterogeneity. Tidal creeks drive heterogeneity in salt marshes, and their formation within restoration projects has been an ongoing

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BOX 15.1: CASE STUDY OF ADAPTIVE RESTORATION AT TIJUANA ESTUARY In the mid-1980s, researchers and managers agreed that the key to sustaining rare and endangered species at Tijuana Estuary (a National Estuarine Research Reserve [NERR]) would be to restore salt marsh in the southern arm where repeated sedimentation events had elevated over two hundred hectares of wetland. Only the northern arm continued to support tidally dependent species that were endangered (lightfooted clapper rail [R. l. levipes], salt marsh bird’s-beak [C. m. subsp. maritimus]) or threatened (annual pickleweed [S. bigelovii], sea-blite [S. esteroa]). Threats increased when tidal flushing ceased in 1984, due to sluggish tidal action and sand accretion in the estuary mouth. Before removing sediments, Williams and Swanson (1987) analyzed hydrological conditions, devised a series of excavations, and incorporated Pacific Estnacine Research Laboratory (PERL) researchers’ request to phase the project so early results could inform later modules. The effort involved all members of the NERR Management Authority (U.S. NOAA NERR System, U.S. Fish and Wildlife Service, U.S. Border Patrol, State Coastal Conservancy, State Parks, County Parks, City of San Diego, and City of Imperial Beach). They agreed to implement the first module near the NERR Visitor Center, where it could demonstrate the value of increasing intertidal habitat. In 1997, about 0.5 hectare of disturbed upland was converted to intertidal habitat. PERL researchers designed the site to support research on biodiversity (would speciesrich plots accumulate more biomass and trap more nitrogen?) and restoration (which species need to be planted, and which would recruit on their own?). The site design and experiment were coordinated—the intertidal bench was graded flat and smooth to accommodate replicate plots, and eighty-seven experimental plots were sized and spaced to vegetate the marsh plain. In March 1997, PERL planted about six thousand seedlings within eighty-seven 2 ⫻ 2–meter plots. Treatments compared

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plantings of zero, one, three, and six species to a plot (at ninety plants per plot). Tidal flushing was reintroduced in April 1997, and researchers assessed canopy development, biomass and nitrogen accumulation, and halophyte recruitment, with funding from the National Science Foundation (NSF). As predicted by biodiversity-ecosystem-function theory, species-rich plantings increased plant growth and nitrogen accumulation (Callaway et al. 2003). In addition, canopies were more complex (Keer and Zedler 2002). Only one species (S. virginica) was able to recruit and dominate without introduction; the remaining seven halophytes needed to be planted or sown as seeds (Lindig-Cisneros and Zedler 2002). The information that S. virginica did not need planting saved considerable money for phase 2. The phase 1 site became fully vegetated within two years, and a clapper rail began feeding along the excavated tidal channel. This highly visible project enhanced support for additional excavation, this time in the south arm, where more land (up to two hundred hectares) was available for restoration. Experimental designs and restoration were planned in coordination for a larger site (eight hectares) where PERL proposed to determine how the addition of tidal creeks would affect fish use and salt marsh development using three replicate areas with tidal creek networks and three without creeks. Additional experiments tested spacing of halophyte transplants and the need to condition the soil. Researchers asked several questions: How does topographic heterogeneity influence ecosystem structure and function? Do creeks need to be added or will they form on their own? Do soils need to be amended and plants tightly spaced to enhance survival? During January 2000, contractors introduced S. foliosa to the “cordgrass zone,” within large plots that did and did not have added soil conditioner (50 percent kelp waste product and 50 percent perlite), and PERL planted thousands of seedlings in plots amended with kelp compost (rototilled in), plots simply rototilled, and unmodified plots. Half were near tidal creeks, and half were thirteen meters away. The same plot spacing

was repeated in approximately one-hectare areas (cells) that lacked a tidal creek network. On February 14, 2000, the eight-hectare marsh plain was opened to tidal flow. While cordgrass survived and expanded, the marsh plain experienced severe stress from drying and hypersalinity followed by flooding and sedimentation (Zedler et al. 2003; Wallace, Callaway, and Zedler 2005). Nearly all the marsh plain seedlings died, so a new experiment was designed with five species known to need replanting and a test of spacing (ten, thirty, or ninety centimeters from the central transplant). By December 2000, all 108 plots were established, and NSF support came through to assess outcomes. Counter to expectations, the soil conditioner had a greater effect on survival than tight clusters or tidal creeks (O’Brien and Zedler 2006). Subsequent research showed that tidal creeks were conduits for fish that feed on the marsh plain (Larkin et al. 2008; Larkin, West, and Zedler In Press). Most important, the creeks kept the marsh from completely filling with sediment from multiple floods (Wallace et al. 2005). In addition, tidal pools formed and served as “oases” for algae, invertebrates, and fish (Larkin et al. 2008; Larkin, West, and Zedler In Press). Overall, topographic heterogeneity enhanced the development of ecosystem function. Phase 3 is being planned, and the findings so far lead to several cautions and recommendations. First, size can be a constraint on salt marsh restoration. Simultaneously exposing eight hectares of former salt marsh soil allows salt crusts to form as tidal waters move over the

dark surface, heat up, and evaporate. Second, seedlings have low survival on hypersaline flats. Under natural conditions, vegetatively reproducing halophytes would creep slowly onto exposed mudflats, or seedlings would germinate in the shade of their ancestors. Third, tidal creeks accelerate ecosystem development, but they might form on their own if the site is excavated to mudflat level. Fourth, if small islands of marsh plain elevation are left within the mudflat, the marsh plain area that needs to be planted would be small, and planting could be very dense, which would allow rapid canopy closure and prevent salt crust formation. Finally, given plenty of inflowing sediment, the islands should expand rapidly, as plants trap and stabilize substrate. All of these recommendations can be incorporated into an experimental design for phase 3. Adaptive restoration allowed tests of ecological theory, answered restoration questions, and restored over eight hectares of salt marsh. Key attributes were close communication among researchers and managers, the U.S. Fish and Wildlife Service and State Coastal Conservancy’s funding for planning and excavation, and NSF’s support of research. The difficulties were in timing—having the site, the plan, and funding. Still, we know of no better way to improve the practice of restoration than by capitalizing on the availability of actual sites and using them to test alternative approaches as replicated treatments. At least one endangered species seems to agree—light-footed clapper rails have been sighted within the cordgrass zone!

challenge. If restored sites are graded to marsh plain elevations, tidal energy is often insufficient to create creeks (Cornu and Sadro 2002; Williams and Orr 2002). Creeks will develop at lower elevations, but lowering the topography is costly and increases the time for vegetation establishment. Hence, managers need to know if tidal creeks aid ecosystem development. To find out, we initiated an experiment with and without creeks at an eight-hectare restoration site in Tijuana Estuary (see box 15.1).

A related challenge for a number of Pacific Coast salt marsh restoration efforts is dealing with areas that have subsided to low intertidal or subtidal elevations. These elevations no longer support vegetated marshes, often due the compaction and oxidation of organic soils following diking. For example, Salmon River Estuary (Frenkel and Morlan 1991) historically supported a large salt marsh, but when the dike was breached to restore tidal flushing, the water was too deep for all the salt marsh to recover. Muzzi

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Marsh in San Francisco also was slow to establish vegetation because elevations were low; eventually enough accretion occurred to support vigorous stands of S. foliosa and California clapper rail (Rallus longirostris obsoletus) (Williams and Faber 2001). There are many questions and trade-offs in dealing with such sites. As noted, they can be left at low elevations to allow sediments to accumulate and creeks to form, but vegetation will recruit slowly. Alternatively, sites can be restored to higher elevations by filling. However, this impedes creek formation, and the material that is used to fill the site (often dredged material) might have inappropriate substrate conditions (e.g., texture and organic content), or it might be contaminated. A third finding is that these substrate conditions determine the degree to which an appropriate elevation can support desired biota. At San Diego Bay, Caltrans was required to provide nesting habitat for the light-footed clapper rail; however, sandy dredge spoils did not supply enough nitrogen for S. foliosa to grow tall and support nests, even though the elevation was correct and the plant was present (Langis, Zalejko, and Zedler 1991; Zedler 1998). Thus, we learned that sandy substrate is unsuitable for restoration of functional salt marsh (Zedler and Callaway 1999). At Tijuana Estuary, the excavation of sediments to expose the historical marsh plain

FIGURE 15.2 Salt crusts which reduced vegetation establishment in the south arm of Tijuana Estuary. Photo by J. Zedler.

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rapidly formed a salt crust (fig. 15.2), and only a tiny minority of plantings survived (Zedler, Morzaria-Luna, and Ward et al. 2003). Beyond these ecological challenges, several management issues complicate restoration, because different agencies have different perspectives. Management aims are sometimes about restoration but often about habitat remodeling. Targets often depend on who makes the choice and what land uses are grandfathered in. In San Diego County, the California Department of Fish and Game manages Buena Vista Lagoon as a fresh-to-brackish water body for waterfowl, even though this wetland would naturally range from brackish to tidal (with an open river mouth) to hypersaline (with impounded seawater). Batiquitos Lagoon was recently dredged to convert what was sometimes shallow water and sometimes a salt flat (with a salt marsh fringe) into a permanently flooded deepwater basin to mitigate damages to fish habitat caused by port expansion in Los Angeles Harbor. NOAA argued strongly for in-kind compensation, which had to occur off site (in this case, ninety kilometers south) because there were no closer options. In some cases, political pressures trump sensible management. In 2005, the U.S. Congress passed a homeland security law (HR 418) that includes $53 million to build a triple border fence along the United States–Mexico border,

purportedly to protect the U.S. Navy facilities in San Diego Bay from terrorists presumed to crawl under or swim around the existing solid steel fence. This project will scrape bluff tops and fill gullies in order to add two more fences and create a no-man’s land the width of a major freeway for Border Patrol personnel to traverse with ease. The project would destabilize enough material that, if mobilized by heavy rainfall or earthquakes, could cover the adjacent wetland in Tijuana Estuary a meter deep. The area between fences would be kept clear of vegetation, ensuring erosion during storms. Despite environmentalists’ pleas to use existing science on sedimentation rates and impacts and to consider alternative, more effective security measures, the fence is expected to go forward, as the law exempts the project from environmental review and prevents legal action. Pacific Coast salt marsh restoration is moving toward large-scale, regional restoration efforts, based in part on recommendations to address projects on a broader scale, rather than a case-by-case basis (Zedler 1996). In addition, restoration managers have realized that projects cannot be designed with only a narrow scientific perspective but that social issues and multiple stakeholders have to be incorporated. For example, the Southern California Wetlands Recovery Project (http://www.scwrp.org) includes over a dozen agencies and works with community task forces to identify restoration priorities in southern California. The San Francisco Bay Habitat Goals project involved more than fifty scientists from agencies, academia, and nongovernmental organizations who identified goals for Bay-wide restoration (Goals Project 1999). These efforts helped set the stage for the largest restoration project on the West Coast, the South Bay Salt Pond Restoration Project (http://www. southbayrestoration.org). The CALFED Bay– Delta Program has the ambitious goal of addressing both water management issues for the state, as well as restoring wetlands in the Bay–Delta region with a large number of collaborating organizations and agencies (http:// calwater.ca.gov). Similar regional efforts are

underway in the Seattle area with the Puget Sound Nearshore Ecosystem Restoration Project (http://www.pugetsoundnearshore.org), a partnership between federal and state agencies, Indian tribes, industries, and environmental organizations. Within the Pacific Northwest, interest in watershed management has been substantial, focusing on the impacts of timber harvesting and dams on migratory salmon (Williams et al. 1997). There is obviously a clear connection to coastal wetlands, and many efforts are underway to restore salmon habitat (e.g., Simenstad and Thom 1996). Another positive trend has been the development of handbooks and guidelines for restoration efforts. Early books gave advice on restoration planning in California, including those by Josselyn (1982), Josselyn and Buchholz (1984), and Zedler (1984). More recently handbooks have been developed for both southern California (Zedler 2001) and San Francisco Bay (Phillip Williams and Associates and Faber 2004). A third trend is toward conducting the restoration as large field experiments (Zedler 2001; Cornu and Sadro 2002). Elsewhere, we have called this approach “adaptive restoration” (Zedler and Callaway 2003; Zedler 2005). The essential steps of the process are (1) identify uncertainties that need to be better understood in order to restore the site; (2) design the large-scale restoration program as a series of phased modules; (3) design and implement each module as a field experiment that tests alternative approaches to achieve specific objectives and reduce key uncertainties; (4) use the knowledge gained from early experiments to improve later modules; and (5) ask new questions of each subsequent module, developing the answers through experimentation. A case study of this approach from Tijuana Estuary is provided in box 15.1. ALTERED PHYSIOGRAPHY

The management focus for tidal inlets of Pacific Coast salt marshes has been to maintain permanent tidal connections, despite the fact that some lagoons may have had naturally intermittent tidal connections (Carpelan 1969). The

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specific benefits and natural occurrence of intermittently tidal systems has not been fully evaluated; however, more permanent connections ensure regular flushing of lagoons and salt marshes and are likely to lead to more stable water quality conditions and higher local biodiversity. Management at Los Peñasquitos Lagoon in San Diego County has focused on maintaining the lagoon’s tidal opening. The condition of the tidal inlet, as well as the lagoon’s water level and water quality, has been monitored continually for State Parks by the Pacific Estuarine Research Laboratory (PERL, San Diego State University) for over a decade. When the mouth of the lagoon closes, water quality data loggers are watched closely for extreme salinity and low dissolved oxygen conditions. Threshold conditions have been established over the last decade, and if these are reached, permits are in place to allow an emergency crew to bulldoze and reopen the mouth of the lagoon. The typical cycle is for the mouth to close in late summer, during periods with naturally low tidal range and little or no river flow. Sediment accumulates, nearshore sand transport builds a berm at the inlet, and the bridge at the lagoon mouth restricts inlet migration. Without intervention, the mouth would remain closed until winter rainfall collects in the lagoon and overtops the berm or until the highest tides in January overtop the berm from the ocean side. The mouth is typically opened mechanically when the tide range is large to ensure maximum flushing and tidal scour of the inlet. A similar approach is used at Crissy Field in San Francisco Bay where long-shore drift of sand within the bay and the small tidal prism of the restored marsh has led to frequent closure of this system (K. Ward, personal communication). As indicated earlier, Elkhorn Slough has experienced the opposite problem, excessive tidal flow, due to hydrological modifications in the 1940s designed to maintain a marina entrance (Caffrey et al. 2002). Historical changes in the Slough have been documented (Caffrey et al. 2002; Van Dyke and Wasson 2005), and the increase in tidal flows throughout the slough 298

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are hypothesized to be the cause of salt marsh loss. Recently, the Elkhorn Slough National Estuarine Research Reserve (NERR) began a management planning process to address this issue, with multiple levels of public participation and scientific review, involving local, state, and federal agencies, academics, and diverse stakeholders, including birdwatchers, the marina district, kayakers, and agricultural interests in the watershed. The NERR staff and advisers are now evaluating a range of alternatives for managing hydrological conditions, sediment inputs, vegetation, and other parameters in an attempt to reduce salt marsh loss (B. Peichel and B. Largay, personal communication). Because upstream impacts drive many of the changes in coastal salt marshes, managers of coastal wetlands realize that a large-scale, watershed-based approach is necessary to manage hydrological issues for coastal salt marshes. However, watershed management is difficult because the scale is large and stakeholders are many. Despite these challenges, managers are realizing that improving upstream management can prevent or reduce problems before they reach the bottom of the watershed (Williams, Wood, and Dombeck 1997). State and local agencies are educating the public about the importance of improving the management of coastal watersheds (e.g., see http:// www.projectcleanwater.org), and interest in managing water and pollutants on a watershed basis also has been promoted widely by federal agencies, including the Environmental Protection Agency (e.g., the Total Maximum Daily Loads program). For southern California’s Los Peñasquitos Lagoon, the State Coastal Conservancy is trying to reduce excessive freshwater and sediment input from the disturbed watershed (J. King and K. Bane, personal communication). An additional challenge for Tijuana Estuary is that 75 percent of its watershed is in Mexico. Binational efforts are underway to improve management and education, including the publication of the Tijuana River Watershed Atlas (Wright and Vela 2005). In 2004, the United States built an $8 million

FIGURE 15.3 Aerial image of restoration in the south arm of Tijuana Estuary, including a sediment basin that was designed to protect the project from watershed delivery of sediments. From Zedler and West 2007.

sediment retention system adjacent to the border to protect an existing eight-hectare restoration project and future restoration efforts (fig. 15.3). Restoration measures in the estuary began with excavation of one to three meters of historically deposited sediments, which have elevated the marsh plain above tidal influence. Even with these efforts, the site has continued to experience record sedimentation rates, especially during winter of 2004–2005 (average deposition of 4.5 centimeters; Wallace et al. 2005). INVASIVE PLANTS

In Washington, large-scale efforts are underway to control Spartina spp., using a variety of methods. The most widely used herbicide has been glyphosate, although initial tests of a new herbicide, imazapyr, suggest that it is more effective (Murphy 2004). Other control methods are crushing, digging, and covering the plants and introducing Prokelisia marginata, a planthopper, as a biocontrol agent (Grevstad et al. 2003; Murphy 2004). Hacker et al. (2001) found an overall reduction of 13 percent for S. anglica in Puget Sound with glyphosate treatment from

1997 to 2000, and the effectiveness of the control varied due to habitat conditions, with the most effective control in high-salinity marshes. Annual control was necessary to cause a decrease in S. anglica distribution, and S. anglica actually expanded its distribution with intermittent treatment (Reeder and Hacker 2004). Within Washington, the use of herbicides for control has been controversial, with ongoing disputes between strong advocates of control methods and others who are concerned about potential negative effects of herbicide use. The recent use of imazapyr has been contested in court, as was earlier use of glyphosate. So far agencies have been allowed to continue using the herbicides, although lawsuits have slowed control efforts. Grevstad et al. (2003) evaluated the use of P. marginata (a planthopper) as biological control, and it reduced S. alterniflora biomass by 50 percent and height by 15 percent in field cages in Willapa Bay. The long-term effectiveness of this method is uncertain due to questions concerning the planthopper’s dispersal and overwintering capabilities in Washington. Within Humboldt Bay, the local community is not interested in herbicide

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use, and control efforts for S. densiflora have focused on mowing, with hand pulling and digging of isolated plants (A. Pickart, personal communication). Control of S. alterniflora within San Francisco Bay is complicated by two critical differences with the situation in Washington. First, there is a native Spartina in San Francisco Bay, S. foliosa, with substantial hybridization between these species (Daehler and Strong 1997). The hybrids exist across a gradient of morphological characters, and it is difficult to distinguish between the native and the hybrid in the field. The only completely reliable way to identify hybrids is to complete genetic testing on individual tissue samples. It may only be possible to control S. alterniflora and the hybrid, by eliminating all Spartina in an area, including S. foliosa, with the potential for large-scale negative consequences for the genetic diversity of S. foliosa populations and the biota that use this habitat. The second complication related to Spartina control in San Francisco Bay is the presence of the California clapper rail, a federally listed endangered species that uses Spartina habitat for foraging and nesting. It appears that the California clapper rail uses patches of both the nonnative and hybrid species, and its numbers may have increased in the short term due to the spread of the invasives. However, there is concern that geomorphic changes associated with the invasives (especially infilling of low marsh creeks) may lead to the long-term loss of habitat for clapper rails (J. Collins, personal communication). Any control efforts that substantially reduce Spartina habitat in the long term will be complicated by the listed species. Even with these challenges, a multimillion-dollar effort to control S. alterniflora in San Francisco Bay is underway with hopes for control over the next five years (see http://www.spartina.org for more details). In contrast to the large-scale efforts to control Spartina species in Washington and San Francisco Bay, little has been done to control transitional high marsh invasives, although attempts to control tamarisk in southern California have been initiated (L. Levin and 300

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D. Talley, personal communication). Given that expansion of most high marsh invasives is linked to increased freshwater inputs into local estuaries, it is unlikely that they could be controlled without some effort to address these inputs. However, a management shift of this sort would require large-scale changes in water use that at present are unlikely to occur. Each case of an invasive species is unique because each is a result of multiple factors, including the characteristics of the invading species, the native community, and the anthropogenic disturbances that may be allowing a particular species to expand its distribution. To control a species, managers need to understand each of these factors and consider stakeholders and local constraints. If a decision is made to control a species, a quick response is much more likely to be successful; however, this is an enormous challenge in typical regulatory settings. Approval for herbicide use against Spartina spp. has taken many years in both Washington and San Francisco Bay.

LESSONS LEARNED AND RECOMMENDATIONS The Pacific Coast is diverse and dynamic. Its salt marshes are few and far between. They occur within lagoons and estuaries of uncertain number and area, due to incomplete census information and variable definitions of coastal water bodies. Total remaining salt marsh area is unknown, but the remnants are 10 to 20 percent of that estimated in approximately 1800. There are a number of challenges that remain for the conservation and restoration of Pacific Coast salt marshes, and we summarize these in the following lessons and recommendations. Lesson 1. The loss of 80 to 90 percent of Pacific Coast salt marsh area is unacceptable for a region that never had “enough” coastal salt marsh. The remaining 10 to 20 percent is not enough to provide basic ecosystem services or ensure the survival of sensitive, wetlanddependent species. Extinction is forever, and

extirpation has nearly the same effect in Pacific Coast wetlands. In this region with its small, discontinuous estuaries and little remaining habitat, an extirpation at one site might not be reversible without human assistance. Lesson 2. Preservation is much easier (and probably cheaper) than restoring salt marsh to compensate for damages. At Tijuana Estuary, plans to bury a four-meter diameter sewer pipe near the surface of Tijuana Estuary were rejected when it became clear that damages would be extensive and mitigation difficult to impossible. Instead, a fifty-meter shaft was excavated near the edge of the estuary, and the pipe was placed in a horizontal tunnel well beneath the estuary. Although construction was expensive, this solution avoided many other costs, including time and resources that would have been spent on detailed environmental assessment, contention over actual damages, mitigation of damages, lengthy monitoring of compensatory efforts, and the potential for litigation. Lesson 3. The severe loss of habitat area and increasing threats of extinction are arguments for “keeping all the parts” (sensu Aldo Leopold) somewhere in the region. This is most readily accomplished by restoration and management of a variety of habitats in a fully tidal wetland, even if they may not have occurred naturally. Some goals, such as waterfowl and shorebird migratory stopovers, are well managed for by retaining unnatural features, such as salt ponds (Goals Project 1999). Lesson 4. Although salt marshes of the Pacific Coast are somewhat linked by water, through ocean circulation and tidal fluxes, and by air, they are not well linked on land. Species that are extirpated can only reestablish if their propagules can arrive by water, by air (e.g., wind or bird dispersal), or via land corridors (animal dispersal). Today, it is roadways and automobiles that link salt marsh remnants, and their effects are overwhelmingly negative. Several species have neared extirpation in Pacific Coast salt marshes. In Tijuana Estuary,

populations of both S. bigelovii and S. esteroa are still rare following an eight-month inlet closure, and reintroductions were unsuccessful (Zedler et al. 2001). Populations of the endangered hemiparasite, C. m. subsp. maritimus, have been highly variable since reintroduction in San Diego Bay (Zedler 1998). The same species declined at Tijuana Estuary (the donor site for seeds) during very dry years. If these two populations are extirpated, the nearest source population would be Upper Newport Bay, which is over 150 kilometers north. Seed banks need to be created and sustained, and species reintroduced as needed to keep all the species in the region. Lesson 5. People care about the remaining salt marshes. Residents of Los Angeles continue a decades-long battle to protect the last 120 hectares of Ballona Wetland (near Los Angeles International Airport) from impacts of upstream development. Their battle to acquire and protect Ballona Wetland is testimony to the value that citizens of Los Angeles assign to the region’s remaining greenspaces for aesthetic enjoyment and passive recreation, as well as an educational resource. A persistent group of San Diego citizens succeeded in getting the city spend over $3 million to purchase a sevenhectare lagoon and salt marsh called Famosa Slough (a tributary to Mission Bay). Once local residents got the site set aside, they immediately began restoring it by removing trash, planting native vegetation, and helping the city acquire grant funds to impound and improve urban runoff at the upstream end. Famosa Slough has earned its “fame” (fig. 15.4). Recommendation 1. A detailed inventory of Pacific Coast wetlands and their salt marsh habitats is needed. The EcoAtlas project of the San Francisco Estuary Institute (http://www.sfei.org/ecoatlas/index.html) is a model; similar efforts are critically needed to document salt marsh remnants and opportunities for salt marsh restoration. Recommendation 2. Comprehensive and coordinated monitoring is needed for Pacific Coast

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(A)

(C)

estuaries. One particular need is to track spatial and temporal patterns of rare and threatened species along the entire Pacific Coast (table 15.1). The shorebird monitoring project of the International Shorebird Surveys (ISS) is an excellent example. A volunteer survey initiative was launched in 1974 by Manomet Bird Observatory (now Manomet Center for Conservation Sciences, or MCCS; http://www.manomet.org/WHSRN/iss.htm). Thousands of observers track bird use and collect data across the globe. Coordinated monitoring for invasive species is also needed with rapid response to any newly detected invasive species in order to effectively control or eradicate

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(B)

FIGURE 15.4 Famosa Slough, in San Diego, California. This project was restored almost entirely by citizens’ efforts, coordinated by the Friends of Famosa Slough. Photos by J. Zedler.

nonnative species. Such an effort is underway for Spartina spp. through the various programs along the Pacific Coast; however, similar efforts are needed for other salt marsh invasive species. Recommendation 3. The move toward largerscale restoration with a wide range of collaborating stakeholders is positive, but restoration efforts need to be accelerated. The limiting factor for restoration is not potential sites but funding. For example, Tijuana Estuary has two hundred hectares available for restoration, a general plan has been developed, it is a research reserve where outcomes can be studied and tracked over the long term, and the efforts would likely help threatened and endangered

species. However, funding for restoration within the estuary is lacking. Recommendation 4. Adaptive approaches are needed. Adaptive restoration begins by stating what is not known that needs to be known to restore a site. Next, the site is subdivided into modules to be restored in phases, and the most urgent questions are matched to the first module. An early question might be, which species need to be planted and in what assemblages? A whole-module experiment is then designed to restore the first module while answering the first question(s). Experimental plots fill the module, so that the entire area is restored, but the different approaches suggest what will work best for subsequent modules. The second module applies the best approach from the first module, within the context of a new experiment that seeks answers to the next most pressing question. For example, how should the soil be amended or the topography manipulated to achieve project goals? As the process continues, a body of knowledge develops, and the area becomes restored to increasing functionality. Subsequent projects in the region can take advantage of these findings. Managers can compare treatments and see the consequences of each alternative restoration procedure. The risk of using the wrong approach over the entire site is avoided. The advantages over the trial-and-error approach greatly outweigh the cost of monitoring and analysis, because whole-site errors are avoided. Acknowledgments. We thank Mark Bertness and Brian Silliman for inviting this contribution and Lisa Levin and two anonymous reviewers for their comments. Research on Tijuana Estuary and writing were supported in part by a grant from the National Science Foundation (DEB0212005) to Zedler and Callaway and in part by an award from Earth Island Institute. REFERENCES Albertson, J. D., and J. G. Evens. 2000. California clapper rail. Pages 332–341 in P. R. Olofson (ed.), Bayland Ecosystem Species and Community Profiles:

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Plants, Fish, and Wildlife. Oakland, CA: San Francisco Bay Regional Water Quality Control Board. Traut, B. H. 2005. Coastal ecotones: A case study of the salt marsh/upland transition zone in California. Journal of Ecology 93: 279–290. Trulio, L. A., and J. G. Evens. 2000. California black rail. Pages 341–345 in P. R. Olofson (ed.), Bayland Ecosystem Species and Community Profiles: Life Histories and Environmental Requirements of Key Plants, Fish, and Wildlife. Oakland, CA: San Francisco Bay Regional Water Quality Control Board. Van Dyke, E., and K. Wasson. 2005. Historical ecology of a central California estuary: 150 years of habitat change. Estuaries 28: 173–189. Wallace, K. J., J. C. Callaway, and J. B. Zedler. 2005. Evolution of tidal creek networks in a high sedimentation environment: A 5-year experiment at Tijuana Estuary, California. Estuaries 28: 795–811. Ward, K. M., J. C. Callaway, and J. B. Zedler. 2003. Episodic colonization of an intertidal mudflat by native cordgrass (Spartina foliosa) at Tijuana Estuary. Estuaries 26: 116–130. Webb, C. K. 1989. Inlet dynamics for southern California. Unpublished MS thesis, San Diego State University, San Diego, CA. Williams, J. E., C. A. Wood, and M. P. Dombeck (eds.). 1997. Watershed Restoration: Principles and Practices. Bethesda, MD: American Fisheries Society. Williams, P. B., and P. M. Faber. 2001. Salt marsh restoration experience in the San Francisco Bay Estuary. Journal of Coastal Research Special Issue 27: 203–211. Williams, P. B., and M. K. Orr. 2002. Physical evolution of restored breached levee salt marshes in the San Francisco Bay estuary. Restoration Ecology 10: 527–542. Williams, P. B., and M. L. Swanson. 1987. Tijuana Estuary enhancement hydrologic analysis. Oakland, CA: California State Coastal Conservancy. Wright, R. D., and R. Vela, eds. 2005. Tijuana River Watershed Atlas. San Diego, CA: San Diego State University Press, Institute for Regional Studies of the Californias. Zedler, J. B. 1980. Algal mat productivity: Comparisons in a salt marsh. Estuaries 3: 122–131. ———. 1984. Salt Marsh Restoration: A Guidebook for Southern California. La Jolla: California Sea Grant College Program, Institute of Marine Resources, University of California. ———. 1992. Invasive exotic plants: Threats to coastal ecosystems. Pages 49–62 in P. M. Grifman and S. E. Yoder (eds.), Perspectives on the

Marine Environment. Los Angeles: University of Southern California Sea Grant Program. ———. 1996. Coastal mitigation in southern California: The need for a regional restoration strategy. Ecological Applications 6: 84–93. ———. 1998. Replacing endangered species habitat: The acid test of wetland ecology. Pages 364–379 in P. L. Fiedler and P. M. Kareiva (eds.), Conservation Biology for the Coming Age. New York: Chapman & Hall. ———, ed. 2001. Handbook for Restoring Tidal Wetlands. Boca Raton, FL: CRC. ———. 2005. Restoring wetland plant diversity: A comparison of existing and adaptive approaches. Wetlands Ecology and Management 13: 5–14. Zedler, J. B., and P. A. Beare. 1986. Temporal variability of salt marsh vegetation: The role of lowsalinity gaps and environmental stress. Pages 295–306 in D. A. Wolfe (ed.), Estuarine Variability. San Diego, CA: Academic Press. Zedler, J. B., and J. C. Callaway. 1999. Tracking wetland restoration: Do mitigation sites follow desired trajectories. Restoration Ecology 7: 69–73. ———. 2003. Adaptive restoration: A strategic approach for integrating research into restoration

projects. Pages 167–174 in D. J. Rapport, W. L. Lasley, D. E. Rolston, N. O. Nielsen, C. O. Qualset, and A. B. Damania (eds.), Managing for Healthy Ecosystems. Boca Raton, FL: Lewis. Zedler, J. B., J. C. Callaway, J. S. Desmond, G. VivianSmith, G. D. Williams, G. Sullivan, A. E. Brewster, and B. K. Bradshaw. 1999. Californian salt marsh vegetation: An improved model of spatial pattern. Ecosystems 2: 19–35. Zedler, J. B., J. C. Callaway, and G. Sullivan. 2001. Declining biodiversity: Why species matter and how their functions might be restored in Californian tidal marshes. BioScience 51: 1005–1017. Zedler, J. B., H. Morzaria-Luna, and K. Ward. 2003. The challenge of restoring vegetation on tidal, hypersaline substrates. Plant and Soil 253: 259–273. Zedler, J. B., and J. M. West. 2007. Declining diversity in natural and restored salt marshes: A 30year study of Tijuana Estuary. Restoration Ecology 16: 249–262. Zedler, J. B., T. Winfield, and P. Williams. 1980. Salt marsh productivity with natural and altered tidal circulation. Oecologia 44: 236–240.

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PART V

International Perspectives

16

Human Modification of European Salt Marshes A. J. Davy, J. P. Bakker, and M. E. Figueroa European coastal salt marshes range from the Arctic to the Mediterranean. Tidal amplitude may be low, as in the brackish marshes of the Baltic and the frequently hypersaline marshes of the Mediterranean. Marshes on the extensive coasts and estuaries of the North Sea, Irish Sea, and Atlantic Ocean tend to be meso- or macrotidal. There are also inland salines, particularly in steppes of central Europe. The diversity of communities that has developed reflects these large-scale gradients of climate, tidal amplitude, and salinity. Virtually all European salt marshes are also the product of thousands of years of human impact. Metal mining in the catchments that drain into the estuarine marshes of southern Europe dates from 5,000 years BP. The conversion of marshes for salt production (salinas) can be traced back 2,600 years. Occupation of coastal marshes by herdsmen occurred at about the same time. Coastal embankment and land claim for grazing and cultivation have been practiced around margins of shallow seas, embayments, and estuaries at least since Romano-British times in northern Europe. More recently, coastal protection schemes (sluiced barriers and dams) have reduced the tidal influence and salinity in large areas of the remaining estuarine marshes, with major effects on community structure and system functioning. Large engineering projects, often associated with developing ports and maintenance of their deep-water channels, have changed the sedimentary environment to promote the rapid development of new marsh systems in some areas and erosion elsewhere. Rising relative sea levels, partly the result of isostatic adjustment but increasingly associated with climate change, are leading to the erosion of ancient marshes, particularly around the southern North Sea. The economics of sea defense there now favor the abandonment of former land claim, by breaching or removing embankments (managed coastal realignment), and regeneration of salt marsh. Increasing recognition of the ecosystem services provided by salt marshes and their importance for biodiversity is also a driver for the reinstatement of tidal influence and restoration of marsh function.

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Grazing by livestock has modified many marshes over centuries. The impacts of grazing by migratory geese have increased dramatically as their population sizes have been affected by changes in agriculture elsewhere. Human introduction of invasive species of Spartina from the Americas has transformed many marshes. As elsewhere, the coastal and estuarine waters of Europe have experienced eutrophication with nitrogen and phosphorus from agricultural runoff, human waste, and industrial processes in the catchments of its great rivers. Similarly, salt marshes have become sinks for diverse industrial and agricultural pollutants, including heavy metals, radionuclides, hydrocarbons, and persistent pesticides. The European Atlantic is the world “hot spot” for oil tanker spills: salt marshes on the coasts of the Bay of Biscay and English Channel have experienced repeated, highly destructive oilings. In the face of this apparent gallimaufry of threats, many salt marsh sites in the twenty-five countries of European Union are now protected by its Habitats and Birds Directives. These have established a large network of protected sites (Natura 2000) in which wild fauna and flora of European interest and their habitats must be maintained at, or restored to, favorable conservation status. There are good prospects for the mitigation of human impacts, depending on local circumstances. Managed coastal realignment in response to sea-level rise provides the opportunity to create entirely new or replacement marshes. These will often not follow the successional trajectories of undisturbed sites but, nevertheless, may be engineered to offer an appropriate range of marsh elevations and environments. In all marshes, highest diversity is guaranteed when the range of pioneer to climax stages can be maintained. Management of an adequate sediment supply is a prerequisite to prevent aging or erosion. In estuarine settings, a broad salt marsh is more likely to provide a gradient from salt to brackish conditions and hence higher diversity. Grazing by livestock is potentially a valuable management tool. It can be applied to inhibit or reverse succession and therefore the effects of eutrophication, by reducing standing biomass and species dominance, and by mobilizing nutrients. High grazing intensity creates a very short turf with good habitat for winter- and spring-staging geese. Intensively grazed marshes also harbor higher plant species diversity than long-term ungrazed marshes, especially at the small scale; at the scale of a whole salt marsh, low stocking rates could generate the highest diversity. On a regional scale (e.g., the Wadden Sea), application of the whole range of grazing pressures might be preferred.

European coastlines stabilized in approximately their current positions after a period of rapid sea-level rise following the last glaciation. Marine sediments have been deposited and consolidated in many places, such that vegetated upper-marsh plains on the Atlantic and southern North Sea coasts of Europe have persisted for more than 6,000 years BP (Funnell and Pearson 1989; Funnell and Boomer 1998; Allen 2000b). The marshes that we see today have been molded by a combination of human 312

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activity and underlying patterns of natural change that have mainly been forced by changing climate and relative sea level. Extensive marshes tend to occur where rising sea levels, often in conjunction with an isostatic drop of the land surface, have accommodated long-term sedimentation. Their persistence, however, depends on the net effect of the various forces promoting sediment accretion and erosion. European salt marshes range from the arctic to the Mediterranean (fig. 16.1) and are thus

FIGURE 16.1 The distribution of main salt marsh sites in Europe (based on Dijkema 1984 and Burd 1989).

highly diverse as environments (Chapman 1977; Dijkema 1984). Those on the coasts of Iceland, northern Scandinavia, and arctic Russia have much in common with the North American arctic marshes (Nordhagen 1954). The marshes of the Baltic Sea (fig. 16.2) are characterized by lower (brackish) salinities and a small tidal range; they tend to be transient because they are on a coastline that is rising, as a result of isostatic adjustment (Tyler 1969; Siira 1985). In the most northern part of the Baltic, uplift reaches nine millimeters per year (Jerling 1999). Many of the more intensively studied salt marsh systems are on the macro- or

mesotidal coasts and estuaries of the North Sea, English Channel, Irish Sea, and Atlantic Ocean; their climates range from cool temperate in the northeastern part of this area, through mild oceanic, to a Mediterranean type in southwest Iberia. Isostatic adjustment means that marshes have developed on a rising coastline around Scandinavia and in the north and west of the British Isles, but elsewhere land levels are falling. Marshes on the Mediterranean coast of Europe have formed under microtidal conditions and experience a combination of arid, hot summers and mild (generally frost-free), wet winters—a climate that can give rise to severe

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FIGURE 16.2 Livestock-grazed sea shore meadow at Gotland (Sweden) in the Baltic. The vegetation is dominated by Armeria maritima with patches of Scirpus maritimus. Barnacle geese breed on small islands along the coast and forage with their young on the short vegetation. Photo by Sandra Van Der Graaf.

seasonal hypersalinity. These European marshes are akin to those of the African Mediterranean coast. The larger marshes tend to be found in the deltas of large rivers, such as the Ebro (Spain), Rhône (France), and Po (Italy), and are also affected by seasonal changes in their flows. Currently there are about 176,000 hectares of salt marsh around the Baltic and Atlantic coasts of Europe (Bakker et al. 2002) and a smaller area around the Mediterranean coast. In addition there are relatively small areas of inland salt marsh, mainly in the steppes area of central and eastern Europe, and associated with gypsum outcrops in the Iberian Peninsula (fig. 16.1). Human influence on European salt marshes can be traced back for millennia. The metal mining that led to the acidification of marshes in the estuary of the Rio Tinto in southwest Spain dates back at least five thousand years, to the Copper and Bronze ages; mining was particularly active in Phoenician and Roman times, and was restarted in the late nineteenth century (Davis et al. 2000). The production of sea salt from artificial salt pans built on coastal marshes can be traced back some 2,600 years to Ostria, on the mouth of the River Tiber, Italy (Marín and d’Ayala 2000). In northwest Europe (the Wadden Sea), herdsmen probably first occupied salt marshes for livestock grazing also around 2,600 years ago, during an interlude of sea314

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level recession, when there was reduced risk of flooding (Bakker et al. 1997, 2002). As sea levels started to rise again, people built dwellings on artificial mounds (fig. 16.3) and eventually began to construct low walls (dikes) to promote sedimentation and reduce saline flooding, thus creating the first “summer polders” (discussed later). Fossil evidence indicates that there were probably no pristine marshes left in the Wadden Sea area by 2,000 years BP (Bakker et al. 2002). Abundant archaeological remains, including Roman fields and ditches, testify to Romano-British occupancy of the marshes flanking the Severn Estuary in Britain (Allen 2000a). Although habitats affected by human activity now dominate the landscape of the Netherlands, the diversity of habitats reconstructed for the Rhine and Meuse delta in its natural state, 3,700 years BP, resembles the patchwork of habitats still present in the Pechora Delta in Arctic Russia (Van Eerden et al. 2005). Although salt marshes tend to retain a more “natural” appearance than most other habitats in developed agricultural and industrial landscapes, they are just as much the product of human exploitation as any other (Adam et al. 2008). The history of deliberate human impacts has included embankment and drainage for land claim, either to create agricultural land or for industrial and urban development, livestock

FIGURE 16.3 Ungrazed mainland salt marsh near the Hamburger Hallig (Germany). The vegetation shows a mosaic of Artemisia maritima (Seriphidium maritimum), Atriplex portulacoides, and Elytrigia atherica. People live on the artificial mound without further protection from a sea wall. It is a small remnant of a previously common type of landscape along the mainland coast of the northern Netherlands and Germany before the embankments started in the twelfth century. Photo by Roos Veeneklaas.

grazing, and harvesting of hay, wildfowl, and fish (Adam 2002). More or less inadvertent impacts can be no less profound, especially if they lead to increased erosion: nutrient enrichment, pollution with agricultural pesticides and industrial effluents, the introduction of invasive aliens, coastal protection engineering, and the dredging of deep-water shipping channels into ports. To this list, we must now add the consequences of anthropogenic, global climate change, with its concomitant sea-level rise, not to mention our responses to them for coastal defense and ecological restoration. Such activities have altered not only the area and distribution of marshes but also the structure and functioning of many of those remaining.

EMBANKMENT AND LAND CLAIM Embankment and land claim for grazing and cultivation have been practiced at least since Romano-British times in northern Europe. Although this represents a large net loss of mature salt marsh, continuing sedimentation seawards of the embankments may permit new marsh colonization; consequently, new marshes have often developed, allowing successive cycles of marsh accretion and land claim (Kestner 1962).

The prime example of land claim is the extensive Dutch, German, and Danish polderland, which includes much former salt marsh from the margins of shallow seas, embayments, and estuaries (Bakker et al. 2002). It was a response not only to address a need for agricultural land and defense from rising relative sea levels, but also to counter a threat from coastal land subsidence associated with medieval peat drainage and digging (Esselink 2000). Initially, the strategy was to enhance natural accretion by digging drainage ditches and building low dams. Low sea walls were generally constructed above the level of the mean high tides, thus leaving a foreland of marsh and mudflat to extend seaward. The marshes behind were drained by a rectilinear network of ditches, with main drains discharging through sluices in the dams. The low dams created summer polders that could be used for grazing or a hay crop but allowed tidal deposition of sediments by high tides and winter storms. Sedimentation could be enhanced with groynes on the intertidal flats along the coast. The most widely adopted procedure involved the construction of sedimentation fields surrounded by brushwood groynes and intersected by a dense system of ditches (the Schleswig-Holstein method; fig. 16.4). Eventually, when accretion was sufficient, the

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FIGURE 16.4 Artificial mainland salt marsh in Schleswig Holstein (Germany). The salt marsh has developed because of the shelter of brush wood groynes that form sedimentation fields. Parallel ditches were dug for drainage. The salt marsh is heavily grazed by livestock, thus featuring a “golf course” dominated by Puccinellia maritima. The left section of the photograph has not been grazed for fifteen years and is dominated by Elytrigia atherica. Photo by Roos Veeneklaas.

embankments could be raised and all tidal influence excluded. Another cycle of land claim might then be started on developing low marshes to seaward (Bakker et al. 1997, 2002; Esselink 2000). Such cycles of subsequent land claims developed especially along the mainland coast with relatively high rates of sedimentation. The back barrier marshes on the islands had low rates of sedimentation. Therefore, little land reclamation took place there. Nevertheless, the area of salt marshes on the back barrier islands has greatly increased during the past century. This can be attributed to artificial sand dikes that provide sheltered sites for salt marsh development. Nowadays, the area of salt marshes in the Netherlands amounts to nearly 9,000 hectares (3,500 hectares on barrier islands, and about 5,300 hectares along the mainland coast, subdivided in 4,000 hectares foreland marsh, 960 hectares summer polder, and 300 hectares de-embanked summer polder (Bakker, Bunje, et al. 2005). In Britain, embankment and land claim of salt marshes have mainly been for agriculture. Initially, this was for grazing land, but many areas were subsequently developed for arable farming. Activity has been greatest in the shallow, sediment-rich waters around the major estuaries, particularly in the Wash (Lincolnshire and Norfolk), the Severn Estuary, the greater Thames Estuary (Essex and north Kent), the 316

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Humber Estuary, the Ribble Estuary (Lancashire), and Morecambe Bay (Cumbria and Lancashire). It is estimated that some 53,400 hectares of salt marsh may have been lost in this way, more than half since the seventeenth century (Doody 1992). Although this is not a net loss, because new marshes have often developed in response to embankment, inevitably the new marshes tend to be lower in the tidal frame, less mature, and less species-diverse than those lost. Potentially they might develop into mature, more diverse marshes, but there are few examples of this. Such replacement marshes are truly anthropogenic. Similarly, the embankment of salt marshes for agricultural purposes was carried out along the coast of northwest France, especially in the estuary near Mont Saint Michel (fig. 16.5; Lefeuvre and Bouchard 2002). The Dutch word polder has even been assimilated into the French language. The conversion of coastal marshes or lagoons to salt pans (salinas) for the production of salt has had an impact particularly on the Mediterranean coast, where an annual water deficit promotes efficient concentration and crystallization; hence, salt can be extracted on an industrial scale. The Messolonghi salt works alone produces about 110,000 tonnes per year, satisfying 40 percent of the Greek domestic demand (Petanidou 2000). The importance of salt to humans is such that, historically, many

FIGURE 16.5 Ungrazed mainland salt marsh near Mont-Saint-Michel (France). The vegetation is dominated by Elytrigia atherica and patches of Festuca rubra. Photo by Johan Van De Koppel.

salinas have also been constructed on the Atlantic and North Sea coasts. Active or abandoned salinas occupy about seventy-five thousand hectares of the littoral area of the European Union (Marín and d’Ayala 2000). They are now recognized as an important European habitat in their own right under both the Birds and the Habitats Directives of the European Commission (European Commission 2003). Significant areas of salt marsh have been claimed for the development of ports, refineries, and other industrial developments at many locations around the coast of Europe; such activities rarely permit any further expansion of the salt marsh.

COASTAL ENGINEERING More recently, coastal protection schemes (sluiced barriers and dams) have reduced the tidal influence and salinity in large areas of the remaining estuarine marshes, with major effects on community structure and system functioning. Only in the twentieth century did the Dutch Wadden Sea reach its present limit in the north, with the large enclosures of the Zuiderzee, now IJsselmeer (1932), and the Lauwerszee Estuary (1969). The Westerschelde in the southwest of the Netherlands reached its present extent with the enclosure of the Braakman in 1952.

After the large inundations in 1953, the coastline was shortened in the southwest Netherlands. As part of this project, the Oosterschelde became partly closed by a semiopen seawall in 1987. It remained an inlet, however, with slightly (10 percent) reduced tidal amplitude. This resulted in salt marsh species present moving down the elevation gradient (De Leeuw et al. 1994). Former tidal inlets that were completely cut off from the tidal influence became saline or freshwater lakes. The terrestrial vegetation was transformed into either scrub and forest under laissez-faire management or into livestock pastures with large populations of geese (Van Wieren 1998). Seashore meadows in the Baltic do not suffer from coastal protection works. On the contrary, they tend to expand naturally as a result of the isostatic uplift (Jerling 1999). The more saline areas only periodically exposed by the nontidal water level support Salicornia spp. In sheltered places or areas with lower salinity, Schoenoplectus maritimus and S. tabernaemontani occur, sometimes together with Aster tripolium and Triglochin maritima. The zone between mean summer water level and the level of highest flooding in summer harbors a turf of Puccinellia maritima under relative high salinity and of Agrostis stolonifera under brackish

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conditions. Higher on the shore, flooding is frequent during winter, and Juncus gerardii and Glaux maritima dominate. At the highest zone, a sward of Festuca rubra with Armeria maritima occurs (Jerling 1999). The isostatic uplift continually creates new land and alters the physicochemical properties of existing land, resulting in a downward shift of plant species (Cramer and Hytteborn 1987). Southern Europe has also been generally less affected by coastal defense works. However, there have been huge engineering projects that have had considerable effects on the salt marshes. Long-term studies of the effects of these projects can be valuable in illuminating some of the processes of salt marsh succession and degeneration. The perturbations to the sedimentary environment can offer experimental manipulations on a scale normally beyond the reach of ecologists. Such a model system was provided by the construction of the Juan Carlos I Dike, which substantially changed physiographic conditions at Odiel Marshes (fig. 16.6), in a Holocene estuary on the coast of the Gulf of Cádiz, in southwestern Spain (Castellanos, Figueroa, and Davy 1994). Nearly fifteen kilometers in length, the raised dike now carries a road across an extensive area of salt marsh and projects into the Atlantic Ocean. It was built to protect the navigable channel to the port of Huelva, in the common estuary of the Odiel and Tinto rivers, from

the eastward drift of sandy deposits. The development of new marshes on low-lying tidal flats with different tidal drainage regimes was initiated in 1977. The early colonist, Spartina maritima, is tolerant of highly reducing sediments and long periods of tidal inundation (Castillo, Fernandez-Baco, et al. 2000). Colonizing clonal clumps rapidly expanded and promoted sediment accretion locally. In one lagoon, where drainage was impeded by a sand spit and sediment redox potentials remained low, the clumps eventually coalesced to form Spartina marsh. However, in a similar lagoon with rapid tidal drainage, complex successional changes ensued. Here, the raised, expanding clumps of Spartina facilitated the invasion and eventual dominance of Sarcocornia perennis, which was more competitive under the less reducing conditions at higher elevation (Castellanos et al. 1994). A more recent invasion, of the tops of the coalescing tussocks, has been by a hybrid of S. perennis with the larger S. fruticosa, which is characteristic of the landward margins of the marshes; hybridization appears to have occurred in situ on the tussocks and is facilitating further successional development (Figueroa et al. 2003). This, however, is only one facet of change at Odiel: elsewhere there has been extensive loss of salt marsh by both vertical erosion and horizontal retreat of channel banks. In the southern part of the marsh, which experiences the greatest human pressure, the banks

FIGURE 16.6 Newly developed salt marsh at Odiel marshes, on the opposite side of the joint estuary of the Rio Tinto and Rio Odiel from industrial development at Huelva, southwest Spain. Photo by A. J. Davy.

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of the main channels are retreating at an average rate of twenty-five centimeters per year, and the marsh surface is deflating by up to five centimeters per year (Castillo, Luque, et al. 2000; Castillo et al. 2002).

SEA-LEVEL RISE AND MANAGED COASTAL REALIGNMENT Certain areas of Europe, such as in large parts of the Baltic and northwest Britain, are currently buffered from the effects of eustatic sea-level rise by isostatic uplift. Elsewhere, ancient marshes are suffering erosion in various parts of Europe (Harmsworth and Long 1986; Castillo, Luque, et al. 2000; Cooper, Cooper and Burd 2001; Castillo et al. 2002; Van der Wal and Pye 2004). At least in the southern North Sea, relative sea levels are rising at about three

millimeters per year, and increasing wave action and extreme water levels are associated with an increase in storminess. The economics of sea defense in parts of this region now favor the abandonment of former land claim, by breaching or removing embankments—managed realignment or retreat (fig. 16.7). By absorbing wave and current energy, the breached sites are expected to provide protection for the hinterland (French 1999) and for the new sea walls built behind them. They also provide opportunities for the reestablishment of salt marsh. Such restoration or reconstruction is probably the most important contemporary human influence on salt marshes in Europe. In contrast, reinforcing the original embankments with expensive “hard engineering” would have been likely to lead to the loss of salt marsh through “coastal squeeze,” although the role of coastal squeeze in

(A)

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FIGURE 16.7 Coastal realignment at Brancaster West marsh, Norfolk, UK. (A) Panoramic view of regenerating salt marsh in 2005, three years after breaching of the sea wall; (B) culvert in the sea wall allowing exchange of tidal waters via the creek system. Photos by A. J. Davy.

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recent salt marsh loss in southeast England is still contentious (Hughes and Paramor 2004; Morris et al. 2004; Wolters, Bakker, et al. 2005). Furthermore, increasing recognition of the ecosystem services provided by salt marshes and their importance for conservation of biodiversity is leading to the reinstatement of tidal influence and restoration of marsh function in certain estuaries. Managed retreat, or setback, has been widely practiced historically in Britain, with archaeological and stratigraphic evidence dating back to Romano-British times (Allen 2000a). Most active deliberate retreat sites are younger than ten years. Much current interest focuses on predicting and managing the development of salt marsh in the longer term. Valuable insights may come from historical, accidental dike breach events; storm surges, particularly those in the North Sea in 1897, 1921, and 1953, caused breaches that were never repaired and thus allowed natural regeneration of salt marsh (Burd 1994, 1995; Crooks et al. 2002; Wolters, Garbutt, and Bakker 2005b). Comparison of two such regenerated sites, a century after dike breach, with adjacent areas of natural marsh in Essex, England, revealed reasonably close convergence in vegetation composition after this period (Crooks et al. 2002), although differences persisted. Good convergence, especially on a shorter time scale, is not the general experience (Onaindia, Albizu, and Amezaga 2001). A large-scale analysis of seventy de-embankment sites around the coasts of England, France, Belgium, the Netherlands, and Germany assessed floristic development of marshes in terms of the proportion of the regional pools of halophytes present (Wolters, Garbutt, et al. 2005). This proportion (saturation index) depends on the range of tidal environments available (expressed as the proportion of vertical difference between high-water spring and highwater neap tides that is occupied by the elevational range of the site) and the area of the site. Saturation index also tended to be higher in sites less than twenty years old than in sites representing later successional stages. This might 320

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be related to the dominance of a single plant species in the absence of livestock grazing (discussed later). Marsh development, or lack of it, in retreat sites also provides model systems for investigating salt marsh succession and function. An early managed realignment experiment at Tollesbury in southeast England (breached in 1995) has been more successful in producing intertidal mud flats than salt marsh so far (Garbutt et al. 2006). Consolidation and oxidization of the original land claim sediments mean that much of their area lies lower in the tidal frame than adjacent salt marshes. In consequence, they may be initially unsuitable for the establishment of halophytes, because of prolonged, tidal inundation and poor drainage. Rapid sedimentation in such young systems (Wolters, Garbutt, et al. 2005b) tends to produce unconsolidated, weak, and saturated substrates. Drainage can be further impeded by a nearly impervious soil pan (aquitard), having been formed during the terrestrial phase, by the absence of a creek system and by a limited number of breaches (sometimes only one) in the embankment for tidal exchange (Crooks et al. 2002). The hypoxic, reducing sediment chemistry may be exacerbated by the decomposition of former glycophytic vegetation. Other limitations may include the absence of a halophyte seed or propagule bank and poor dispersal by water currents from the nearest sources of propagules (Wolters, Garbutt, et al. 2005a). Much remains to be learned about the sedimentary and ecological processes involved in the regeneration of salt marshes.

CLIMATE CHANGE The recent increase in the rate of sea-level rise is one manifestation of human impact on global climate (International Panel on Climate Change [IPCC] 2001). Salt marshes and their biota are also likely to be affected directly or indirectly by increasing atmospheric carbon dioxide concentrations, higher temperatures, changing patterns of rainfall, and increased irradiation with

ultraviolet-B (UV-B) resulting from destruction of the stratospheric ozone layer. Most European salt marsh species have C3 photosynthesis, but certain of the widely distributed and productive species have evolved the C4 pathway (e.g., Spartina spp.), with its greater water use efficiency (advantageous in saline and arid environments) and the virtual elimination of photorespiratory losses (more advantageous at higher temperature, but the advantage may be negated by increasing atmospheric carbon dioxide concentration). Predicting the outcome of climate change is complex, as it also needs to take into account effects of temperature on, for example, phenology and photochemical inhibition, as in the study of Spartina anglica and Puccinellia maritima by Long (1990). The appropriate long-term field experiments have not been carried out on European salt marshes. There is evidence that increasing UV-B irradiation may depress photosynthesis and growth more in Aster tripolium than in S. anglica (Van de Staaij et al. 1990) and that Elymus athericus will be more severely affected by increased UV-B with a concomitant increase in carbon dioxide concentration (Van de Staaij et al. 1993).

GRAZING The tradition of livestock grazing at least the higher parts of salt marshes with sheep and cattle is probably as old as human habitation (fig. 16.8). Few mature European marshes will not have been grazed at some stage of their history, and the modifications to their vegetation are often very apparent. The Baltic seashore meadows have been heavily influenced by cattle grazing (Tyler 1969). Large areas of “saltings” pastures, composed mainly of Puccinellia maritima, Agrostis stolonifera, and Festuca rubra, are found in northern France and, on a smaller scale, in most of the bays and estuaries of the west coast of England and Wales from the Solway to the Bristol channel (Ranwell 1972). Centuries of high stocking rates in some areas have produced “billiard table smooth” or “golf course” swards of these grasses, as in

FIGURE 16.8 Livestock-grazed salt marsh at the island of Schiermonnikoog (The Netherlands) featuring a short sward with Festuca rubra. Elytrigia atherica that dominates the vegetation inside the thirty-year-old exclosure. Photo by Roos Veeneklaas.

Morecambe Bay, United Kingdom, or Schleswig Holstein, Germany, although a surprising diversity of other species can often persist in the short turf (Adam 2000). Elsewhere, heavy grazing has reduced species diversity by eliminating dicotyledonous herbs and taller species (especially Atriplex portulacoides, Elytrigia atherica, and Limonium spp.) in favor of the low-growing halophytic grasses (especially Puccinellia). Experimental studies by Ranwell (1961) showed that even S. anglica marsh growing on firm sediment could be converted to Puccinellia by sheep grazing in five to ten years. In Britain, there are many more marshes, which although ungrazed today, still show evidence of former exploitation for grazing (Doody 1992). In the same way as centuries of land claim have given way to managed retreat, there has been a reversal in salt marsh management. The trend in recent decades has been toward the abandonment of traditional grazing practices. It might be imagined that this would herald a return to more “natural” systems, but in reality, it often leads to a new layer of anthropogenic effects. Fine-grained vegetation on increasingly eutrophic sediments, molded by long-term grazing, becomes highly susceptible to invasion by single, dominant species (Bakker 1985, 1989; Bakker et al. 2002). Loss of biodiversity is associated with increasing biomass and litter accumulation. Brackish seashore meadows in the Baltic become

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FIGURE 16.9 The salt marsh of the island of Schiermonnikoog (The Netherlands) without livestock grazing. Plots show the effects of five years of exclusion from grazing by hares and geese (block at the right) and exclusion only of geese (block on left), which shows no differences from the control area where both species graze. Photo by Jaap de Vlas.

dominated by the tall reed Phragmites australis (Dijkema 1990; Jutila 1999). The tall, coarse grass Elytrigia atherica (Elymus athericus) now dominates the higher parts of many European marshes. Mature plants are relatively unpalatable to grazers, but, once established, they can readily invade stands of shorter species such as Festuca rubra. Grazing by hares and geese (fig. 16.9) can, however, delay the establishment of its seedlings, particularly at the lower end of a productivity gradient (Kuijper, Nijhoff, and Bakker 2004; Kuijper, Dubbeld, and Bakker 2005). In more productive salt marshes, geese and hares lose control, and tall plant species take over. Tall vegetation can only be removed by livestock grazing, thus facilitating the grazing of smaller herbivores such as geese and hares (Kuijper and Bakker 2005). Tessier et al. (2003) found that vegetational trajectories in responses to the cessation of grazing on marshes of the Atlantic coast of France were very dependent on local drainage conditions: a middle marsh area with impeded drainage remained dominated by Puccinellia, whereas Atriplex (Halimione) portulacoides rapidly replaced it on a well-drained low marsh. However, Elytrigia began to increase in abundance on the middle marsh after seven years without grazing and is anticipated to become dominant. Spread of Elytrigia might depend on the rate of sedimentation and hence nitrogen input (Schröder, Kiehl, and Stock 2002). 322

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Loss of biodiversity is associated with increasing biomass and litter accumulation after longterm cessation of livestock grazing. Plant species diversity is lower at marshes without livestock in the Wadden Sea (Bos et al. 2002). Goose use of 164 coastal sites around the Wadden Sea is negatively related to canopy height and positively to livestock density (Bos et al. 2005). The characteristic halobiontic invertebrate community of salt marshes, including insects in the Wadden Sea (Andresen et al. 1990) and spiders near Mont Saint Michel, France (Pétillon et al. 2005), can be replaced by a nontypical ruderal community. Most of the debate to date has focused on the presence or absence of grazing; in fact, good experimental and long-term data on intermediate and rotational grazing are lacking. The numbers of migratory geese have increased dramatically as their population sizes have been affected by changes in agriculture elsewhere (Van Eerden et al. 2005). Their impact is still a matter for debate. Grubbing Greylag geese affect the salt marsh habitat by complete removal of the vegetation in the Dollard marshes at the border of the Netherlands and Germany (Esselink 2000). Grazing of Barnacle geese seems to result in secondary pioneer vegetation in the Dollard. However, this only occurs in the presence of livestock grazing to facilitate feeding by the smaller herbivores such as geese (Esselink 2000).

INVASIVE SPECIES The introduction by man of invasive species of Spartina from the Americas has transformed many European marshes. The most widespread consequences to date, affecting mainly northern Europe, arose from the accidental introduction of S. alterniflora, probably in shipping ballast sediments, from North America to southern England in the early nineteenth century. Its hybridization with the native S. maritima to form S. ⫻ townsendii and the subsequent doubling of chromosomes to produce the invasive, fertile amphidiploid S. anglica is a classic, textbook example of speciation by hybridization (Gray, Marshall, and Raybould 1991), with S. alterniflora known to be the maternal parent (Ferris, King, and Gray 1997; Ayres and Strong 2001; Baumel, Ainouche, and Levasseur 2001). S. anglica has proved to be a vigorous invader that can colonize mudflats seaward of most other salt marsh vegetation, including its native parent, which is now very rare in Britain and the Netherlands. It produces extensive monospecific stands, reminiscent of its American maternal parent. Human influence was not only in the species introduction. The value of the species for stabilizing mudflats and enhancing sedimentation for coastal protection, future grazing marsh, or land claim was rapidly recognized (Carey and Oliver 1918), and clonal offsets were planted extensively around the northwestern European coast from northern Denmark to southwestern Netherlands, and, indeed, many other places in the world; it appears to have spread naturally around the French coast (Goodman, Braybrooks, and Lambert 1959; Ranwell 1967). It is now one of the European salt marsh dominants, particularly on fine muds in lower parts of the tidal frame, where it may adopt an early successional role and promote rapid accretion (Ranwell 1964). In contrast, the native S. maritima has been in decline over the same period in northern Europe; as it displays very little genetic variation, its decline may be related to a virtual restriction to vegetative propagation at higher latitudes rather than

any direct competition from S. anglica (Yannic, Baumel, and Ainouche 2004). Spartina patens has also been introduced, probably originally from North America to the Mediterranean, as a packing material. It has spread more recently to the Galician coast of northwest Iberia, where it has formed extensive, dense stands that threaten the native biodiversity of the higher parts of the marshes (San Leon, Izco, and Sanchez 1999). The South American Spartina densiflora was similarly introduced to the estuaries of Atlantic southwest Spain and is now flexing its muscles as an ecosystem engineer. It has invaded the marshes of the Gulf of Cádiz, which include some of the largest and most diverse in southern Europe. S. densiflora has become very abundant there, displacing indigenous species, particularly on the middle elevations of the marsh, where it can form monospecific stands (Nieva et al. 2001, 2005). Its dense tussocks retain large standing crops of litter that deprive competitors of light, and it can spread rapidly, by both copious seed production and clonal growth. No native species can colonize such a wide range of sediment and tidal conditions (Nieva et al. 2005), although S. densiflora is less tolerant of the highly reducing sediments low in the tidal frame than the native S. maritima (Castillo, Fernandez-Baco, et al. 2000). The threat posed by this invasion to southern European and northern African marshes cannot yet be fully assessed but is probably already too late to prevent its spread. The invasive character of Phragmites australis on brackish seashore meadows in the Baltic and that of Elytrigia atherica on higher marshes as widespread as the Wadden Sea, Britain, and France has been referred to previously in the context of responses to the cessation of grazing. The extension of E. atherica on to well-drained lower marshes appears to have been substantially limited by competition with Atriplex portulacoides (Bockelmann and Neuhaus 1999). Recently, genetically differentiated populations have succeeded in colonizing lower marshes, which suggests that it has sufficient genetic variation to

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extend its range further (Bockelmann et al. 2003). It is also favored by nutrient enrichment, especially by nitrogen (Kuijper et al. 2005).

EUTROPHICATION AND POLLUTANT LOADING FROM RIVER CATCHMENTS As elsewhere, the coastal and estuarine waters of Europe have experienced eutrophication with nitrogen and phosphorus from agricultural runoff, human waste, and industrial processes in the catchments of its rivers. Similarly, anthropogenic heavy metals and pesticides are discharged via the rivers. Thus, accreting salt marshes are likely to become sinks for industrial and agricultural pollutants. The metal mining and refining industries of the Rio Tinto and Rio Odiel catchments of southwestern Spain are a conspicuous example. Discharges from the rivers have produced high heavy metal loadings in the estuary. These loadings are augmented by contaminants released directly from the Huelva Industrial Centre, sited on the estuary itself (fig. 16.6). The Odiel Marshes were declared a Biosphere Reserve by UNESCO in 1983 and include a Special Protection Area for birds (a designation of the European Union). Luque et al. (1999) have reported elevated concentrations of nine heavy metals in the tissues of eight common halophytes in these marshes. Particularly high concentrations of metals were found in Zostera noltii, Spartina maritima, and S. densiflora, and samples from species on lower parts of the marsh tended to have the highest concentrations. The tissue concentrations of copper, lead, and zinc would have been very toxic to other species, but no signs of toxicity were observed at Odiel. Nevertheless, the accumulation and fluxes of these metals in the food webs of the marsh may have consequences that are currently unknown. Metal loadings in the UK Blackwater Estuary are much lower (Emmerson et al. 1997), but the sediments of newly regenerated salt marshes at a managed retreat site there have already been shown to be sinks for anthropogenic lead, chromium, and copper (Macleod et al. 1999). 324

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Major reductions in input and concentration of metals in the Wadden Sea mainly occurred in the late 1980s until the early 1990s and have continued more moderately until 2002. The proposed background level in sediment was exceeded by mercury (by a factor of three to ten) and lead (by a factor of two to four). The proposed background level in blue mussel Mytilus edulis was exceeded for cadmium (a factor of two to three), lead (a factor of two to nine), and lead (a factor of two to four) (Bakker, Van der Heuvel-Greve, and Vethaak 2005a). Effects on salt marshes are not known. The persistence of most xenobiotics presents the largest problem for achieving the goal of downward trends of concentrations in the sediment and biota of the Wadden Sea. PCBs are still widespread, but concentrations have decreased over the past twenty years. However, they still exceed agreed background levels by many factors. Lindane and its metabolites are mostly decreasing, although ␤-HCH is persistent in the higher trophic levels. Concentrations of DDT, DDE, and HCB also decreased in the 1990s, the latter with a relatively late response in the higher trophic level (birds’ eggs; Bakker et al. 2005b). Again, effects on salt marshes are unknown. Riverine nutrient input also decreased gradually between 1997 and 2002. For example, winter phosphate concentrations in the Wadden Sea have decreased since the mid-1980s to about 1 ␮M. The decreasing nutrient input by the Rhine and Meuse rivers has had a significant effect on phytoplankton biomass in summer. Long-term data from the western Dutch Wadden Sea show a decreasing trend in the duration and extent of algal blooms. The decreasing nutrient input also had a significant effect on the autumn inorganic nitrogen (NH4 ⫹ NO3) and therefore on the organic matter turnover (Van Beusekom et al. 2005). These data have not been related to the salt marsh ecosystem. Data collected along a productivity gradient on a barrier system revealed that nitrogen in tidal waters contributed most to ecosystem nitrogen accumulation rate in early successional stages, whereas atmospheric

deposition was more important at later stages. Tidal influence was low at high marsh elevation sites. Here, atmospheric deposition was the dominant exogenous nitrogen source both in young and old marshes (Van Wijnen and Bakker 2000). In general, European salt marshes tend to be nutrient limited, and numerous nutrient addition experiments have demonstrated the potential for exogenous nitrogen and phosphorus to bring about vegetational change (table 16.1). During the past century, the Baltic Sea has been transformed from a clean, oligotrophic water body to a contaminated, mesotrophic one. Its drainage basin supports a human population of about eight-five million in fourteen countries. The phosphorus load has increased about eightfold, and the nitrogen load about fourfold. This has increased pelagic primary

production by an estimated 30 to 70 percent and sedimentation of organic carbon by 500 to 1,000 percent, resulting in oxygen depletion (Snoeijs 1999). The very limited water exchange with the ocean retains pollutants inside the Baltic Sea for a long time. Furthermore, chemical and biological degradation of pollutants is slow as a result of low temperatures and long winters, especially in the northern basins (Snoeijs 1999). There is evidence that recent increases in the use of persistent herbicides may be having insidious effects on the stability of salt marshes on the coast of southeast England. Marsh sediments appear to act as a sink for the widely used triazine herbicides (e.g., simazine and atrazine; Leggett, Bubb, and Lester 1995; Meakins, Bubb, and Lester 1995). Mason et al. (2003) showed that concentrations of herbicides within ranges

TABLE 16.1 Responses (⫹, positive; ⫺, negative) of European salt marsh species to nutrient additions in field experiments

Location

Primary Limitation

Baltic Sea

N (NH4) or P

Additional Responses

Main Species Responding

Source

Juncus gerardii (⫹) Plantago maritima (⫹)

1

Wadden Sea

N

Low marsh: Suaeda maritima (⫹) Puccinellia maritima (⫺) High marsh: Spergularia maritima (⫹)

2

Wadden Sea

N

At only one site of three: Elytrigia atherica (⫹)

3

Wadden Sea (Schiermonnikoog)

N

P

Elytrigia atherica (⫹) Seriphidium maritimum (⫹) Limonium vulgare (⫺)

4

North Norfolk, UK (upper marsh)

N

P

Atriplex portulacoides (⫹) Limonium vulgare (⫹) Puccinellia maritima (⫹) Salicornia europaea (⫹) Spergularia maritima (⫹) Suaeda maritima (⫹) Armeria maritima (⫺)

5

Mont Saint Michel, northern France

N

Puccinellia maritima (⫹) Suaeda maritima (⫹)

6

SOURCES: 1, Tyler (1967); 2, Kiehl et al. (1997); 3, Leendertse (1995); 4, Van Wijnen and Bakker (1999); 5, Jefferies and Perkins (1977); 6, Tessier et al. (2003).

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observed in the field reduced the photosynthetic efficiency of both epilelic diatoms and vascular salt marsh plants. Such sublethal concentrations reduced the growth rates and production of extracellular polymeric substances (EPS) in diatoms, with parallel reduction of sediment shear strength (stability). Further reductions of vascular plant photosynthesis resulted from sediment deposition on the leaves near creek banks, a more likely occurrence with eroding sediments. Clearly, there are complex interactions with considerable potential for destabilizing marsh surfaces, especially in the face of rising sea levels and increased wave action.

OIL SPILLS Spillages of oil from ships and refineries frequently come ashore on salt marshes, where they become trapped (Baker 1979). The lighter oils are highly toxic to plants and animals, whereas heavier crude oils tend to have a smothering effect. This is a chronic problem near refineries and major ports, where repeated discharges and effluents can destroy the vegetation or alter its composition in favor of the relatively tolerant species. Catastrophic releases of oil from wrecked tankers occur periodically (table 16.2), and the resulting slicks can have long-lasting effects on salt marshes. The Atlantic coast of Europe is the worst-affected region worldwide for such incidents (Vieites et al.

2004), with the important salt marshes on the Brittany coast of France having suffered repeated oilings. The notorious sinking of the Torrey Canyon, off the Isles of Scilly in 1967, heralded the era of supertanker spillages and their environmental consequences. Aggressive attempts to clean up stranded oil, with the use of chemical dispersants or by removing sediment from the marsh surface, proved to be more damaging in the long term than allowing natural recovery (Baker 1999). Twelve years after the Amoco Cádiz disaster, heavily oiled marshes on the Brittany coast that had received no cleanup showed good recovery, but those that had been cleaned by sediment removal were extensively altered (Gilfillan et al. 1995). Unfortunately, aesthetic considerations and the economics of tourism may outweigh ecological priorities. The more recent wreck of the Prestige off the Iberian Peninsula in 2002 (Andrade et al. 2004) undoubtedly caused Spain’s worst environmental disaster to date.

RELEASE OF RADIONUCLIDES Nuclear reprocessing facilities, such as Sellafield in northwest England and La Hague in northern France, are allowed low-level discharges of radionuclides with long halflives into coastal waters. These radionuclides inevitably are trapped in the fine-grained deposition of salt marshes, which thus retain a

TABLE 16.2 Major tanker oil spills that have polluted salt marshes on the coast of Europe

Year

Tanker

Marshes Affected

Size (Tonnes) Pollutant

Source

1967

Chryssi P. Goulandris Torrey Canyon Amoco Cadiz Tanio Sea Empress Erika Prestige

Milford Haven, southwest Wales

250–500

1, 2

Brittany, France; southwest England Brittany, France Brittany, France Milford Haven, southwest Wales Loire Atlantique and Vendee, France Galicia, northwest Spain

132,000 223,000 26,000 72,000 20,000 63,000

1967 1978 1980 1996 1999 2002

Arabian light crude Kuwait crude Light crude Fuel oil North Sea crude Heavy fuel oil Fuel oil

1, 2 2, 3 4, 5 5, 6 5, 7 8

SOURCES: 1, Baker (1979); 2, Baker (1999); 3, Gilfillan et al. (1995); 4, Berne and Bodennec (1984); 5, http://www.itopf.com/stats.html; 6, Sime and Edwards (1998) ; 7, Poupart and Meudec (2005); 8, Andrade et al. (2004).

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stratigraphic record of the discharge histories (Brown et al. 1999; Cundy et al. 2002). However, there is evidence that the more labile elements can be redistributed within the marsh (Morris, Butterworth, and Livens 2000). After decades of discharge from Sellafield, manmade radionuclides are widely distributed in the marshes around the Irish Sea, and their concentrations are highest in the marshes of Esk and Ribble estuaries, close to the discharge site (Howard, Livens, and Walters 1996; Sanchez et al. 1998). In addition, there has been episodic fallout of radionuclides from atmospheric weapons testing, which peaked in 1963, and the radioactive plume resulting from the Chernobyl disaster, which passed over northwestern Europe in 1986. Indeed, the 137Cs peak from Chernobyl has been used as a marker to estimate recent rates of salt marsh accretion in the Baltic and North seas (Callaway et al. 1996). While this accumulation of toxic and radioactive substances might be a cause for concern, the long-term ecological consequences for salt marshes, if any, are unknown.

CONSERVATION NEEDS AND PRIORITIES The continuing direct and indirect threats to salt marshes from human activity highlight the need for both their conservation and restoration. The conservation of European salt marshes, like other habitats, is co-coordinated using a common framework that encompasses all member states of the European Union. Currently, there are twenty-seven member states. The Habitats Directive (Council Directive 92/43/EEC on the conservation of natural habitats and of wild fauna and flora) aims to (1) contribute to ensuring biodiversity through conservation of natural habitats and species of wild fauna and flora of European interest; and (2) maintain or restore their favorable conservation status. One of the more important requirements of the Habitats Directive is the selection, designation, and protection of a network of Special Areas of

Conservation (SACs). Together with Special Protection Areas (SPAs) designated under the earlier Birds Directive (79/409/EEC), they constitute the Natura 2000 network of sites. The list of habitat types considered of European conservation importance (annex 1 to the Habitats Directive) includes six categories that span the whole range of coastal or inland salt marshes (table 16.3). Some of the annex 1 habitat types are designated “priority” because they are considered to be particularly vulnerable and are mainly, or exclusively, found within the European Union. In relation to salt marshes, however, this applies only to inland marshes and inland salt and gypsum steppes (table 16.3). A large number of salt marsh Natura 2000 sites has already been designated, although few in the more threatened habitat types (table 16.4). In addition, countries maintain many salt marsh nature reserves of national or international importance that have not been designated as SACs. The requirement under the Habitats Directive to maintain favorable conservation status means that the natural range and area covered by a habitat type cannot be allowed to decline. This is currently one of the drivers of salt marsh regeneration and restoration on land claim sites, using managed coastal realignment: in Britain alone, it is estimated that 2,100 hectares will need to be regenerated to compensate for the predicted losses by 2014. Whether such replacement marsh can also meet the quality criteria for favorable status is still far from clear.

CONCLUSION European salt marshes are among the most intensively studied and well known in the world, a feature they share with those of North America. They differ from marshes in the New World and Australasia, however, in having a much longer, sustained history of human impact. There are virtually no genuinely natural salt marshes in Europe, and human activity has affected both

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TABLE 16.3 European salt marsh habitats designated in Annex I to the Habitats Directive of the European Commission. European Commission (2003)

natura 2000 code

habitat type and subtypes

characteristic plant species

Atlantic and Continental Salt Marshes and Salt Meadows 1310

Salicornia and other annuals colonizing mud and sand. Subtypes: Glasswort stands (Thero-Salicornietalia)

Salicornia spp., Microcnemum coralloides, Suaeda maritima

Mediterranean halo-nitrophilous pioneer communities (Frankenion pulverentulae)

Frankenia pulverentula, Suaeda splendens, Salsola soda, Cressa cretica, Parapholis incurva, P. strigosa, Hordeum marinum, Sphenopus divaricatus

Atlantic sea-pearlwort communities Sagina maritima, S. nodosa, Cochlearia (Saginion maritimae) danica, Gentiana littorale, Bupleurum tenuissimum Central Eurasian crypsoid communities 1320

Crypsis spp., Cyperus pannonicus, Spergularia media, S. marina, Salicornia spp., Lepidium latifolium, Chenopodium spp., Atriplex spp.

Spartina swards (Spartinion maritimae), Subtypes: Flat-leaved cordgrass swards

Spartina maritima, S. alterniflora

Rush-leaved cordgrass swards

Spartina densiflora

1330

Atlantic salt meadows (Glauco-Puccinellietalia)

Puccinellia maritima, Atriplex portulacoides, A. pedunculata, Aster tripolium, Armeria maritima, Glaux maritima, Plantago maritima, Frankenia laevis, Seriphidium maritimum, Festuca rubra, Agrostis stolonifera, Juncus gerardii, Carex extensa, Blysmus rufus, Eleocharis spp., Spergularia marina, Puccinellia distans, P. fasciculata, P. retroflexa, Triglochin maritima, Potentilla anserina, Elyrigia atherica, Atriplex littoralis, A. hastata, Beta maritima, Matricaria maritima

1340

Inland salt meadowsa

Aster tripolium, Atriplex hastata, A. pedunculata, Elytrigia atherica, Juncus gerardii, Plantago maritima, Puccinellia distans, Salicornia spp., Spergularia salina, Suaeda maritima, Triglochin maritima

Mediterranean and Thermo-Atlantic Salt Marshes and Salt Meadows 1410

Mediterranean salt meadows (Juncetalia maritimi). Subtypes: Tall rush salt marshes dominated by Juncus maritimus/J. acutus

Juncus maritimus, J. acutus, Carex extensa, Aster tripolium, Plantago cornuti, Scorzonera parviflora

Short rush, sedge, and clover salt marshes (Juncion maritimi) and humid meadows behind the littoral, rich in annual species and Fabaceae

Hordeum nodosum, H. maritimum, Trifolum squamosum, T. michelianum, Alopecurus bulbosus, Carex divisa, Ranunculus ophioglossifolius, Linum maritimum

Mediterranean halo-psammophile meadows

Plantago crassifolia, Blackstonia perfoliata, Centaurium tenuiflorum, Orchis coriophora ssp. fragrans

TABLE 16.3 (Continued)

natura 2000 code

habitat type and subtypes

characteristic plant species

Mediterranean and Thermo-Atlantic Salt Marshes and Salt Meadows Iberian salt meadows

Puccinellia fasciculata, Aeleuropus littoralis, Juncus gerardii

Halophilous marshes along the coast and coastal lagoons

Puccinellia festuciformis

Humid halophilous moors with a shrubby stratum dominated by Artemisia coerulescens

Artemisia coerulescens

Cyprus subtypes

Eleocharis palustris, Puccinellia gigantea, Arthrocnemum macrostachyum, Aeleuropus littoralis, Centaurium spicatum, Cressa cretica, Crypsis factorofskyi

1420

Mediterranean and thermo-Atlantic halophilous scrubs (Sarcocornetea fruticosae)

Atriplex portulacoides, Inula crithmoides, Suaeda vera, Sarcocornia perennis, S. fruticosa, Arthrocnemum macrostachyum, Halocnemum strobilaceum, Limonium virgatum, L. diffusum, L. ferulaceum, L. densissimum, L. girardianum. L. bellidifolium, L. gmelinii, Aeleuropus littoralis, Aster tripolium, Limoniastrum monopetalum, Artemisia gallica

1430

Halo-nitrophilous scrubs (Pegano-Salsoletea)

Perganum harmala, Artemisia herba-alba, Lycium intricatum, Capparis ovata, Salsola vermiculata, S. genistoides, S. oppositifolia, Suaeda pruinosa, Atriplex halimus, A. glauca, Camphorosma monspeliaca, Haloxylum articulatum

Salt and Gypsum Inland Steppes 1510

Mediterranean salt steppes (Limonietalia)a

Halopeplis amplexicaule, Hymenolobus procumbens, Limonium spp., Lygeum spartum, Microcnemum coralloides, Salicornia patula, Senecio auricula, Sphenopus divaricatus

1520

Iberian gypsum steppes (Gypsophiletalia)a

Centaurea hyssopifolia, Gypsophila hispanica, G. struthium, Helianthemum squamatum, Herniaria fruticosa, Lepidium subulatum, Ononis tridentata, Reseda stricta, Teucrium libanitis

1530

Pannonic salt steppes and salt marshesa

Artemisia santonicum, Suaeda pannonica, Lepidium crassifolium, Puccinellia peisonis, Aster tripolium, Salicornia prostrata, Camphorosma annua, Plantago tenuiflora, Juncus gerardii, Plantago maritima, Cyperus pannonicus, Pholiurus pannonicus, Festuca pseudovina

NOTE: This supranational legislative instrument provides for a common framework for the conservation of plant and animal species and their natural habitats within the 25 member states of the European Union. This classification is the basis for the creation of a European network of Special Areas of Conservation, a major constituent of Natura 2000. a

Indicates a priority habitat (see text).

TABLE 16.4 Numbers of European salt marsh sites protected under the Natura 2000 network recorded in the European Nature Information System Database (EUNIS 2006)

Natura 2000 Code 1310 1320 1330 1340 1410 1420 1430 1510 1520 1530

Habitat Type and Subtypes

Number of Sites

Salicornia and other annuals colonizing mud and sand Spartina swards (Spartinion maritimae) Atlantic salt meadows (Glauco-Puccinellietalia) Inland salt meadows Mediterranean salt meadows (Juncetalia maritimi) Mediterranean and thermo-Atlantic halophilous scrubs (Sarcocornetea fruticosae) Halo-nitrophilous scrubs (Pegano-Salsoletea) Mediterranean salt steppes (Limonietalia) Iberian gypsum steppes (Gypsophiletalia) Pannonic salt steppes and salt marshes

632 599 625 30 629 610 15 14 6 1

NOTE: See table 16.3 for details of habitat classification.

the quality and quantity of seminatural habitat. Historically, the predominant influence on quality has been grazing by livestock. Recently, increased intensification of agriculture and overproduction in Europe have led to widespread abandonment of land more marginal for productivity. The cessation of agricultural practices on salt marshes will eventually result in the dominance of a few plant species with subsequent loss of biodiversity. This trend can only be exacerbated by progressive eutrophication. In many places we now need to consider the use of grazing in the repertoire of management techniques to restore integrity and biodiversity. Recent changes in agricultural practices in North America have had devastating, if less direct, repercussions on the salt marshes of the Arctic, by increasing goose populations reaching the summer breeding grounds (see chap. 5, this volume). The pressure on space and natural resources exerted by dense human populations over much of Europe has been the main determinant of quantity of salt marsh. Until recently, the overwhelmingly most important impact was loss of marsh area to land claim. The imperative to occupy tidal lands has been much less pressing in North America, where exploitation for fisheries and hunting has also helped their continued existence. Again, only in the past decade 330

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has the historical trend of land claim begun to be reversed, as more rapidly rising sea levels render current sea defenses uneconomic to maintain in some of the more densely populated areas of Europe. As we have seen, “managed realignment” (retreat) of the coast brings opportunities for the reactivation of long-lost marshes and even the creation of new ones. It is also fraught with potential threats, as changed sedimentary environments can promote erosion, and the erection of new, landward sea defenses may still lead to “coastal squeeze.” Outcomes well short of those desired are possible, including conversion to low-lying intertidal mudflats and a net loss of salt marsh. In general, coastal subsidence and rising sea levels may be less of an issue for North American salt marshes, except in key areas such as the Mississippi Delta. Salt marshes suffer many modifications and threats that are similar throughout the world. They have shared with the oceans the dubious role of being deliberately or unwittingly used as dumps for all kinds of waste. They have often been in the way of dredged channels for industrial ports or just “reclaimed” for something perceived as more useful. Until recently, society has been regrettably slow to recognize the value of their landscapes, their importance for biodiversity, the ecosystem functions they can

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17

Human Impacts and Threats to the Conservation of South American Salt Marshes Cesar S. B. Costa, Oscar O. Iribarne, and Jose M. Farina Except for pre-Columbian civilizations established in the Andes high plains, most indigenous Indian tribes of South America were hunter-gatherers, with rudimentary agriculture that had little structural impact on coastal salt marshes (e.g., construction of small ponds for salt production). Since the beginning of the European colonization of South America in the fifteenth century, most of the main cities have been established in bays, coastal lagoons, and estuaries at the expense of coastal wetlands. During the nineteenth century, after the long struggle for independence from Spain and Portugal, agriculture and ranching were the main drivers of habitat modification in the coastal regions of recently established countries. In the twentieth century, the expansion of agribusiness, forestry, and mining has supported the economic development of many South American countries. After World War II, the population explosion, rapid industrialization, and increased dumping of solid and liquid waste in waterways have become the primary drivers of human impacts on coastal marshes. Although there is significant geographic variation in the intensity and distribution of these human influences across South America, our activities have either directly (by habitat use) and/or indirectly (by pollution, basin modification, and changes of soil use and hydrology)

affected the structure and function of South American salt marshes. Intertidal marsh communities extend along the coast of the continent from Cartagena Bay (Colombia, 10⬚24⬘ N) to the Magallanes region (southern Chile, 55⬚ S). South American salt marshes show marked geographic differences in important physical forcing factors, plant composition, and food web structure. This variation means that we must consider site-specific forces controlling community organization to (1) properly characterize and evaluate the impacts of human disturbances and (2) assess the resilience and recovery needs of humanimpacted salt marshes. To avoid conversion to environmentally degraded dumping places or human development sites, tropical and temperate salt marshes must be geographically identified, mapped, and protected by federal and local laws. In addition, public education and internationally coordinated actions should also be used to preserve the integrity and ecological character of South American salt marshes.

CHARACTERIZATION OF SOUTH AMERICAN SALT MARSHES Unlike for North America, few published wetland inventories provide information on the extent of salt marshes in South America, and studies on 337

the dynamics of these systems are entirely lacking. Information that is available indicates that South American marsh vegetation is biogeographically distinct from other areas around the world (West 1977; Costa and Davy 1992; San Martín et al. 1992; Canevari et al. 1997; Adam 2002; Isacch et al. 2006). Plant composition and dominant species of tropical, warm temperate and cold temperate salt marshes are markedly different. Although South American salt marshes function as reproductive and feeding grounds for many migratory and resident birds and hot spots of avifauna diversity (Antas 1994; Chesser 1994), except for the southwest Atlantic (but see Iribarne, Bortolus, and Botto 1997; Iribarne 2001; Silliman and Bortolus 2003; Botto et al. 2005), little is known about other salt marsh animals of South America. TROPICAL SALT MARSHES

In areas not colonized by mangrove propagules or where mangrove growth is inhibited (Tomlinson 1986), two kinds of well-developed salt marsh are found at tropical and subtropical latitudes. The first are narrow (a few meters wide) fringe marshes characterized by monospecific stands of grasses on the seaward edge of mangrove forests. Fringe marshes occur on coastlines where the intertidal zone is protected from wave action and there is ample sediment supply to allow the accumulation of marsh soils (e.g., river mouths). Except for the semiarid coast of northeastern Brazil where Sporobolus virginicus fringes the low intertidal zone (C. S. B. Costa, personal observation), the cordgrass Spartina alterniflora commonly dominates these marshes. At these tropical latitudes, both grasses also form stands within mangroves in the high intertidal zone (Adaime 1978; Wakabara et al. 1996; Flynn, Wakabara, and Tararam 1998; Netto and Lana 1997, 1999). Most conspicuous are large stands of S. alterniflora on the coast of Guyana and Suriname (8⬚ N; West 1977) and in several of the states of Brazil: Maranhão (1⬚ S; Souza-Filho and Paradella 2002; Cohen and Lara 2003), Espírito Santo (20⬚ S; Carmo 1987), São Paulo (25⬚ S; Adaime 1978), and Santa 338

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Catarina (27⬚ S; Soriano-Sierra 1990; Panitz 1992). Mangrove fringe marshes of S. alterniflora support higher densities and diversities of macrobenthic invertebrates (Lana and Guiss 1991; Netto and Lana 1997, 1999; Flynn et al. 1998) and fishes (Wakabara et al. 1996) than adjacent unvegetated intertidal areas. The second type of tropical salt marsh is vegetated salt flats colonized by salt-resistant herbaceous communities (referred to as salinas, salitrales, salares, or apicum) that occur in areas flooded only twice a month by spring tides (Tomlinson 1986; Rutzer and Feller 1987; Menezes and Peixoto 2000; Souza-Filho and Paradella 2002; Cohen and Lara 2003). They are common in the Atlantic dry tropics (5⬚ S to 15⬚ S) but also to the north and south where high evaporation coupled with relatively infrequent flooding create conditions favoring the formation of hypersaline soils (ranging between 50 and 140 parts per thousand; Souza-Filho and Paradella 2002, 2003; Cohen and Lara 2003; Cohen et al. 2004). Plant assemblages on salt flats have been related to particular flooding frequencies (Cohen and Lara 2003; Cohen et al. 2004) and rainfall evaporation patterns. The forbs Batis maritima, Sesuvium portulacastrum, and Blutaparon vermiculare and the grass S. virginicus are common in these marshes from Guyana (Martyn 1934; Mears 1982), to southeastern Brazil (Mears 1982; Lacerda and Hay 1982; Menezes and Peixoto 2000). B. vermiculare has also been recorded near Guayaquil in Ecuador (Mears 1982). Most studies indicate that these marshes harbor lower macrobenthic species richness than other tropical sites. Kober (2004) found fifty-five different benthic taxa associated with open mudflats of large intertidal creeks in northeastern Brazil. Although polychaetes represented the most diverse (twenty taxa) and numerically abundant group (59 percent of up to 677 individuals per square meter found; Kober 2004), seven decapod crustaceans and one gastropod accounted for more than 95 percent of total epifaunal biomass of open mudflats (Koch and Wolff 2002; Kober 2004). The most abundant species on these mudflats

is the fiddler crab Uca maracoani. Potentially due to its tolerance of high temperatures (Koch, Wolff, and Diele 2005), this species dominates the mid- and low intertidal zones. Salt flat benthic communities also support more than 127 species of birds, with densities of 19 per hectare (Kober 2004). In addition to these two kinds of coastal marsh, in the highland (Altiplano) basin, that extends for approximately a thousand kilometers from southern Peru across Bolivia and into Argentina and Chile, continental salt marshes are found associated with large salt flats, salt lakes, and lagoons (salares). In northern Argentina (27⬚ S to 30⬚ S; provinces of Santa Fé, Cordoba, and Santiago del Estero), short grasses dominate the plant communities of continental salt marshes (Carnevale et al. 1987). The most abundant halophytic species of these prairies are Paspalum vaginatum, P. dilatatum, Distichlis spicata, and Stipa hyaline, although S. densiflora is occasionally present. Surrounded by semiarid lands, these marshes are biodiversity hot spots and key areas for migrant birds (Antas 1994; Chesser 1994; Bucher and Chani 1997). TEMPERATE ATLANTIC MARSHES

From 31⬚ S to 43⬚ S latitude, over two thousand square kilometers of salt marshes occur along the hot temperate coasts of southern Brazil, Uruguay, and northeastern Argentina (Isacch et al. 2006). Along this coast, a combined gradient of decreasing mean annual rainfall (from 1,200 to 196 millimeters) and increasing mean tidal range (from less than 0.5 to more than 2.5 meters) determines marked differences in marsh vegetation, from brackish sedges and rushes in the north, to halophytic meadows of Spartina and Sarcocornia in the south (West 1977; Costa and Davy 1992; Canevari et al. 1997). In euryhaline areas of the southwest Atlantic, low marshes are dominated by monospecific stands of S. alterniflora (Cagnoni and Faggi 1993; Costa 1997; Muniz and Venturini 2001; Isacch et al. 2006), Spartina densiflora (Iribarne et al. 1997; Bortolus and Iribarne 1999; Clara and Maneyro 1999; Bortolus,

Schwindt, and Iribarne 2002; Isacch et al. 2006) and/or Spartina ⫻ longispicula (believed to be a hybrid between S. alterniflora and S. densiflora; Mobberley 1956). When both S. alterniflora and S. densiflora are present, the former usually occupies most of the low marsh, while S. densiflora occupies the upper, drier levels of the marsh (Cagnoni and Faggi 1993; Costa, Marangoni, and Azevedo 2003). In the absence of S. alterniflora, Sarcocornia perennis become an important salt marsh pioneer (Perillo and Iribarne 2003, 2004). High and low salt marshes of the southwest Atlantic are grazed by rabbits (e.g., Oryctolagus spp.) and by large (Myocastor coypus and Hydrochoerus hydrochaeris) and small (Scapteromys tumidus, Holochilus brasiliensis, Oligoryzomys flavescens, O. delticola, Akodon azarae) wetland rodents (Clara and Maneyro 1999; Bo et al. 2001). Main predators for these rodents are the thick-tailed opossum (Lutreolina crassicaudata) and the crab-eating fox (Cerdocyon thous). Among coastal birds, oystercatchers, seagull, and terns are important marsh predators of polychaetes (Nereidae, Nephtyidae, and Capitellidae), pelecypods, and crustaceans inhabiting mudflats and tidal pools (Botto et al. 1998; Iribarne and Martinez 1999; Delhey et al. 2001). During low tides, marine and coastal birds feed on marsh invertebrates; during high tides, they rest on higher ground (Blanco 1998), becoming vulnerable to terrestrial predators. The great abundance of the intertidal burrowing crab Chasmagnathus granulatus is the most distinctive characteristic of southwest Atlantic marshes (Iribarne et al. 1997; Silliman and Bortolus 2003; Rosa and Bemvenuti 2004). These crabs (about three to four centimeters in carapace width) inhabit soft nonvegetated bottoms of the upper beaches vegetated by S. densiflora and S. perennis (Iribarne et al. 1997). Field observations and gut content analysis indicate that C. granulatus is a deposit feeder in soft bottom sediments and a herbivore-detritivore in salt marshes (Iribarne et al. 1997; Bortolus and Iribarne 1999). Regardless of habitat, however, stable isotopic analyses show that Spartina is human impacts in south america

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their primary food source (Botto et al. 2005). As in North America (on the U.S. Atlantic coast: Silliman and Zieman 2001; Silliman and Bertness 2002; Silliman et al. 2005), on the southwestern Atlantic coast of South America, there are invertebrate grazers (crabs) that exert strong control on marsh plant growth (Argentina: Bortolus and Iribarne 1999; Silliman and Bortolus 2003; Brazil: Costa et al. 2003). Herbivory by C. granulatus (consumption of young shoots) often decreases the aerial biomass of S. densiflora by more than 80 percent (Bortolus and Iribarne 1999) and could limit the zonation of S. alterniflora in bare areas (Costa et al. 2003). Given the widespread distribution of C. granulatus, and its ability to suppress marsh plant growth and distribution (Bortolus and Iribarne 1999; Costa et al. 2003; Silliman and Bortolus 2003), C. granulatus could play an important role in determining marsh plant community structure at regional spatial scales. Cold temperate salt marshes spread from middle to high latitudes along the South American coast, where climate changes dramatically. Between 36⬚ S and 56⬚ S, due to the strong influence of the Falklands cold current (Costa and Davy 1992) and the effect of cool frontal systems originating in Antarctica (Acha et al. 2003), annual average air temperatures on the Atlantic coast of Argentina decrease from 12⬚C to 3⬚C. Following this trend, Spartina species become less abundant (Isacch et al. 2006). At the southern extreme, in Tierra del Fuego, marshes are characterized by extensive mud/sandy flats determined by the macrotidal regime with amplitude of more than fourteen meters (Collantes and Faggi 1999; Bujalesky, Coronato, and Isla 2001). In this zone, low marshes are dominated by S. perennis (Bianchiotto et al. 2003), while high marshes are covered by Puccinellia spp. and pastures (P. magellanica, P. biflora, P. glaucescens) with the presence of Juncus acutus and/or several forbs (Frankenia chubutensis, Limonium brasiliense, Suaeda argentinensis, and Armeria maritima), as well as plants of the nearby halophilous steppes of central Patagonia 340

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(Ledidophyllum cupressiforme, Atriplex lampa, Hordeum halophilum, Elymus sp., and Poa spp.; Dollenz 1977; Collantes and Faggi 1999; Bujalesky et al. 2001). In the Patagonian region, Lama guanicoe graze several high marsh species of short grasses (particularly Puccinellia spp.), and the gray fox (Dusicyon grideus) is the top predator in this area (Marcenaro et al. 1984; Bonino et al. 1991). In low marshes, birds are the top predators, although sea lions, dolphins, and penguins occasionally visit these areas to eat crustaceans and fishes (Collantes, Anchorena, and Koremblit 1989). TEMPERATE PACIFIC MARSHES

Most of the Pacific coast of South America is under the influence of the Humbolt cold current, and the steep topography (rocky shores with cliffs forty to two hundred meters high) restricts the presence of salt marshes to the mouths of small-flow rivers and bays (West 1977; Haloua et al. 1999; Bujalesky et al. 2001), especially bordering the desert areas of southern Peru and northern Chile (6⬚ S to 28⬚ S). Because of the absence of rain, some of the rivers that reach the sea have sand barriers blocking their direct entry into the ocean that opens only occasionally. High evaporation typically induces hypersaline conditions in the surrounding deltaic soils. These environments are dominated by Sarcocornia fruticosa (≈ S. perennis) on the low intertidal level and by Distichlis spicata and occasionally Scirpus californicus on higher ground (Arana and Salinas 2003; San Martín et al. 2001). Below 28⬚ S, rainfall increases and S. densiflora increases in representation, whereas S. fruticosa decreases (Ramírez and Añazco 1982; San Martín et al. 1992; Hauenstein et al. 2002). Less saline marsh areas tend to be occupied by Scirpus americanus, Typha angustifolia, and the South African invader Cotula coronopifolia (San Martín et al. 1992, 2001). In marshes associated with sand plains and coastal lagoons, the Chilean flamingo (Phoenicopterus chilensis), austral oystercatchers (Haematopus ater), and imperial cormorants (Phalacrocorax atriceps) are common

(Collantes et al. 1989; Collantes and Faggi 1999). Pacific marshes are also visited by migratory shorebirds, particularly during the spring and summer. Relatively little is know about the impact of permanent and migratory birds on the invertebrate communities of temperate Pacific marshes of South America. However, it is worth mentioning that due to the low incidence of protected shores, in contrast to Atlantic marshes, in most Pacific marshes marine invertebrates are practically absent (J. M. Fariña, personal observation).

HUMAN-INDUCED THREATS TO SOUTH AMERICAN SALT MARSHES LANDFILL

Although marshes were hunting grounds and provide several herbs for the traditional medicine of Amerindians, European colonists viewed South American marshes as wastelands: unproductive and unhealthy sites that were breading grounds for deadly diseases (e.g., yellow fever, malaria, etc.). Given such low status, destroying wetlands via landfill was seen as a positive action. Over the past four hundred years, the filling of intertidal and shallow water flats in bays, coastal lagoons, and estuaries for agricultural, livestock, port, residential, and

industrial developments has decreased or destroyed vital salt marsh habitats along the South American coast (fig. 17.1). Apart from the examination of historical records (maps, reports, etc.), which might allow quantification of marsh losses by landfill, there is no estimate of the absolute scale of this impact in South America. Seeliger and Costa (2002) reported that between 1909 and 1914, during the expansion of the Rio Grande port (in southern Brazil), 8,776,000 cubic meters of dredge material were generated. This material served for the construction of islands and was deposited along the lower Patos Lagoon estuary margins. Indeed, during the nineteenth and twentieth centuries, landfill destroyed about 10 percent of Patos Lagoon salt marshes (fig. 17.1). Similar impacts can be seen in the Bahia Blanca Estuary (Argentina), where a large portion of the S. perennis marsh is being used for landfill (O. Iribarne, personal observation). The loss of tropical and temperate salt marshes to landfill likely further contributes to the continuing decrease of landings of both fully or overexploited estuarine fishes. Salt marsh habitats provide shelter and food for commercially important crustacean and fish species, which represent almost 65 percent of the mean annual artisan catch (21,500 tons) in southern Brazil (Costa et al. 1997).

FIGURE 17.1 Land reclamation for agriculture, road construction, and aquaculture in a salt marsh area of southern Brazil.

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FIGURE 17.2 Crabs eating Spartina at Mar Chiquita salt marsh in Argentina.

Some of the coastal lagoons in Uruguay and Argentina are semiclosed by the construction of bridges, which decrease saltwater input (damping the tidal wave) and freshwater discharge (Isla and Gaido 2001). This is the main environmental disturbance at the Mar Chiquita coastal lagoon, Argentina (Iribarne 2001), which is affecting the distribution of salt marshes and increasing the invasion of brackish plant species (Iribarne et al. 2001). DIVERSION OF FRESHWATER

Diversion of freshwater for energy generation, irrigation, industry, and construction of coastal highways is becoming an issue in South America and already affects major coastal watersheds. Water withdrawal and diversion contribute to increased soil salinity of mangroves and landward salt intrusion in the Magdalena River delta (Colombia; Botero 2002) and in northeastern Brazilian estuaries (Marins et al. 2002). For example, during the twenty-first century, decreased freshwater runoff from the Jagaribe River (northeastern Brazil) during the rainy season (from 200 to 20 m3 s⫺1) has created extensive new saline environments far from the mouth of the river that were subsequently colonized by herbaceous halophytes and mangrove trees (Marins et al. 2002). Between 1969 and 1997, the population and the area of rice cultivation increased in the Patos Lagoon watershed of southern Brazil by more than 37 percent and 120 percent, respectively, 342

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and today as much as 13 percent of natural freshwater runoff (≈ 9.8 cubic kilometers) may be diverted during a drought year (Seeliger and Costa 2002). Ever-decreasing freshwater discharge into this estuary modifies seasonal variation in flow rates that may be essential for flushing and maintaining the balance of salt and nutrients in the estuary (Seeliger and Costa 2003), which are key determinants of structure in salt marsh plant communities (Costa 1997). In southern Chile, the expansion of the food industry (Universidad Austral de Chile 2005) and forestry (Hauenstein et al. 2002) have increased freshwater demand and may also be changing salinity in marsh soils as well as the biogeochemical cycles of some elements. SALT FLAT CREATION AND DESTRUCTION IN THE TROPICS

The utilization of tropical salt marshes, coastal salt flats, or the construction of ponds in the intertidal for salt production spread throughout the Colombian Caribbean, following artisan methods of the Guajiro Indians (Alvaréz-Leon 1993). During the twentieth century, this practice on an industrial scale resulted in conversion of thousands of hectares of mangrove into salt flats along the arid northeastern coast of Brazil (Kjerfve and Lacerda 1993). At the beginning of the 1990s, conversion was totally prohibited, and several salt companies went bankrupt. Several of the abandoned saltpans were colonized by herbaceous halophytes and become permanent features of the upper intertidal zone. Many, however, have not recovered because soil salinities remain too high (over one hundred parts per thousand). Cutting of mangrove trees for coastal development can lead to elevated soil salinities in the intertidal zone, at times resulting in salty bare areas or replacement of the forest by an herbdominated salt flat. Degradation of the mangrove community in Rio de Janeiro (Brazil) may produce a persistent herbaceous community dominated by Sesuvium portulacastrum and Blutaparon vermiculare (Lacerda and Hay 1982). Batis maritima is also observed in cleared

mangrove areas, both toward the seaward margin and in the upper, midlittoral zone (C. S. B. Costa, personal observation). Additionally, the cosmopolitan fern Acrostichum aureum and its American relative Acrostichum danaefolium may also become dominant on disturbed mangroves (Tomlinson 1986; Lebigre 1999; Soares 1999). These changes could be related to intense sediment deposition that buries mangrove seedlings and elevates the topography (Soares 1999). Human destruction of mangroves and subsequent conversion into salt flats may also benefit organisms other than herbaceous halophytes. Salt flats are important to migratory shorebirds such as plovers and sandpipers and migratory waterfowl such as blue-winged teal and several waders (Masero et al. 2000). Although salt flats have benefited from human activities in the past, these habitats are now the main targets for conversion to shrimp farming. Extensive salt flats (salitrales) were found naturally on the Pacific coast of Ecuador, particularly south of Bahia de Caraquez, associated with an extremely dry climate (less than two hundred millimeters of rainfall per year) and the macrotidal regime of the Gulf of Guayaquil. Originally, and until about 1969, more than ninety-two thousand hectares of salt flats were found along the Ecuadorean coast. By 1991, Ecuador had lost nearly forty-three thousand hectares of salt flats due to the construction of ponds for shrimp farming (Botero 1993). More recently, on the north and northeast coast of Brazil, the shrimp farm industry has destroyed thousands of hectares of tropical salt marshes associated with salt flats that were converted to shrimp farms. The low price of salt marsh land is also very attractive for the fastgrowing tourism industry and new resorts. LIVESTOCK GRAZING

Large-scale utilization of South American salt marshes for grazing livestock has existed since European colonization of southern Brazil, Uruguay, and northeastern Argentina in the seventeenth century. Along 1,500 kilometers of coast, the small tidal range, gentle marsh topog-

raphy, and the shallow creeks facilitate access by cattle and horses, allowing trampling and easy access to succulent plants (Seeliger and Costa 1997, 2002). This grazing can be a determining factor for the short-vegetation physiognomy of several southwest Atlantic marshes. Cattle and sheep farming and intentional or accidental burning open the canopy of the dominant Spartina and Juncus species and allow secondary plant species invasion. Salt-tolerant short grasses such as Paspalum distichum and P. vaginatum are particularly abundant in salt marshes strongly affected by cattle and horse grazing (Carnavale et al. 1987; Costa et al. 1997). Additionally, the absence of a tall canopy can enhance evaporation and salinization of the sediment during low tide, decreasing the chances of survival of the intertidal crab Chasmagnathus granulatus in high marsh areas (Bortolus et al. 2002). With population densities up to one hundred burrows per square meter, these crabs graze Spartina shoots (Silliman and Bortolus 2003; Costa et al. 2003), strongly prey on infauna (Iribarne et al. 1997), and are important forces regulating detritus fragmentation and remobilization of sediment and belowground macrophyte biomass (Botto and Iribarne 2000; Botto et al. 2005). Thus, reduction of crab populations due to grazing may influence the recycling of organic matter in the marshes (fig. 17.2). The grazing livestock agribusiness also threatens the food sources and habitats of other native marsh herbivores and their main predators. Traditional ranching reduces the cover of high and low salt marshes of the La Plata River that otherwise are grazed by rabbits and by large and small wetland rodents (Clara and Maneyro 1999). Decreases in herbivore populations in turn directly affect the populations of predators (thick-tailed opossum and crab-eating fox). Sheep grazing prevails from north to south along the Patagonian coast and constitutes an important agent of erosion (Gaiero et al. 2002). In this region, sheep grazing indirectly threatens both guanacos, Lama guanicoe, which feed on several high marsh short grasses, and their main predator, the gray fox (Bonino 1991).

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There is little information concerning the importance of traditional grazing on bird populations that live in very important conservation areas such as Lagoa do Peixe (southern Brazil) and Mar Chiquita Biosphere Reserve (Argentina). Grazed marshes are resting sites of migrant birds (Botto et al. 1998; Clara and Maneyro 1999; Iribarne and Martinez 1999; Isacch and Martinez 2001). They support large populations of permanent residents (Vooren 1997; Clara and Maneyro 1999; Collantes and Faggi 1999; Yorio et al. 2001) and can periodically benefit both sea- and coastal birds. Since tall vegetation seems to prevail after the cessation of grazing (Carnavale et al. 1987; Costa et al. 1997), the increasing efforts of environmental agencies to eliminate grazing in these conservation areas should be considered with caution. This policy, rather than being guided by basic research, is based only on the orthodox conservation goal: “to return these environments to their natural condition.” Following only this path may compromise essential bird habitat in many regions. Large-scale experiments with removal of cattle livestock (fig. 17.3) are needed to assess impacts on migratory bird populations, as well as native components of the more resident food web (e.g., foxes, invertebrates, fish) (fig. 17.4).

FIGURE 17.3 Cattle exclusions at El Yali salt marsh in Chile reveals conspicuous changes on salt marsh plants species composition and biomass.

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FIRE

The impacts of intentional (for land clearance) and accidental burning of South American salt marshes are described only for southwest Atlantic grasslands. In southern Brazil, accidental burning may cause high mortality of invertebrates and nesting birds, but it opens the canopy of mid-upper marshes dominated by S. densiflora and J. kraussii to invasion by secondary short species (the legume Vigna luteola and the forbs Senecio tweedii and Aster squamatus) and increases rates of nutrient availability and recycling (Costa and Marangoni 2000). In the salt marshes of the Pampas region of Argentina, fire management is used as a tool to improve cattle forage and avoid accidental fires (Isacch and Martinez 2001; Isacch et al. 2004). In comparison with the Argentinean land act, this seems a better approach than marsh conversion, but the structure of burned marshes is highly simplified. Vegetation structure is an important component of avian habitat selection, and the effect of burning is very important for bird diversity. Immediately after burning, areas are colonized by species common in open spaces, but then the recovery of bird diversity and abundance follows the recovery of plant structure. Comparison of plant cover and bird

community of salt marshes of Mar Chiquita (Argentina) before and after extensive burning showed that during the first year after the fire, the canopy of Juncus acutus completely recovered and that the bird community returned to its preburn structure (Isacch et al. 2004). One the other hand, after the same time, marshes dominated by S. densiflora were unable to reach the height of the preburned canopy. The red-capped wren-spinetail (Spartonoica maluroides), a rare bird dependent on tall grass that lives mainly in Spartina marshes, never reached the abundances of preburned marshes. Thus, given that excessive burning reduces habitat for endangered or rare species, prescribed burns were not recommended in this region (Isacch et al. 2001). EUTROPHICATION

Most urban centers along the South American coast either lack or have inadequate sewage treatment facilities, which tend to overflow during heavy rains and discharge raw sewage directly into bays, lagoons, or estuaries (Lacerda, Kjerfve, et al. 2002; Tagliani et al. 2003). Excessive nutrient loading in estuarine zones cause eutrophication, with an increase on primary producer biomass and changes in phytoplankton, phytobenthos, macroalgae, and salt marsh plant species composition. Extreme cases of eutrophication occur at the Guanabara Bay (Rio de Janeiro state, southeastern Brazil) and near the Buenos Aires Province (Lacerda, Kjerfve, et al. 2002). In southern Brazil, estuarine nutrient levels are occasionally elevated due to remobilization of bottom sediments in the lower estuary (Niencheski and Windom 1994), local discharge of industrial and domestic effluents (Almeida, Baungarten, and Rodrigues 1993), and agricultural runoff. Threefold increased concentrations of dissolved nitrate and ammonium in eutrophic bays seem to explain the doubled primary production of the marsh grasses S. alterniflora (it went from 670–823 g dry weight m⫺2 yr⫺1 to 1,707 g dry weight m⫺2 yr⫺1; Peixoto et al. 1997; Seeliger et al. 1997) and S. densiflora (1,390 g to 2,390 g dry weight

m⫺2 yr⫺1; Silva, Pereira, and Dorneles 1993; Peixoto and Costa 2004) in comparison to pristine marshes. These growth increases could cause changes in plant zonation and invasion of eutrophic high marshes by low marsh dominants, as a result of the release of nutrient competition (e.g., Pennings and Bertness 2001; Silliman and Bertness 2005), but no obvious signs of this have yet to be observed in South American marshes. Additionally, an increasing amount of floating organic and industrial debris was observed in the last few decades of the twentieth century (Acha et al. 2003), resulting in extensive drift line deposits. Plastics and plastic bags were the main debris types, and debris deposition suffocates dominant rhizomatous plants growing just above the mean water level and opens the canopy to nitrophilous annual species such as Chenopodium album, Atriplex patula, Atriplex hastata, and Apium graveolens, as well as the perennial grass Eriochloa punctata (Costa et al. 1997). Large deposits of solid waste in degraded marshlands make them very vulnerable to human development activities. Similar functional and structural effects of solid waste disposal and liquid effluents are observed even on the UNESCO Man and the Biosphere Reserve and a Ramsar site (Acha et al. 2003), and on supposed pristine Patagonian river estuaries, such as near the city of San Antonio Oeste (twenty-five thousand inhabitants) (O. Iribarne Pers. Obs.). The fast expansion of intensive shrimp farming on both the Pacific and Atlantic coasts of tropical South America is resulting in the discharge of significant amounts of nutrients into adjacent waterways. Brazilian shrimp farming extended over fifteen thousand hectares and produced ninety thousand tons of shrimps during 2003. However, 90 percent of the tanks were located in the northeastern region of the country, and each ton of produced shrimp released fifty thousand to sixty thousand cubic meters of effluents rich in organic particulate matter, microorganisms, and macronutrients into estuarine and coastal waters (Boeger et al. human impacts in south america

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2005). This pollution caused marked episodes of fish and invertebrate mortality associated with low oxygen concentration in the water but also with the outbreak of viral and bacterial diseases in shrimp farms (Schaeffer-Novelli 2002). Animals of tropical marshes and salt flats also have been affected by eutrophication. Beginning in 1998, massive mortalities of the intertidal crab Ucides cordatus throughout most of northeastern Brazil have been reported associated with symptoms, such as lethargy and poor motor control. This lethargic crab disease (LCD) resulted in reduction of up to 84 percent in collection rates by artisan fishermen, and it was recently associated with an ascomycete fungus (Boeger et al. 2005). METAL AND OIL POLLUTION

Heavy nutrient and metal loads on South American marshes are related to the proximity of “megacities” that can directly discharge effluents in the coastal zone (i.e., Rio de Janeiro and Buenos Aires) or indirectly contribute through the catchments that carry their urban waste (i.e., Caracas and São Paulo; Lacerda, Kjerfve, et al. 2002). Sepetiba Bay in southeastern Brazil is an example of the significance of basin-generated impacts on the coastal zone and how tropical salt marshes act as traps for trace metals. Fluvial input from the Paraíba do Sul river basin, which is diverted to the Sepetiba Bay basin for Rio de Janeiro city’s water supply, brings trace metals to the Sepetiba Bay basin (Lacerda, Marins, and Barcellos 2002). Within the bay, trace metals are transported and mostly associated with suspended matter. High metal concentrations can be found in sediments close to the river mouth (e.g., up to 11, 1,100, and 137 parts per million of cadmium, zinc, and mercury, respectively). Sediment oxidation around S. alterniflora roots, due to air transport through aerenchyma tissue, induces cadmium and zinc precipitation with iron and manganese hydroxides (Lacerda, Freixo, and Coelho 1997) but releases mercury previously precipitated with sulfides under reducing conditions (Marins et al. 1997). 346

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Trace metal concentrations of the estuarine waters that flood Patos Lagoon salt marshes (southern Brazil) are close to background levels reported for estuarine systems elsewhere (Seeliger and Knak 1982; Niencheski and Windom 1994). However, further south, the Bahia Blanca Bay, which has 1,150 square kilometers of tidal plains and salt marshes and estuarine soils affected by industrial effluents, shows high contents of zinc (up to 780 parts per million), copper (up to 278 parts per million), and lead (up to 158 parts per million) (Andrade et al. 2002). In addition, Ferrer et al. (2003) found dissolved concentrations near to the norms for marine and estuarine waters but higher mean values in the suspended matter (copper, 36 parts per million; zinc, 205 parts per million). Their ninety-six-hour, semistatic acute assays (for viability) on recently hatched larvae of Chasmagnathus granulatus demonstrated that zinc (172 parts per million) is more toxic toward larvae than copper (219 parts per million). Thus high metal concentrations in soil and suspended matter may inhibit crab recruitment. On the Patagonian coast of Argentina (38⬚ S to 52⬚ S), there are eight permanent rivers, and Gaiero et al. (2002) showed that, in general, average heavy metal concentrations in their suspended matter are lower than those commonly found in nonpolluted rivers. These authors found high concentrations of copper and zinc (290 and 460 parts per million, respectively) only in the suspended matter of the Negro River, where extensive farming occurs. In southern Peru and northern Chile, human threats to salt marshes mostly emerge from increasing demand for water and habitat for industrial activities and the dumping of copper mine tailings in coastal areas. The mining industry utilizes water from the highlands of the Andes to transport mineral to coastal harbors, where effluents accumulate and high values of porewater copper are found (up to 1,450 parts per million; Lee and Correa 2005). The main approach used to deal with lead in polluted waters has been to pump the contaminated water into tree plantations created in the

desert near the mouth of occasional rivers, in order to evaporate excess effluent. However, the overflow of these systems during stormy or melting periods carries contaminants to neighboring marshes. Chronic petroleum losses during docking operations and large spills are major threats to South American salt marshes. The major spills that have affected salt marshes were the accidents of the Metula in the Magellan strait, the San Jorge near Punta del Este (Uruguay), and the Estrella Pampeana in the access channel of Buenos Aires Harbor. The Bahía Blanca Bay port system (the most important in Argentina) has a petrochemical industrial complex that has dumped hydrocarbon raw materials and refined products in the surrounding marsh areas via atmospheric pollutants and effluents for almost thirty years. The hydrocarbon content in soil samples ranged between 3.3 and 158.3 parts per million, and it was highly correlated with total content of lead (Andrade et al. 2002). Gaiero et al. (2002) also found evidence of lead contamination in the Colorado and Gallegos rivers and associated it with oil extraction activity in their headwaters. A major area of petroleum exploration and oil tanker traffic is located near extensive salt marshes in Tierra del Fuego (Dollenz 1977; Collantes and Faggi 1999). Most of the oil-contaminated marshes show patches of dead plants, with all or most of the aboveground vegetation coated. Heavy crude and refined oils can severely affect marsh plants and animals by coating their roots and respiratory systems, thus preventing gas exchange. The acute effects of light refined oils have not been examined on South American marshes, but both light and heavy oils can become mixed deeper into the substrate via polychaetes tubes or crab burrows, extending the damage both in degree and duration. ORGANOCHLORINE CONTAMINANTS

Although data on the effects of agriculture-derived compounds on estuarine organisms are lacking, the excessive application of pesticides (e.g., malathion and deltamethrin) over vast

areas of agricultural lowlands of the southern Atlantic is likely to contribute, through runoff, to elevated concentrations in the estuaries. Organochlorine contaminants (OC) are not important in Argentinean marshes, but they have been detected. At Mar Chiquita coastal lagoon (Menone et al. 2000, 2004), polychlorinated biphenyl (PCB) concentrations were relatively low. Heptachlor epoxide, dieldrin, endosulfan sulphate, chlordane compounds, DDT and its metabolites, and hexachlorocyclohexanes (HCHs) were the major pesticides detected. Higher concentrations of SOCC were found in sediments of South American cordgrass marsh (S. densiflora) than in mudflats, with heptachlor epoxide being the most abundant OC pesticide in Mar Chiquita lagoon. The total amount of heptachlor epoxide in the underground biomass of South American cordgrass (extending over thirty-two square kilometers) was estimated as 2.4 kilograms (Menone et al. 2000). Thus, South American cordgrass is an important distribution agent of persistent contaminants (particularly sediment-bound OC), and a significant portion of the total budget of these hydrophobic compounds is located in Spartina biomass. Some salt marsh species (i.e., Chasmagnathus granulatus and Uca uruguayensis), which strongly influence energy transfer in submerged and emerged habitats (D’Incao et al. 1992; Botto et al. 2005), appear to be highly susceptible to pesticide concentrations (Ferrero, Gutiérrez, and Cervelini 2001), but this is mainly in warm waters during the summer (Monserrat and Bianchini 1995). Chamagnathus granulatus living in S. densiflora beds showed higher concentrations of OC in their tissue than crabs from neighboring mudflats. Concentration of cyclodienes ⬎ HCHs ⬎ DDTs were found in crabs (Menone et al. 2004). These more water-soluble compounds are excreted by the feces and finally removed by tidal flushing. Thus, it is evident that crabs also play a role in the distribution of sediment-bound OC, and crab beds are modifiers of the dynamics of organic pollutants in southwest Atlantic estuarine areas.

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SEA-LEVEL RISE AND CLIMATE CHANGE

Extensive shore erosion along the northern and northeastern Brazilian coast, as well as the northeastern and eastern coast of Argentina, has been reported. Between 1972 and 1997, there was a net loss of coastal mangroves of 0.8 kilometer per year along 166 kilometers of the elevated areas of Bragança’s coastline (northern Brazil; Cohen and Lara 2003). Meanwhile, mangrove forests have migrated landward, invading 3.4 square kilometers (about 38 percent) of the elevated hypersaline herbaceous flats. These current dynamics of vegetation coverage change seem to be compatible with a long-term trend related to rates of relative sea-level rise along the Brazilian coast. This rise has lead to an increase in inundation frequency, reduction in soil salinity, and the transport of mangrove seeds into more elevated marsh flats (Cohen and Lara 2003) (fig. 17.5). Evidence of erosive processes along the Patos Lagoon margins and the coastline of southern Brazil (Tomazelli and Villwock 1989; Villwock 1994) suggest an estimated rate of sea-level rise between two and three millimeters per year. Lanfredi, Pousa, and D’Onofrio (1998) analyzed the tide height records of the Argentinean coast between Buenos Aires and Puerto Madryn and also found estimates of relative sea-level rise of between 1.4 and 3.5 millimeters per year. Additionally, analysis of historical data (1854–1979) showed a strong secular warming trend of sea surface temperature (SST; up to 1.6⬚C 100 year⫺1) on the southwestern Atlantic coast near southern Brazil (Zavialov, Wainer, and Absy 1999). Elsewhere, a causal relationship between increasing extreme storm frequency and increasing sea surface temperature (SST) has been posited (Emanuel 2005; Webster et al. 2005). On March 25, 2004, the first tropical cyclone of hurricane intensity in the satellite era developed in the South Atlantic and struck the southern coast of Brazil in Santa Catarina State (Hughes and Merckle 2005). Indeed, over Paraguay, southern Brazil, Uruguay, and northeastern Argentina, northeasterly circulation associated with the subtropical Atlantic anticyclone 348

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has increased since 1954 (Mata et al. 2001). If rates of relative sea-level rise and the increase in intensity of cyclones are confirmed for the flat, wave-dominated coast of the southwest Atlantic (Costa, Cordazzo, and Seeliger 1996; Perillo et al. 2005), storm overwash will more frequently affect areas further inland than in the past (Costa et al. 1996). Simulations also show alarming increases in water levels associated with storm surges (Lanfredi et al. 1998) and widespread beach and marsh erosion. Coastal marshes and wetlands are likely to migrate further inland (Costa and Davy 1992), and salinity in upper estuarine areas might increase. Rainfall and the discharges of major rivers have been increasing over the last fifty years in the southeastern Atlantic and southwestern Pacific estuaries, and it is still unknown whether sea-level rise will compensate for the retreat of the salt front and salt marshes to lower estuarine zones. Large-scale phenomena such as El Niño/La Niña–Southern Oscillation (ENSO) are important sources of interannual climatic variability in South America, generating notorious changes in rainfall (Kiladiz and Diaz 1989; Ropelewski and Halpert 1996; Mata et al. 2001). However, increases in the long-term mean rainfall are also being noted (e.g., Argentina: Viglizzo et al. 1995; Hulme and Sheard 1999; Brazil: Genta, Pérez, and Mechoso 1998; Uruguay: Krepper, Garcia, and Jones 2003), although there are only scarce and geographically restricted studies on their community effects (Vilina and Cofre 2000; Jaksic 2001). In central Argentinean pampas, rainfall regimes show a long-term cyclic behavior that during the last four decades has increased over the historic annual mean (Viglizzo et al. 1995, 1997; Lucero and Rozas 2002). Annual rainfall in the coastal part of this region increased from an average of 751 millimeters from 1900 to 1950 (range, 396 to 1,231 millimeters), to an average of 934 millimeters during the 1950–2004 period (range, 588 to 1,826 millimeters; Canepuccia 2005). This region is characterized by an extensive, flat landscape, where high rain frequently causes flooding (Frenguelli 1950; Soriano et al. 1991). These events are

FIGURE 17.4 The combined effects of cattle grazing and drought severely affected salt marsh plants distribution at Peixe Lagoon National Park in Brazil.

likely to modify the availability of wetland habitats and thus habitat diversity. However, given the correlation between freshwater input and dominance of S. alterniflora or S. densiflora (Isacch et al. 2006) this increase in rainfall could promote increased dominance of S. densiflora. INCREASING ULTRAVIOLET RADIATION

The gradual degradation of the ozone layer over Antarctic regions is causing an increase in the transmission of ultraviolet-B (UV-B; 280 to 320 nanometers) radiation and poses a threat to cold and warm temperate latitudes in the Southern Hemisphere (Tarasick et al. 2003). Ozone-poor stratospheric air is spreading to northern Argentina, Uruguay, and southern Brazil and has already caused a 13 to 20 percent decrease of ozone at 30⬚ S during spring (Santee et al. 1995; Kirchhoff et al. 1996). Guarnieri et al. (2004) showed that a 1 percent reduction in ozone in southern Brazil resulted in an increase of about 1 percent in UV-B radiation for most solar zenith angles observed in an annual cycle. Not only have the spring levels of UV-B radiation doubled in the last twenty years, but also the number of days that South America is under the influence of the ozone hole has increased (Rousseaux et al. 1999). Field experiments using UV-B opaque and transparent plastic filters in southern Brazil and

Patagonia showed a relatively small effect of UV-B radiation on the growth of salt marsh plants with basal meristems (Juncus and Spartina) and a marked effect in plants with apical meristems (Sarcocornia) (Bianciotto et al. 2003). The main growth response to full exposure of Sarcocornia plants to ambient UV-B radiation was the reduction of shoot branching (60 to 85 percent) and shoot density (up to 30 percent) of Brazilian and Tierra del Fuego populations, respectively (Costa et al. 2006). A three-year UV-B exclusion field experiment with the Patagonian dominant salt marsh plant, Salicornia ambigua (⫽ Sarcocornia perennis), in Tierra del Fuego (Bianciotto et al. 2003) showed that exposure to natural UV-B significantly reduced plant biomass by 17 percent. Cuticle thickness and the concentration of UV-shielding pigments (absorption at 305 nanometers) in Sarcocornia shoots were up to 48 and 40 percent higher in plants receiving ambient UV-B than in plants protected from to UV ambient light, respectively. UV exposure also increased shoot mortality, and postflowering senescence was thirty days earlier. Slight changes in the relative composition of Sarcocornia to Puccinellia in the marsh were seen. Thus, increasing UV-B is likely to pose major ecological problems for the southwestern Atlantic and Patagonean marshes by reducing plant productivity and modifying tissue composition and palatability.

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INVASIVE SPECIES

In South America, as elsewhere (Silliman and Bertness 2004), the introduction of nonnative plant species has affected the structure and dynamics of salt marsh communities. On the Pacific coast, there is a strong relation between the degree of perturbation (by agricultural activities and city development) and the number of nonnative species found on coastal marshes (Hauensteien et al. 1988). On the most perturbed areas, salt marshes usually show marked increment of exotic species, especially weeds (Ramírez, Finot, et al. 1991). According to San Martín et al. (2001), salt marshes from the central-southern part of Chile with low agricultural perturbation and high levels of human disturbances had 31 percent (of thirty-seven species) and 76 percent (of seventy-three species) of their communities composed by exotic plants, respectively. The role of invasive species in plant communities of coastal marshes can be controversial, and an illustrative case study has been recently described for “Nature’s Santuary Rio Cruces” at Valdivia City in southern Chile. During an earthquake in 1960, the whole coastal area of this river (around sixty square kilometers) dropped almost two meters, forming a salt marsh that was rapidly colonized by several

FIGURE 17.5 After road construction, two species of invasive plants species colonized the salt marsh in northern Brazil.

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species (Weischet 1960). Ramírez, San Martín, et al. (1991), describing the flora of this system, found that twenty-six of eighty plant species (32.5 percent) were exotic but that one invasive species coming from Brazil and northern Argentina (Egeria densa) represented 35 percent of plant community aerial biomass. For forty years, E. densa has been the most important dietary item for many herbivores, including some charismatic species such as the black-necked swan (Cygnus melancoryphus), grebes (Podiceps major), and coots (Fulica armillata) (Schlatter 1991). In November 2004, possibly due to the effect of the pulp mill industry (UACH 1995) or of many different waste sources (Center for Advanced Studies in Ecology and Biodiversity [CASEB] 2005), a drastic reduction on E. densa caused the mortality and massive migration of swans and grebes. The Chilean environmental agencies now have the dilemma of deciding whether to reintroduce this invasive plant in a restoration project to recover the populations of the charismatic bird species (Di Marzio and McInnes 2005). Since the South American salt marsh fauna is largely unknown, very little information is available for invasive species. The Pacific oyster (Crassostrea gigas) was illegally introduced in 1982 for gastronomic purposes to the northern

coast of Argentinean Patagonia, in Bahía Anegada (39⬚50⬘ S to 40⬚40⬘ S; Orensanz et al. 2002). This species has been successful in open sandy-muddy flats and also in areas inhabited by S. alterniflora (Escapa et al. 2004). Although the bioengineering ability of this suspension feeder is well known, there are no reports of negative or positive effects on salt marshes in South America. During 1999, the golden mussel (Limnoperna fortunei) was introduced by ballast water to Patos Lagoon in southern Brazil (Mansur et al. 2003). From June 2000 to January 2004, above-normal rainfall in the head of the lagoon resulted in a long period of freshwater conditions in the estuarine zone (mean salinity below five parts per thousand) and colonization of the intertidal of soft-bottomed enclosed bays by the golden mussel (C. E. Bemvenuti, personal communication). Although higher salinities returned during the summer of 2004 and the distribution of the golden mussel retreated to nearby small rivers, the limnification event showed that this mussel is able periodically to exploit shallow bays and marshes.

RECOMMENDATIONS According to the UN Population Division of the Department of Economic and Social Affairs (2004), South America has approximately 346 million people concentrated in cities (80 percent of the population), with most of the large cities located within one hundred kilometers of the coast. By 2050, the population is expected to reach 535 million, generating a high demand in coastal areas for infrastructure, urbanization, tourism, and the conversion of coastal habitats for uses such as agriculture and aquaculture. This scenario might be not very different from that in North America, but the strikingly different social structure in South America and large percentage of the population below the poverty line make it much more difficult to avoid further losses of salt marshes for development. In contrast to the historic situation in other parts of the world that, decades

ago, suffered the consequences of uncontrolled human impacts, the information now available may still help prevent further deterioration of South American salt marshes. Filling, salt flat conversion, and coastal erosion impacts are the most important human impacts on salt marshes of both coasts of South America, although other issues may be even more important at local levels. Marsh restoration is still a very young science in South America, is extremely expensive, and is a very low priority in countries facing tremendous health problems, including the general lack of sewage and proper disposal of solid wastes. Thus, prevention of marsh filling and salt flat conversion should have high priority for coastal managers. However, as elsewhere, the economic and societal value of South American salt marshes need to be recognized by the public and decision makers so that they can be afforded protection by environmental laws. National or regional classification of wetlands including salt marshes (sensu Cowardin et al. 1979) needs to be completed and considered by the legislation. The endless discussion of whether vegetated salt flats are salt marshes or successional stages of mangrove forests does not help the conservation of these habitats. The large-scale destruction of Ecuadorean “salt flats” by conversion to shrimp farms (Botero 1993) is a good example of how dangerous the gaps left by poorly developed environmental laws are. The first necessary step—to create an inventory of salt marshes—has been taken for a few South American countries, but mainly through the initiative of researchers and environmentalist groups (Dieges et al. 1995; Canevari et al. 1997; Isacch et al. 2006). Throughout South America are several lines of evidence pointing to significant coastal erosion, related or not to relative sea-level rise. In the cases of northeastern Brazil and Buenos Aires Province (Argentina), reduction of sediment discharge delivered to the coastal zone by the main rivers, due to water diversion, intensifies coastal erosion (Lacerda, Kjerfve, et al. 2002). It is necessary to have a supranational

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scientific evaluation for the future impacts of global change on large hydrological systems that include many countries and sustain extensive marshlands. For example, the La Plata River estuary has over eight hundred square kilometers of salt marshes, distributed along its Uruguayan and Argentinean coasts, subjected to increasing urban pollution, ranching, and sea-level rise. Managers must consider the scenario of climate change and its effects on coastal hydrology to plan the maintenance of waterfront marshes as a first line of defense against sea-level rise. Native plant species such as S. alterniflora and S. densiflora are aggressive colonizers and efficient bioengineers that might cope with rising sea level. Even in an optimistic scenario, with human population moving inland, marsh creation could be an option to meet conservation objectives (coastal biodiversity and fisheries habitat availability). Detailed classification and evaluation of human impacts already exist for the major coastal watersheds of South America (Lacerda, Kjerfve, et al. 2002) and can be used to list sites subject to similar environmental threats and degradation problems. Public education and multinationally coordinated actions should be used to preserve the integrity and ecological character of South American salt marshes. The designated authorities must also consider the need to protect selected sites against external land use changes and land-based sources of pollutants. Indeed, there are very few protected areas of coastal environment in South America. However, traditional management of salt marshes must be also considered as a conservation practice. Extensively grazed marshlands have been preserved by their value for local communities. The role of horse and cattle grazing in the structuring of a salt marsh plant community can have negative and positive effects on local biodiversity, but very little specific information is available for South America. Thus, experimental studies that exclude cattle and horses from traditionally grazed marshes must be carried out. Recently, there is strong pressure to stop grazing inside conservation 352

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units of the southwestern Atlantic coast (e.g., Peixe Lagoon, Brazil; Mar Chiquita, Argentina), as stated by environmental laws. These are resting sites of migrant birds that have being grazed for at least two hundred years, and no one knows how the vegetation will respond to the absence of grazing or how it will affect the attractiveness of the salt marshes to birds. The conservation of South American salt marshes can continue to develop only by educational efforts at the national level, backed by good science (Adam 2002), allowing public identification and valuation of these environments. Questionnaires given to sixth- and seventh-grade classrooms of schools in coastal cities of southern Brazil and Tierra del Fuego (Argentina) showed that most pupils could not recognize the importance of neighboring salt marshes or did not feel social responsibility for marsh degradation by litter disposal and waste tipping (C. S. B. Costa, personal communication). Even in countries such as Brazil, where specific federal and state laws protect both salt marshes and mangrove forests, legislation is not always enforced due to the lack of qualified environmental guards and attorneys. Thus, public education will be a major tool to reduce losses of salt marshes in the course of further urban development of South America. Acknowledgments. This project was partially supported by grants from the Universidad Nacional de Mar del Plata, the Fundación Antorchas, CONICET (Argentina), ANPCyT (Argentina PICT13527), and GEF Patagonia (all granted to O. Iribarne). C.S.B. Costa’s work was supported by grants from CNPq and PROBIO/MMA (Brazilian government). REFERENCES Acha, E. M., H. W. Mianzan, O. Iribarne, D. A. Gagliardini, C. Lasta, and P. Daleo. 2003. The role of the Rio de la Plata bottom salinity front in accumulating debris. Marine Pollution Bulletin 46: 197–202. Adaime, R. R. 1978. Estudo da variação sazonal do “standing-crop” e do repovoamento em um banco de Spartina alterniflora Loiseleur, no complexo

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Carnevale, N. J., P. Torres, S. I. Boccanelli, and J. P. Lewis. 1987. Halophilous communities and species distributions along environmental gradients in southeastern Santa Fé province, Argentina. Coenoses 2: 49–60. Center for Advanced Studies in Ecology and Biodiversity. 2005. Comentarios sobre el Informe Final de la Universidad Austral de Chile para la Dirección Regional de CONAMA X Región de los Lagos, “Estudio sobre el Origen de Mortalidades y Disminución Poblacional de Aves Acuáticas en el Santuario de la Naturaleza Carlos Anwandter, en la Provincia de Valdivia.” Santiago, Chile: Author. Chesser, R. T. 1994 Migration in South America: An overview of the austral system. Bird Conservation International 4: 91–107. Clara, M., and R. Maneyro. 1999. Humedales del Uruguay: Ejemplo de los humedales del este. Pages 68–80 in A. I. Malvarez (ed.), Tópicos sobre Humedales Subtropicales y Templados de Sudamérica. Buenos Aires: UNESCO-MAB. Cohen, M. C. L., and R. J. Lara. 2003. Temporal changes of mangrove vegetation boundaries in Amazonia: Application of GIS and remote sensing techniques. Wetlands Ecology and Management 11: 223–231. Cohen, M. C. L., R. J. Lara, C. Szlafsztein, and T. Dittmar, T. 2004. Mangrove inundation and nutrient dynamics from a GIS perspective. Wetlands Ecology and Management 12: 81–86. Collantes, M. B., J. Anchorena, and G. Koremblit. 1989. A soil nutrient gradient in magellanic Empetrum heathlands. Vegetatio 80: 183–193. Collantes, M. B., and A. M. Faggi. 1999. Los humedales del Sur de Sudamérica. Pages 15–25 in A. I. Malvarez (ed.), Tópicos sobre Humedales Subtropicales y Templados de Sudamérica. Buenos Aires: UNESCO-MAB. Costa, C. S. B. 1997. Tidal marshes and wetlands. Pages 24–26 in U. Seeliger, C. Odebrecht, and J. P. Castello (eds.), Subtropical Convergence Environments: The Coast and Sea in the WarmTemperate Southwestern Atlantic. Berlin: Springer. Costa, C. S. B., R. Armstrong, Y. Detrés, E. W. Koch, M. Bertiller, A. Beeskow, L. S. Neves, G. M. Tourn, O. A. Bianciotto, L. M. Pinedo, A. Y. Blessio, and N. San Roman. 2006. Effect of UV-B radiation on salt marsh vegetation: Trends of the genus Salicornia along the Americas. Photochemistry and Photobiology 82: 878–86. Costa, C. S. B., C. V. Cordazzo, and U. Seeliger. 1996. Shore disturbance and foredune plant distribution. Journal of Coastal Research 12, no. 1: 133–140. Costa, C. S. B., and A. J. Davy. 1992. Coastal salt marsh communities of Latin America. Pages

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Anthropogenic Threats to Australasian Coastal Salt Marshes Mads S. Thomsen, Paul Adam, and Brian R. Silliman Australasian coastal salt marshes have experienced intensive human modifications on a relatively short time scale compared to many modified marshes in America, Africa, Asia, and Europe. In this chapter, we review anthropogenic threats first to New Zealand and then Australian salt marshes and provide recommendations to abate these threats and improve restoration efforts of these important coastal communities. Initial impacts by Australian aboriginals (about forty thousand years ago) and New Zealand Maori (about eight hundred years ago) were likely to have been small and were limited to harvesting of plants and shellfish. However, following a few centuries of colonization, indirect effects of Maori avian hunting and coastal forest clearance could have reduced herbivore browsing considerably and caused increased runoff, with increased sedimentation and nutrient inputs in New Zealand marshes. With European colonization about two hundred years ago, reclamation of marshes became a major focus of human development (an imported habit from England), and large salt marshes were converted to urban structures (roads, buildings, ports, marinas) and agricultural grassland. Today, reclamation threats are largely under control and generally prohibited by planning legislation. Still, a suite of indirect anthropogenic stresses (e.g., livestock grazing, eutrophication, and invasions by nonindigenous species) threaten the marshes that remain. Quantitative distribution data of specific marsh species are sparse, descriptive studies dominate the literature, and only a few processoriented experimental studies have been conducted to elucidate factors controlling community structure. Thus, human impacts on Australasian salt marshes and how they may impact community organization are largely based on anecdotal evidence, older semiquantitative surveys, and gray literature reports. Imminent visible threats include livestock grazing/trampling, human trampling and usage of recreational vehicles, waste dumping, invasions by nonindigenous species, and mangrove expansion. Less visible threats include storm water runoff and contamination with persistent organic molecules (e.g., pesticides) and heavy metals. Enhanced sedimentation and eutrophication are major problems in many larger estuaries with intense urbanization, but little is known about how these stressors affect bordering salt marsh structure and function. The consequences of climatic change (sea-level rise, temperature rise, 361

increased storminess, ozone destruction) will affect future Australasian salt marshes, although there is much to learn about both the changes to the environment and the biotic response. Clearly, large-scale inventories are still very much needed to track changes, but the application of high-resolution remote sensing and geographic information system (GIS) techniques will assist the analyses of changes. We suggest that managers outline and carry out long-term monitoring programs that capture present and future distribution ranges of key marshes, plants, and animals, but also that researchers supplement such programs with controlled experiments that test for effects of grazing, competition, facilitation, nutrient limitation, soil texture, and various human stressors, to ensure that we can detect cause-and-effect relationships and thereby create means to ameliorate and mitigate future threats. Ideally, there should also be continued collaboration and data interchange between stakeholders to ensure synergistic benefits for both the managers (e.g., use the experimental data to adjust monitoring programs and provide causation to their distribution data) and researchers (use the monitoring data as background information for experiments and to extrapolate findings to larger spatiotemporal scales).

Coastal salt marshes and salt meadow (hereafter salt marshes) are some of the most important habitats in the world. Marshes protect coastlands from erosion and floods, act as feeding grounds and shelter for aquatic and terrestrial species, provide carbon sinks that reduce global warming impacts, and are natural systems of wastewater treatment that reduce adverse effects of eutrophication and pollution (Chapman 1974; Adam 1995; Valiela and Cole 2002; Bertness, Silliman, and Jefferies 2004). Salt marshes are also important ecosystems within Australasia (New Zealand and Australia) and are likely to be increasingly impacted by humans as populations grow, urbanization increases, climatic change accelerates, and the coastal zone is further utilized for human development, fisheries, and aquaculture. To be able to protect marshes from these threats, a review of what is known about Australasian marshes and human threats to these marshes is timely. We address the review in separate sections for New Zealand and Australia partly because the two countries have been geographically separated for more than seventy million years and are distinctively different with respect to natural history (e.g., they vary widely in geology, tectonic activity, geomorphology, topography,

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climatology, oceanography, animal and plant distribution patterns, biogeography, degree of endemics), partly because of the two countries’ different colonization histories (e.g., Australia was colonized forty thousand years ago, vs. New Zealand only about eight hundred years ago), degree of human-induced extinction rates, and different legislation and administration. During the review, we noted that quantitative scientificbased knowledge regarding New Zealand salt marsh ecology was relatively sparse. Hence, to advocate and stimulate salt marsh research from this minicontinent, we included a more thorough review of both the human colonization history and of the existing salt marsh literature from this country (e.g., we here include Web-based information and a questionnaire to local managers).

NEW ZEALAND SALT MARSHES, HISTORY, MANAGEMENT, AND HUMAN THREATS LATE HUMAN COLONIZATION

New Zealand is considered the last large landmass to be colonized by humans (Diamond 1999). East Polynesians (ancestral Maori) discovered and colonized New Zealand only seven

to eight hundred years ago, first settling around estuaries (King 2003; McFadgen 2003)—that is, in the vicinity of coastal salt marshes. The first few settlers likely had only a minor impact on salt marshes, mainly hunting birds, fishing, and collecting snails, mussels, oysters, and coastal flax for weaving. However, with a rapidly growing population, indirect effects on the marshes could have been dramatic within a few centuries (McGlone 1989). For example, in less than three hundred years from first colonization, more than thirty endemic bird species, many of which likely foraged in marshes— including coastal geese, swans, pelicans, rails, and eleven moa species—were probably hunted to extinction. An estimated thirty thousand to ninety thousand moa were killed around the mouth of the Waitaki River alone (King 2003). In addition, millions of nesting seabirds and coastal seals were killed (the former also by introduced rats and dogs) and large areas of coastal forest burned, probably for hunting purposes (85 percent forest cover reduced to 55 percent before European settlement; Hutching 1998; King 2003). Removal of an abundant avian megafauna coupled with alteration of drainage basin properties (e.g., increased sedimentation would be expected following coastal deforestation) is likely to have altered salt marsh community structure and ecosystem functioning. For example, in the Northern Hemisphere, it has been shown that geese can have dramatic impacts on marshes, including both a negative effect of grubbing rhizomes and a positive effect associated with nutrient enrichment (Jefferies 1988; Esselink et al. 1997; Dormann and Bakker 2000; Bos 2002; Bos et al. 2002; Jefferies and Rockwell 2002). Thus, the extermination of the prolific populations of groundbrowsing birds has potentially created the present view of “naturally nongrazed” marshes in New Zealand. This removal of herbivore top-down control is analogous to the nearextermination of megaherbivores by humans in the Caribbean (manatees and turtles; Jackson 1997) and on the North American Plains,

although the latter extinctions are also attributed to climate changes (Owen-Smith 1989). Utilization and alterations of the coastal zone accelerated with the arrival of European whalers and farmers in the late eighteenth century. For example, nine additional birds species became extinct, seals and coastal birds in more remote areas were hunted to near-extinction, and intensive large-scale farming and urbanization commenced. Today, after only eight hundred years of turbulent human history, around four million people inhabit New Zealand’s approximate 270,000 square kilometers, most living in close proximity to its 11,000-kilometer coastline (McLay 1976; Hume 2003) (some authors estimate up to 18,000 kilometers; Hutching 1998; Bell, Hume, and Hicks 2001; King 2003; Rouse, Nichol, and Goff 2003). This short human history has greatly modified the natural landscape, with examples such as clearance of more than 85 percent of buffering coastal forest (Auckland Regional Council 2000a). Such removal can have large-scale impacts on salt marshes by releasing unfiltered nutrient-rich waters into the marsh (Bertness, Ewanchuk, and Silliman 2002; Silliman and Bertness 2004); destruction of more than 85 percent of wetlands (Taylor and Smith 1997; Auckland Regional Council 2000b); “harvest” of millions of marine mammals and coastal birds (often to local extinctions); extinction of half of the endemic bird fauna (King 2003); introduction of twenty-five thousand nonnative plant species, of which about two thousand survive without human aid (Holland 2001); reduction of the native forest to a quarter of its prehistoric extent (Holland 2001); and conversion of almost fifteen million hectares to farmland (Statistics New Zealand 2002) inhabited by millions of sheep, cattle, and deer (Hutching 1998). Such short-term and large-scale landscape alterations can only have had dramatic impacts on the extent, community structure, and ecosystem functioning of New Zealand salt marshes. However, data showing direct linkages between this activity and changes in marsh

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community structure, as has been shown elsewhere, is largely lacking for New Zealand. SALT MARSHES OF NEW ZEALAND

More than 80 percent of New Zealand’s coastline is “energetically exposed” with sandy or rocky substratum. Still, some 300 to 350 lowenergy bar-built estuaries, drowned river valleys, river mouths, and lagoons exist (hereafter estuaries) (McLay 1976; Bell et al. 2001), where siltation, wave protection, and semidiurnal tides, typically of one to three meters, provide conditions for salt marsh development. Sea level was about 120 meters below mean sea level (MSL) 18,000 years ago but stabilized at its present level about 6,500 years BP (Hume 2003), leaving a relatively short time for siltation, dispersals, and establishment of salt marshes along the present coastline. The estuaries vary from a few hectares to more than fifteen thousand hectares, with two-thirds being less than five hundred hectares (McLay 1976). The areal extent of salt marshes at regional and national scales has been determined from Landsat 7 images (30 ⫻ 30 ⫺ meters resolution, 1996–1997 images, LCDB1 category Coastal Wetland ⫽ salt marsh ⫹ coastal flax [Phormium tenax], minimum habitat size ⫽ 1 hectare). From this large-scale inventory, it is estimated that 10,400 hectares exist on the North Island and 9,700 hectares on the South Island (http:// www.maf.govt.nz/statistics/primaryindustries/ landcover). This total area of 20,100 hectares is slightly smaller than the extent of the mangrove Avicennia marina (⫽A. resinifera), only found north of 38⬚ S of 22,500 hectares. However, because numerous New Zealand marshes are fringing and distributed in small pockets and along thin lines, cumulative large areas are likely to remain undetected by the Landsat resolution. A few large-scale semiquantitative inventories (Braun-Blanquet methodology) have described national distributions of species and communities (Thannheiser and Holland 1994; Haacks and Thannheiser 2003). According to Haacks and Thannheiser (2003), seventy-four salt marsh plant species exist in New Zealand, with 364

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almost half being nonnative and only seven species endemic (Carex litorosa, Lachnogrostis littoralis, Leptinella [⫽ Cotula] dioica, Leptocarpus [⫽ Apodasmia] similis, Plagianthus divaricatus, Poa cita, Puccinellia walkeri). This is a low number of endemics, compared to the national average of 80 percent endemics of all native species (Holland 2001), but not atypical for salt marshes. Many New Zealand salt marsh plant genera are cosmopolitan, and many species are shared with Australia. In general, distribution patterns resemble patterns from southern Australia (Chapman 1974; Adam 1990), with a typical zonation consisting of (from low to high elevation) the seagrass Zostera Capricornil (below midtide), the mangrove A. marina (only above 38⬚), Sarcocornia quinqueflora/Samolus repens (low marsh), Juncus kraussii (referred to in most New Zealand accounts as J. maritimus)/Leptocarpus similis (middle-high marsh), and bordered upland by the flax P. tenax and coastal shrubs (Plagianthus divaricatus, Coprosma spp., and Leptospermum scoparium) (fig. 18.1; Morton and Miller 1973; Chapman 1974; Thannheiser and Holland 1994; Haacks and Thannheiser 2003; Wardle 1991). S. repens and S. quinqueflora are considered pioneer species (Chapman 1974; Thannheiser and Holland 1994), and they can also be found at higher elevations—for example, in disturbed patches, and scattered on open shingle, sand, and mudflats. Vertical zonation pattern is not always as clear-cut as described from North American and European coastlines; and because species limits often are reversed, species distribution is sometimes referred to as “mosaic-like” (Chapman 1974; Thannheiser and Holland 1994). For example, L. similis is not found in all marshes and is sometimes found as the low marsh species, with S. quinqueflora occurring in the higher marsh in disturbance patches (e.g., in some marshes in the Catlins). It is possible that the shifting vertical patterns and competitive hierarchies among plants are modified and/or flip-flopped by local differences in freshwater input (Wilson et al. 1996), soil textures (Partridge and Wilson 1989), climatic

FIGURE 18.1 Border between Samolus repens (front, low marsh) and Juncus kraussii (back, middle marsh) zones. Note how drift log can accumulate in marshes (center), causing disturbances and opening up spaces for pioneer species like Sarcocornia quinqueflora and S. repens.

conditions (Haacks and Thannheiser 2003), and nutrient levels (Levine, Brewer, and Bertness 1998). For example, it is likely that S. quinqueflora gains a “competitive edge” over S. repens on coarse sandy soils and L. similis over J. kraussii with high freshwater inputs (personal observation). Similar freshwater–saltwater dominance reversals have been documented experimentally for North American salt marsh plants (Crain et al. 2004; Pennings, Grant, and Bertness 2005). Additional common native species include Selliera radicans, Baumea juncea, Suaeda novae-zelandiae, Leptinella dioica, Schoenoplectus pungens, Bolboschoenus medianus, Mimulus repens, Puccinella stricta, P. walkeri, Triglochin striatum, and Apium prostratum (Johnson 1989; Wilson et al. 1996; Haacks and Thannheiser 2003). Several accounts exist from local salt marshes (but unfortunately often not presenting biomass values with standard errors)—for example, Pollen Island, Auckland (Chapman and Ronaldson 1958); Avon-Heathcote Estuary, Christchurch (Mason 1969; Knox 1992; Webster 1997; Thomsen, Marsden, and Sparrow 2005); Lake Ellesmere, Canterbury (Evans 1953), Hapuka Estuary, West Coast (Dickinson and Mark 1999); and Nelson Haven, Nelson/ Marlborough (Davies 1931; Doak 1931). The bestdescribed marshes are from the Otago region

where it has been shown that vertical zonation patterns correlate with salinity tolerances, freshwater inflow, soil texture, and “local peculiarities” (Paviour-Smith 1956; Partridge and Wilson 1987a, 1987b, 1988a, 1988b, 1989; King, Wilson, and Sykes 1990; Wilson et al. 1996). In one of the few published manipulative experiments, Partridge and Wilson (1988a) showed that in Otago, most species survived upelevation transplantation but typically died following down-transplantation, suggesting competitive limitation upward and physiological stress limitation downward. However, because neighbor presence (i.e., competitive and facilitative interactions) was not manipulated, this conclusion is not definitive. There are few data on animals utilizing New Zealand salt marshes, and this research area is therefore wide open for eager scientists. Due to a seventy- to eighty-million-year-long geographic isolation following the break-off from Australia and dramatic ongoing changes in geology and climate (e.g., thirty-five million years ago, New Zealand was reduced to a few small flat islands), mammals (except bats) never occupied New Zealand (Harvey 2001). Thus, browsing would have been by the recently extinct moa (Anomalopterynginea spp.) and geese (Cnemiornis spp.) (Atkinson and Greenwood 1989). The few

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scientific papers we are aware of that quantify salt marsh fauna showed that amphipods, nematodes, oligochaetes, and diptera larvae were the most common in terms of biomass in an Otago salt meadow (Paviour-Smith 1956) and that the marine invertebrate fauna of Christchurch marshes is sparse, with only four mollusks, five crustaceans, and five polychaetes. None of these species were found exclusively in salt marshes (Marsden and Heremaia 1998). Qualitative observations and data from adjacent estuarine ecosystems suggest that key salt marsh species with potential large-scale ecological effects likely are the burrowing crab Helice crassa and the gastropods Amphibola crenata, Ophicardelus costellaris, and Potamopyrgus estuarinus (due to their ubiquity throughout New Zealand—personal observation; Morton and Miller 1973; Jones and Simons 1983; Juniper 1986; Marsden and Heremaia 1998). The crab (H. crassa) oxygenates soils and increases bioturbation (Morrisey et al. 1999; Williamson et al. 1999; Gibbs, Thrush, and Ellis 2001) and as a consequence may increase primary production particularly in poorly drained soils (Bertness 1985), whereas the gastropods are likely to control decomposition and/or bacterial and primary production (Juniper 1987a, 1987b; Silliman and Bortolus 2003; Silliman and Newell 2003). Few data exist on how birds utilize New Zealand salt marshes, but PaviourSmith (1956) note that migrating dotterels winter on salt meadows and that gulls, oystercatchers, stilts, and godwits commonly feed in salt meadows or seek shelter behind sedges/ rushes at high tide and under strong winds (we have observed similar patterns in both North and South Island salt marshes). Also, Lowe (1997; Marsden and Heremaia 1998) has shown that densities of oystercatchers were highest in a marsh not separated by stop banks and that stilts were most abundant in a newly artificially formed marsh. In this study, densities, foraging time and prey capture efficiency varied between bird species, salt marsh sites, and seasons, but not in any simple consistent manner. Also, little is known about fish utiliza366

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tion, but Nairn (1998; Marsden and Heremaia 1998) found abundant juvenile flounders and mullets in Christchurch marsh channels during summer months. We expect fish utilization to be as important as found in New South Wales (NSW), Australia, where up to sixteen fish species, six of commercial importance, were caught in pop nets at high tide in the Sarcocornia zone, with a mean density of 0.6 individual per square meter (Mazumder, Saintilan, and Williams 2005a, 2005b). Similar measurements should be made in different New Zealand salt marsh zones, preferentially coupled with experiments that document interactions between visiting fish and birds and permanent salt marsh inhabitants. MANAGEMENT OF NEW ZEALAND SALT MARSHES

New Zealand salt marshes are mainly managed by regional councils and some city councils, but a large proportion is also under management by the Department of Conservation (some marshes are also under private ownership). For example, to conduct scientific research in a salt marsh, resource consent must be obtained from the appropriate council; and if the marsh is also a Department of Conservation wildlife reserve, an additional research permit must be obtained. Some councils, the Department of Conservation, and local trusts have carried out salt marsh restoration projects (Bergin 1994; Auckland Regional Council 2000a, 2000b; fig. 18.2), and seeds of the most common salt marsh species can be bought from commercial nurseries (e.g., J. kraussii, L. similis, New Zealand Tree Seeds), obtained from local voluntary groups/trusts (e.g., by the Guardians of Pauatahanui Inlet, personal observation), or attained from council nurseries (Thomsen et al. 2005). It has been shown that transplantation success of J. kraussii and L. similis depended on initial transplant size (the larger the better) but not on transplant spacing or nutrient additions (twenty-five grams of slow-release NPK pellets added to each transplant; Bergin 1994). The study by Thomsen et al. (2005) confirmed that

Zealand salt marshes, such as by testing for competition and facilitation, repeated nutrient additions, timing of transplantation, and effects of multispecies versus single-species plantings. ANTHROPOGENIC THREATS

FIGURE 18.2 Example of a restorations project in Pauatahanui inlet near Wellington. Juncus kraussii and Leptocarpus similis have been transplanted to restore local marshes.

FIGURE 18.3 Leptocarpus similus is relatively easy to transplant to high marshes, particularly around freshwater sources. Several councils and trusts grow the species in nurseries, making this species readily available for restoration projects.

these two key species are suitable for use in restoration (fig. 18.3), whereas Schoenoplectus pungens, despite an initial transplant survival, failed to regenerate the next spring following seasonal dieback. Thomsen et al. (2005) also showed that the survival and growth of J. kraussii and L. similis did not depend on soil type (dredged estuarine soils vs. marsh soils) but that plants from natural populations had higher biomass than plants out-transplanted from a local nursery, presumably because the nursery plants were “softer”—that is, less resistant to wind, currents, and waves and/or were more exposed to rabbit grazing. Still, much applied ecological information is needed to ensure efficient restoration, reestablishment, and management of New

Without information on the prehuman extent of nationwide salt marsh area, it is not possible to quantify with certainty the area of salt marsh lost. However, in Bay of Plenty, an estimated 56 percent of estuarine wetlands on harbor margins has been lost/reclaimed (calculated from Park 2000 and the New Zealand Land Cover Data Base, http://www.mfe.govt.nz/ issues/land, accessed 2006). This regional loss is less than intensively human alienated North American coastal areas (e.g., up to 80 percent lost salt marshes in New England; Bertness et al. 2002) and much less than the more than 90 percent salt marsh loss suggested by Marsden and Heremaia for the whole of New Zealand (1998, but no data and/or references follow this high suggested loss). Anthropogenic impacts to New Zealand salt marshes can be grouped into four general lines of threats: (1) land reclamation and impacts associated with utilization of watershed (e.g., livestock grazing, eutrophication, and enhanced sedimentation); (2) pollution and human trampling effects; (3) biological invasions; and (4) climatic changes. Because of the scarcity in the literature of hard data, we submitted a questionnaire to managers in the regional councils regarding local salt marsh management and perceived threats. Six councils replied, and we have included their perception about threats to local and regional salt marshes in the text. LAND CLAIM AND HUMAN WATERSHED UTILIZATION

Land claim has probably had the most severe impact on New Zealand marshes (Haacks and Thannheiser 2003; see estimates presented earlier). For example, six major ports and numerous smaller harbors have been built in New Zealand estuaries, and these constructions typically involved considerably foreshore

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reclamation (Hume 2003). Also, at least 164 causeways have been constructed for roads and railways in estuaries (Hume 2003). The causeways reduce/cutoff tidal flow and the hinterland marshes have often been converted into farmland or urban buildings (personal observation). Even where wetlands remain behind the causeway, vegetation composition will typically be very different with degraded water quality, decreased salt water input, and reduced abundance of fish and macroinvertebrates (Roman, Garvine, and Portnoy 1995; fig. 18.4), as has also been documented in tidally cutoff or restricted mangroves (Layman et al. 2004; Layman, Arrington, and Blackwell 2005). Most large New Zealand urban areas are coastal, and the development of these areas would certainly have included reclamation of salt marsh areas. The growing population of New Zealand (e.g., estimated 4.4 million in 2021, “Part 9: Sub-national Demographic Projections,” Demographic Trends 2001, Statistics New Zealand), and a trend of people seeking coastal residency will continue to put pressure on existing salt marshes. Still, many salt marshes are under management by the Department of Conservation and/or regional councils, and this is probably why contemporary land reclamation for urban development are presently only considered a relatively low risk by

FIGURE 18.4 A causeway in Pauatahanui inlet that cut through a marsh, reducing tidal flow and habitat connectivity. The vegetation composition in the marsh behind the causeway will typically be different, sometimes with degraded water quality, decreased saltwater input, and reduced abundance of fish and macroinvertebrates.

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managers (average 2.2 [out of 2, 2, 3, 1, 2, 2 [where score 1 ⫽ not considered a problem, score 2 ⫽ considered a minor problem, score 3 ⫽ considered a major problem, and score ? ⫽ potentially a problem, but considered unknown]). Agricultural land claims have probably also taken a high toll on New Zealand salt marshes. Enormous areas of New Zealand’s landscape have, in a few centuries, been converted to pasture (11.5 million hectares) and other agricultural practices (an additional 4 million hectares) (Statistics New Zealand 2002). Again, it is unknown how much salt mash has been converted to grassland, but agricultural land claim was considered slightly less of a threat than urban land claims by regional managers (average 1.8 [2, 3, 2, 2, 2, 1]). In addition to conversion of salt marshes to pastoral grassland, agriculture also affect salt marshes directly by livestock grazing of existing salt marsh vegetation and by indirect nutrient runoff effects (Jensen 1985; Andresen et al. 1990; Kiehl, Esselink, and Bakker 1996; Kiehl et al. 1997; Levine et al. 1998; Esselink, Fresco, and Dijkema 2002). For example, in Europe, grazing at low to medium livestock densities increase species richness by reducing cover of the tall Elymus athericus and Atriplex portulacoides and thereby decreasing competition and allowing for the coexistence of smaller species like Puccinellia maritima, Triglochin maritima, and Plantago maritima (Bos et al. 2002). Following land claim, intensive utilization of the watershed is likely to have had the secondmost dramatic impacts on New Zealand salt marshes (Haacks and Thannheiser 2003). The main livestock are sheep (39.5 million), cattle (9.7 million), and deer (1.6 million) (Statistics New Zealand 2002), with the latter mainly kept under strict fenced conditions. These large numbers correspond to some of the highest densities in the world (Taylor and Smith 1997), and because farmers often do not restrict sheep access to marshes, browsing and trampling likely have important effects on salt marshes, particularly as they evolved without browsing mammals. This is reflected in livestock grazing being considered a medium to high risk

(average 2.3 [3, 2, 3, 2, 2, 2]). More specifically, Hacks and Thannheiser (2003) suggest that C. coronopifolia is facilitated by grazing but that S. repens and S. radicans are relative insensitive to grazing. They also observed that S. quinqueflora, S. novae-zealandica, M. repens, and Puccinellia spp. were found very rarely in intensively grazed areas. Wraight (1964) reports that in a salt meadow in Lake Ellesmere, abundant A. stolonifera and Trifolium fragiferum were found with little grazing, whereas J. kraussii, Hordeum marinum, S. radicans, P. coronopus, and C. dioica were more abundant under moderate to heavy grazing. Finally, A. prostratum, Atriplex patula, M. repens, and T. fragiferum were most impacted by grazing. Despite these observations, it was concluded in a recent review undertaken to estimate livestock effects on New Zealand wetlands that little information was available (not a single reference was found that documented effects on salt marshes; Reeves and Champion 2004). Clearly, it is important to conduct large-scale correlative surveys to quantify the extent of salt marsh grazing and couple these surveys with cage exclusion experiments and feeding preference trials. Given that cattle tend to be relatively unselective grass “tearers” compared to sheep, which are more selective “biters” (Bos et al. 2002), it cannot be assumed that the effect of grazing on New Zealand salt marsh plants is similar for sheep and cattle grazing. A by-product of intensive large-scale agriculture is nutrient enrichment, which has become an increasing problem in New Zealand freshwater and coastal waters (Taylor and Smith 1997). Point source sewage from the larger urban areas has previously been a large-scale nutrient contributor but is today somewhat limited due to sewage treatment, although storm water runoff still causes frequent peaks of high nutrient outflow (Taylor and Smith 1997). Due to the high livestock densities grazing on steep and shallow soils (more than 70 percent of New Zealand is considered steep topography; Hutching 1998) with high applied fertilizer rates, which accelerates nutrient runoff, the occurrence of algal blooms are an increasing

problem in freshwater and estuarine habitats (Taylor and Smith 1997; Parkyn et al. 2002). This is probably why eutrophication is considered a medium to high threat to New Zealand salt marshes by the regional managers (average 2.3 [?, 2, 2, 2, 3, ?]). Although no studies have linked nutrient availability to New Zealand salt marsh ecology, we expect that enhanced nutrient supply will favor epibenthic diatoms and ephemeral green algae (Fletcher 1996; Raffaelli, Raven, and Poole 1998), facilitate nonnative opportunistic marsh plants (Silliman and Bertness 2004), increase salt marsh productivity (Kiehl et al. 1997; Silliman and Zieman 2001; Brewer 2003), and alter competitive hierarchies and thereby dominance patterns (Levine et al. 1998). Indeed, it is important to reiterate that salt marshes are buffer zones between land and sea, typically facilitating nutrient uptake and transformation as well as sediment deposition (see later discussion) and thereby reduce these stressors on adjacent, more susceptible seagrass, macroalgal, and oyster beds (Valiela and Bowen 2002; Valiela and Cole 2002). Thus, maintaining and/or restoring these buffer zones could be a partial management option to reduce local problems of enhanced sedimentation and eutrophication. New Zealand watersheds are prone to erosion and high sediment transport, given the steep topography combined with relatively soft bedrocks, relatively high precipitation rates (typical values of 800 to 1,200 mmy⫺1 but up to 11,000 mmy⫺1 in Fjordland; Stuarman, Owens, and Fitzharris 2001), and a high tectonic activity (i.e., situated on the convergence zone of the Australian continental plate and the Pacific oceanic plate). However, due to a high forest cover, prehuman estuarine sedimentation rates were relatively low (less than one millimeter per year; Hume 2003). Following forest clearings, urbanization, and intensive livestock farming, it is today estimated that more than eighteen million hectares are threatened by erosion (Hutching 1998), and estuarine sedimentation rates have typically doubled or tripled (Hume 2003). We are not aware of any studies that have

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tested for effects of sedimentation on New Zealand salt marsh community structure, but Swales, MacDonald, and Green (2004) showed that, at wave-exposed sites, Spartina facilitates shell accumulations, which feedbacks to again stabilize Spartina. In other countries, it has been shown that sedimentation can enhance salt marsh development (Gibblin, Valiela, and Teal 1983; DeLaune et al. 1990), presumably by reducing drowning and enhancing nutrient supply. New Zealand salt marshes have been shown to accumulate three to twelve millimeters per year (Lee and Partridge 1983, although this is for invasive Spartina species), documenting that at least this species does well under medium sedimentation levels. Today, sedimentation is considered the most important “contaminant” in the coastal zone (Williamson et al. 2003) and is also considered a high risk to salt marshes by regional managers (average 2.7 [3, 3, 2, 2, 3, 3]). Nevertheless, we disagree with this survey and predict salt marsh plants to be relatively robust to sedimentation (DeLaune et al. 1990; French and Spencer 1993; Kastler and Wiberg 1996), for example, compared to rocky shores, oyster reefs, seagrass beds, and ecosystems inhabited by filter feeders (Airoldi 2003). Clearly, specific tests should be conducted to evaluate the sediment sensitivity of native New Zealand salt marsh plants and animals. An indirect effect of sedimentation on salt marsh plant distribution is competitive displacement by A. marina in northern estuaries. Here, A. marina has been observed to expand onto open mudflats (Young and Harveya 1996; Nicholls and Ellis 2002; Ellis et al. 2004; Park 2004) and potentially also salt marshes. The expansion has been associated with high estuarine sedimentation rates (“infilling,” partly natural, partly human caused), but additional factors such as altered precipitation patterns, sea-level rise, eutrophication, and temperature rise are possible alternative causes (Saintilan and Williams 1999). If A. marina expands onto higher ground and the hinterland is fixed (e.g., by stop banks and causeways), salt marshes will be squeezed and eventually outcompeted 370

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(sea-level rise–coastal squeeze scenarios). No studies have tested for competitive effects between A. marina and salt marsh plants, and we are only aware of one report that shows subtle mangrove expansion into a salt marsh area (Park 2004). Still, mangrove expansion was considered a high risk by northern council managers (average 2.7 [3, 3, 2]). The construction of about eighty large hydrodams throughout New Zealand (providing nearly 75 percent of the national electricity demand; Taylor and Smith 1997) has clearly also altered watershed properties. Such large-scale control of flow conditions likely affect coastal river mouth salt marshes by controlling the extent, duration, and frequency of freshwater flooding, but we are not aware of any study that have linked dam-induced altered freshwater flow patterns to salt marsh ecology. Alteration of river mouth salinity regimes will ultimately control species patterns in fringing marshes, as most species have optimal growth in freshwater, but different tolerances to salinity and thereby different competitive advantages; for example, reducing freshwater river outlets will favor the slower-growing, more salt-tolerant species (Partridge and Wilson 1987a, 1987b; Wilson et al. 1996; Crain et al. 2004; Pennings et al. 2005). POLLUTION AND TRAMPLING

Pollution with heavy metals, oil, pesticides, and persistent organic pollutants (POP, such as PAH, PCB, and DDT) is relatively low compared to more populated and heavily industrialized Northern Hemisphere countries (Williamson et al. 2003). Heavy metals have been discharged from mining operations, tanneries, fertilizer work, and other industries primarily in the period from 1890 to 1960. Today, many point source pollutions have been cleaned up or routed through sewage treatment plants, and surface sediment concentrations are lower than in the past because pollutants have been diluted or buried by less polluted sediments (Williamson et al. 2003). Contemporary sources of heavy metals are mainly urban storm water runoff, industrial spills, boating and antifouling

paints, and geothermal discharges (Roper, Thrush, and Smith 1988; Taylor and Smith 1997). Pollution by toxic organic chemicals is also considered relatively low in New Zealand, with slowly degrading organochlorine pesticides mainly being applied in agriculture between 1940 and 1970, but today phased out by legislation. For example, use of PCB has been illegal since 1995 and its pollution effect is therefore diminishing as residuals are transformed and/ or buried. In contrast, it is expected that polycyclic aromatic hydrocarbons released from combustion of fossil fuel and from oil spills will increase in importance due to continued coastal urbanization and an increasing demand for fossil fuels (Williamson et al. 2003). The ultimate fate of heavy metal and POP discharges is incorporation into in- and offshore sediments, and data on metal and POP levels in estuarine sediments are available both in regional council reports and refereed literature (Roper et al. 1988), but little is known about concentrations in New Zealand salt marsh sediment. Despite most metals and POP showing increased levels in near-shore sediments compared to preindustrialization levels, it has been concluded that compared to impacted overseas areas, New Zealand contamination is relatively low (although highly impacted sites have been found around harbors; Williamson et al. 2003). Still, POPs from urban storm waters and agricultural runoff will continue to enter estuaries and salt marshes in the future. Accreting salt marshes are generally considered sinks for these pollutants (Leendertse, Scholten, and van der Wal 1996), although resuspension events during storms can still transport contaminated sediments out of the marsh. A net pollutant loss is particularly expected in wave-exposed retreating salt marshes (Swales, MacDonald, and Green 2004; Swales, Ovenden, MacDonald, Lohrer, et al. 2005) or salt marshes disturbed by human or livestock trampling. In general, we expect elevated levels of pollutants in New Zealand marshes, but we are not aware of data that verify this or of New Zealand field studies about

how marsh structure or functions are affected by pollutants. Indeed, typical salt marsh invertebrates (H. crassa, A. crenata, and P. estuarinus) were found to be relatively resistant to the herbicide mixture dalapon/weedazol in laboratory toxicity tests (Gillespie 1989). Still, it is too simplistic to extrapolate the laboratory LC50 tests to in situ food web structures and community interactions. Nevertheless, the regional managers consider pollution and pesticide runoff to be only medium threats to the salt marsh health (average 2.0 [2, 2, 2, ?, 2, ?] and 1.8 [1, 2, 2, ?, 2, ?], respectively). Dumping of human rubbish is a much more visual pollutant. Dumping is illegal and also considered a medium threat to salt marshes (average 2.0 [1, 2, 3, 2, 2, 2]). Ecological effects are probably minor as long as the dumping are of small quantities and do not contain degrading toxic substances, but the aesthetic impacts are disproportionately large, particularly because marshes typically are considered “wild nature” primarily visited by naturalists. We are not aware of any data that show the extent or effects of rubbish dumping in New Zealand marshes. Of potential more harm is the usage of salt marshes for recreational vehicles and trampling where it can take many years for a marsh to recover (Adam 2002). These negative effects are partly mitigated in marshes managed by the Department of Conservation by designated boardwalks, information boards, and special legislation. Again, we are not aware of any studies that have quantified the extent or impacts of these stressors, but it is considered a medium to high salt marsh threat by the regional managers (average 2.4 [?, 2, 2, 3, 3, 2]). BIOLOGICAL INVASIONS

Almost half of the New Zealand salt marsh plants are nonnative. This matches the general level of nonnative angiosperms where 2,000 nonnative species persist without human aid, compared to 2,300 native angiosperms (Holland 2001). Of all the nonnative plant species, around two hundred are considered pests that pose a risk to native plants. Of the nonnative salt marsh

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species, Spartina anglica and S. alterniflora in particular are considered highly invasive. S. alterniflora is mainly found in North Island estuaries, whereas S. anglica can be found occasionally in estuaries throughout the country (Partridge 1987; Swales, Ovenden, MacDonald, Lohrer, et al. 2005). Spartina was deliberately planted in New Zealand from the beginning of the twentieth century to reduce erosion and claim land, and the first scientific reports focused on these “advantages” (“converting mudflats to productive grass lands”; Allan 1924, 1930; Harbord 1949). From the late 1960s, this viewpoint was challenged, and researchers began to emphasize problems associated with mudflat destruction and alteration of native salt marsh communities (Bascand 1968; Bascand 1970). Today, Spartina is considered a pest species, and most regional councils have implemented eradication programs, typically spraying with herbicides (Lee and Partridge 1983; Partridge 1987; Jamieson 1994; Roper et al. 1996; Shaw and Gosling 1996, 1997; Turner and Hewitt 1997; Swales, Ovenden, MacDonald, Burt, et al. 2002; Swales, MacDonald, and Green 2004; Swales, Ovenden, MacDonald, Lohrer, et al. 2005). Despite many reports on Spartina in New Zealand (see previously listed references), we are not aware of any scientific studies that test for effects of Spartina on native salt marsh species. Other common nonnative plants include Paspalum vaginatum, Agrostis stolonifera, Atriplex prostrata, Festuca arundinacea, Juncus acutus, J. gerardii, Plantago coronopus, Plantago australis, Puccinellia distans, and P. fasciculata (it is sometimes argued that Cotula coronopifolia is nonnative, but until this has be rigorously established, we consider it a native) (Johnson 1989; Wilson et al. 1996; Graeme and Kendal 2001; Haacks and Thannheiser 2003; Shaw and Allen 2003). Several of these species can be locally abundant (Graeme and Kendal 2001). The list of salt marsh invaders continues to grow, with, for example, the recent detection of Limonium companyonis in the Avon-Heathcote Estuary (Heenan et al. 1999; McCombs and von Tippelskirch 2004). It is interesting that despite New Zealand 372

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managers and scientist being highly aware and respondent to problems associated with nonnative species, such as Spartina eradication programs (Forrest, Taylor, and Hay 1997; Shaw and Gosling 1997; Taylor and Smith 1997; Clout 1999; Jaya, Moradb, and Bel 2003; Hewitt et al. 2004) only a single study has quantified the abundance of nonnative salt marsh plants as a whole (Wilson et al. 1996), and no studies have tested for impact of nonnatives on the distribution and performance of native salt marsh species. This high awareness of problems with nonnative invasive species is also reflected in the high rankings given by the regional managers (average 2.8 [2, 3, 3, 3, ?, 3]). Little is known about nonnative animals in salt marshes, but it is possible that introduced rabbits graze on salt marsh seedlings (Thomsen et al. 2005). Again, controlled experiments are called for, such as mammal exclusion and nonnative plant removals. CLIMATE CHANGE

Climate change is primarily associated with anthropogenic pollution with infrared-absorbing gases (International Panel on Climate Change [IPCC] 2001). Projected climatic changes include rising temperature and sea levels, increased evapotranspiration, higher frequency and/or intensities of storms, and altered precipitation patterns (Michener et al. 1997; Bell et al. 2001; IPCC 2001). Most of these effects are likely to stress salt marshes (although higher temperature and carbon dioxide levels may stimulate primary production; van de Staaij et al. 1993), for example, by increasing immersion and mangrove competition at low elevations and hindrance to upward expansion by stop banks, sea walls, causeways, and other urban structures (coastal squeeze). However, the overall community effects in specific marshes are expected to be complex and to depend on latitude, specific community structure, and other local environmental stressors (Vestergaard 1997; Donnelly and Bertness 2001; Simas, Nunes, and Ferreira 2001; Bertness and Ewanchuk 2002). In New Zealand, potential coastal climatic changes and impacts have been reviewed (relative sea-level

rise in New Zealand is expected to be 0.14 to 0.18 meter by 2050 and 0.31 to 0.49 meter by 2100), resources allocated, and research encouraged to provide information for future coastal management (National Science Strategy Committee 2000; Bell et al. 2001). On a more “salt marsh–positive” note, it is expected that increased salinization of lands around estuaries will transform terrestrial land into salt meadows (including previously reclaimed land), salt marshes, and eventually estuary mudflats, particularly in areas without seawalls or stop banks (planned retreat strategy; Bell et al. 2001). Again, no specific studies have yet related climatic changes to New Zealand salt marsh ecology. In addition to global warming, the degradation of the ozone hole is another humaninduced “diffuse” global climate change. The degradation of the ozone layer, mainly due to CFC emissions, has caused an increase in ultraviolet-B (UV-B) radiation, particularly in the Southern Hemisphere as the ozone layer is thinnest over Antarctica. Thus, New Zealand has high UV-B radiation (Howard-Williams et al. 1997; McKenzie, Connor, and Bodeker 1999), partly contributing to a high incidence of human skin cancer (about 1,800 melanoma cases and 45,000 nonmelanoma cases confirmed by laboratory tests, plus 20,000 nonmelanoma skin cancer cases treated without laboratory tests; Cancer Society of New Zealand 2004). There are no New Zealand studies that link high UV radiation to salt marsh ecology, but given quantified adverse growth effects on South American Sarcocornia plants (Bianciotto et al. 2003) and Dutch Elymus athericus (van de Staaij et al. 1993), we expect similar negative effects in New Zealand. Due to the complexity of the problems associated with multifactorial climate changes, it is obviously difficult to link these stressors to salt marsh performance (Caldwell and Flint 1994) and is therefore not surprising that the regional managers considered these threats likely, but generally with unknown consequences to local salt marshes (global warming average 2.5 [?, 2, 3, ?, ?, ?] and ozone average 2 [?, ?, 2, ?, ?, ?]).

AUSTRALIAN SALT MARSHES, MANAGEMENT, AND HUMAN THREATS SALT MARSHES OF AUSTRALIA

Unlike New Zealand, Australia is relatively flat, with low runoff. Rivers are mostly small, and a particular feature of southern Australian coasts is the number of intermittently open coastal lagoons (Brearley 2005). Northern Australia is tropical, with a strongly seasonal rainfall pattern with a pronounced summer maximum. Tropical coasts are exposed to major storms (cyclones) in mid- to late summer. In more temperate latitudes, rainfall is generally lower than in the tropics and more evenly distributed throughout the year, although in parts of southern Australia, there is a winter rainfallmaximum Mediterranean climate. There are extensive stretches of arid coastline, with no permanently discharging rivers, but where topography permits, there are salt marshes on these apparently inhospitable coasts. In northern Australia, the upper part of the intertidal zone often takes the form of extensive hypersaline flats, with vascular plants either very sparse or absent, and the sediment surface encrusted with microalgae and cyanobacteria (Saenger et al. 1977). The lower part of the intertidal zone in many mainland estuaries and sheltered soft open coasts is occupied by mangroves. The most species-rich mangroves occur in northeast Queensland, with species richness declining both westward and southward. At the highest latitudes, only a single species, Avicennia marina, occurs. There is also a marked decline in stature with increasing latitude. In northeast Queensland, mangroves may be up to thirty meters or more; but in southern Victoria Avicennia, they take the form of low shrubs about one meter tall. There are no mangroves in Tasmania. The total area of salt marsh in Australia is not known with any degree of certainty. Adam (1995) used the estimate presented by Bucher and Saenger (1991) of between thirteen thousand and fourteen thousand square kilometers (see also Bucher and Saenger 1989, 1994). The

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majority of this area occurs in tropical Australia (Queensland, the Northern Territory, and northern Western Australia). Bucher and Saenger (1994) point out that their inventory was not able to distinguish between “true salt marsh, with its flora of salt-tolerant herbs and grasses, and the bare saline clay pan which may have a seasonal plant cover.” The difficulty of distinguishing the two is affected not only by seasonal variation but also by the timing of image accession relative to tidal flooding and rainfall events. There are very extensive, frequently hypersaline, upper intertidal salt flats (a habitat referred to elsewhere as sabkha) in northern Australia, probably at least as extensive in area as the permanently vegetated salt marshes. It is not worth investing too much effort for purposes of inventory into distinguishing between salt marsh and salt flat as both are habitats of conservation value that are vulnerable to anthropogenic impacts. Indeed, it would be unfortunate if, in an attempt to protect salt marsh, developments were relocated to salt flats. Even in temperate Australia, determination of the area of salt marsh is affected by the spatial scale of investigation, the nature and quality of the images assessed, and differences in interpretation between operators. In NSW, the frequently cited figure for total salt marsh area of 5,716 hectares was derived by West et al. (1985) from air photo interpretation and ground truthing of some 130 estuaries; but a later study by Williams et al. (1998) identified some 950 water bodies discharging from NSW into the Tasman Sea, and they suggested that because of the dynamic nature of the NSW coast (as a result of both human and natural influences), “The actual number of water bodies present on the NSW coast may never be known.” More recently, Williams (2006) has highlighted substantial differences in both total area and number of recognized salt marsh sites between studies of the same estuaries utilizing different methodologies so that meaningful detection of trends is difficult.

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(A)

Unlike mangroves, in which species diversity decreases with increasing latitude, in Australian salt marshes, species richness is highest in Tasmania and lowest in the tropics. General accounts of the flora and vegetation of Australian salt marshes are provided by Saenger et al. (1977) and Adam (1990). More detailed studies are reported in Kirkpatrick and Glasby (1981); Bridgewater (1982); Adam, Wilson, and Huntley (1988); Cresswell and Bridgewater (1998); Kirkpatrick and Harris (1999); and Jaensch (2005). The pattern of regional variation in Australian salt marshes can be placed in the context of a global scheme (Adam 1990) and particularly in temperate latitudes, where there are similar patterns of species and generic occurrence across Gondwana continents (southern South America, South Africa, Australia, New Zealand) (Adam 1990). Most of the species or genera on temperate Australian marshes are also found in New Zealand, although there are some striking differences. The three species of Wilsonia (Convolvulaceae) are endemic to Australia, and the important grasses of southern Australian salt marshes, Sporobolus virginicus and Distichlis distichophylla, are notably absent from New Zealand. Thus, the native grasslands, which can be extensive in Australian salt marshes, are not a feature across the Tasman. The characteristic shrub of the upper marsh in New Zealand, Plagianthus divaricatus, does not have an obvious parallel in Australia.

FIGURE 18.5 Black swans (Cygnus atratus) can occasionally be found in salt marshes. Swans are common in Pauatahanui inlet, New Zealand, where they are often seen resting (and sometimes foraging) on Samolus repens patches (A). The numerous swan droppings in the marsh (B) provide a visible fertilization impact that likely facilitates marsh growth.

(B)

There are relatively few studies of fauna on Australian salt marshes. Tropical salt marshes provide habitat for the saltwater crocodile (Crocodylus porosus), which may be a disincentive for fieldwork! Macropods (kangaroos and wallabies) are obvious grazers at many sites, and on urban fringes salt marsh may provide an important refuge for macropods. The native water rat (Hydromys chrysogaster) and the false water rat (Xeromys myoides) both utilize salt marsh and mangrove habitats. A recent study has shown that temperate salt mashes are likely to be important foraging sites for bats (Laegdsgaard, Monamy, and Saintilan 2004). More study, including of tropical marshes, is required, but the previously noted findings strengthen the case for salt marsh conservation. Salt marshes provide important high-tide roosting for migratory wading birds, and a number of sites have been incorporated in conservation reserves for this reason. Unlike many Northern Hemisphere sites, Australian salt marshes are not visited by migratory waterfowl. Australian ducks are nomadic rather than migratory and are predominantly birds of the inland. The chestnut teal (Anas castanea) occurs in both

inland and coastal sites, but of Australian ducks, it is the species that shows a preference for coastal habitats including salt marsh, although it is rarely found in large flocks. Black swans (Cygnus atratus) frequent coastal lagoons, particularly during the molt, and feed primarily on sea grasses, but they can occasionally be found on salt marshes (fig. 18.5). The Cape Barren Goose (Cereopsis novaehollandiae) has a restricted distribution in southern Australia, but population numbers have been increasing, and locally it can exert heavy grazing pressure on vegetation, including salt marsh. One of Australia’s rarest birds, the orange-bellied parrot (Neophema chrysogaster), overwinters on the mainland on salt marshes, where it feeds on the seeds of chenopod (Cousins 1989; Garnett and Crowley 2000). A number of small passerine birds utilize salt marsh habitat including cisticolas (Cisticola juncidis, C. exilis) and chats (Ephianura albifrons and the very rare E. crocea var. macgregori). The invertebrate fauna of salt marshes includes both marine and terrestrial components, of which the marine element has been better studied, even though the total number

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of studies is small. A diversity of marine mollusks (Robinson and Gibbs 1982) and crabs (Mazumder et al. 2005a, 2005b) are characteristic of salt marshes. Given the distribution of salt marsh in the upper intertidal above the mangrove zone, and the paucity of creek and pan systems in most Australian salt marshes (Adam 1997), until recently the value of Australian salt marshes as fish habitat was assumed to be small. Several studies (Morton, Pollock, and Beumer 1987; Connolly, Dalton, and Bass 1997; Mazumder et al. 2005a, 2005b) have dispelled this myth. Although utilization by fish is temporally limited, a diversity of fish species, including a number of commercial importance, has now been shown to be present in salt marsh during flooding tides (Mazumder et al. 2005a, 2005b).

form of aquatic or marine reserve managed by fisheries agencies. Some sites are part of wetlands listed under the Ramsar Convention on Wetlands. Despite the diversity of state laws and policies affecting salt marshes, local government also has considerable influence, not only in approving (or refusing) some development within marshes but, more important, through control of development in the catchment. The research permit system vary between states. In NSW, salt marshes are generally classified as Endangered Ecological Communities, and a permit is needed from the Department of Environment and Conservation. In addition, if the research impinges on fisheries matters, a permit from the Department of Primary Industry would be required.

MANAGEMENT OF AUSTRALIAN SALT MARSHES

ANTHROPOGENIC THREATS

Tenure and regulatory control of Australian salt marshes is complex. In general, freehold and leasehold titles extend to high water (in most cases defined as mean high water springs [MHWS]), so that many salt marshes straddle the boundary between private and Crown land. There are, however, numerous exceptions where, because of historical accidents or inaccurate early surveys, private ownership continues below the current MHWS. Even where salt marsh is in private ownership, there may be controls on development applied through the planning system. For example, in NSW, all salt marsh is listed as an Endangered Ecological Community under the Threatened Species Conservation Act, and many sites are individually mapped under State Environment Planning Policy 14—Coastal Wetlands. The Fisheries Management Act in NSW and the Fisheries Act in Queensland have provisions that give the fisheries agencies considerable control over any development that may occur in salt marshes. Around Australia, a number of salt marshes are incorporated in conservation reserves under the control of national park agencies or in some

European Australia is essentially a coastal nation, the overwhelming majority of the population living on, or close to, the coast. All state capital cities are situated on estuaries (only the national capital, Canberra, is inland). This means that the coast and estuaries have been very heavily impacted by development (Turner et al. 2004). The majority of impacts has been in temperate and subtropical regions, the regions where salt marsh is more floristically rich, but where it would also have been relatively limited in extent. The much more extensive tropical salt marshes have received relatively little direct impact, although there are some tropical coasts where there has been extensive port and industrial development, most notably in northwest Western Australia (Turner et al. 2004), where in addition to development of ports and associated infrastructure to export iron ore, salt production facilities have resulted in loss of salt flats and salt marsh. In southern Australia, much of the loss of salt marsh has been through infilling for urban and industrial development (e.g., Fotheringham 1994). Unlike northern Europe, where the reason, over centuries, for much reclamation has

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DEVELOPMENT

been for agriculture, farming, although it has driven some reclamation, has not been such a major factor in Australia. From the 1970s onward, a major activity resulting in salt marsh loss, often in locations that at time of development were distant from major cities, was the development of canal estates (see illustrations in Turner et al. 2004; Brearley 2005). Many canal estate developments were controversial and viewed by many conservation groups as the epitome of the worst excesses of real estate entrepreneurs. In response to public concerns, planning regulations in NSW currently prohibit any new canal estates, but construction is still possible in other states (e.g., Western Australia; personal observation). Australia is a dry continent, so that management of water resources has been a major preoccupation of governments for the last two hundred years. For much of this period, concepts of ecological sustainability were not considered; even today, when environmental concerns are high, we still have the legacy and infrastructure of past decisions. Freshwater flows into estuaries have been affected both by dam construction and abstraction of groundwater. Tidal flow of saline water into upper estuaries and tributary creeks is often restricted through construction of weirs, floodgates, and other barriers. In NSW, Williams and Watford (1996) identified more than four thousand such barriers. Any areas of salt marsh cutoff from tidal influence by such barriers are likely to change to a more brackish vegetation type, although individual plants of more halophytic species such as Sarcocornia quinqueflora or Suaeda australis may survive at low density among tall Phragmites, Bulboschoenus, or Schoenoplectus for many years (personal observation), suggesting that removal of the barriers may permit reestablishment of salt marsh. Australia’s largest river system is the MurrayDarling, with a catchment covering 14 percent of the continent. Water from the Murray-Darling supports the population in the catchment and three-quarters of Australia’s irrigated farmland

(Turner et al. 2004). In addition, water is transported out of the catchment to supply Adelaide (the capital of South Australia) and the major industrial centers in South Australia. Flow to the mouth of the river has been reduced by about 80 percent compared to the pre-European period (Turner et al. 2004) such that the mouth is now frequently closed. As a result of barrage construction and reduced flow, the area of the estuary has been reduced by 90 percent. Not only is the volume of freshwater reaching the estuary reduced, but the timing of its arrival has been modified as a result of the management of release from upstream impoundments (Turner et al. 2004). The estuaries of all managed rivers will have experienced modified freshwater inputs, although less dramatic than those in the Murray. The impacts on salt marshes have, however, been little studied. Although inputs of water into estuaries may have been reduced, within urban areas or adjacent to major roads, inputs into individual marshes may have increased due to storm water discharge. Except in the most recently constructed examples, there are rarely gross pollutant traps. Storm water is thus a source of a range of pollutants, including heavy metals, oils, pesticides, and larger plastic pieces and other rubbish. Discharge also affects the salinity regime within the marsh. In many locations, storm water pipes are associated with the establishment of patches of Phragmites or Typha within marshes (see Zedler, Paling, and McComb 1990). Spread of Phragmites changes the structure and diversity of salt marshes, but, unlike the situation in New Zealand (McCombs 2004), it is not generally treated as a weed in Australia. EUTROPHICATION

The Australia landscape has been transformed over the last two hundred years by the development of agriculture. On a continent of largely nutrient-poor soils, this has involved extensive use of fertilizers. While the nutrient loads now entering estuaries have undoubtedly increased,

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any consequences for salt marshes have been little studied. The most dramatic impacts have been in southwestern Western Australia, particularly in the Peel-Harvey Estuary (Deeley and Paling 1998; Hodgkin and Hamilton 1998; Brearley 2005), where, as a result of considerable increases of nutrients but particularly of phosphorus, algae productivity greatly increased, smothering fringing salt marshes and, on dying, producing a malodorous mess that substantially reduced the amenity of the area. This problem has been addressed not only through education and better practices of farmers in the catchment, but by major engineering works to construct a new opening to the estuary (the Dawesville Cut), altering the circulation and flushing patterns (Hodgkin and Hamilton 1998). In addition to fertilizer runoff, there may be runoff of agricultural chemicals, but, again, there has been no study of possible impacts on salt marsh biota. OIL POLLUTION

Salt marshes are potentially at risk from oil or chemical spills. These spills could contribute to low background levels of pollution (e.g., from routine port operations and refueling) or be a consequence of major accidents. In addition to shipping accidents, road or rail transport accidents resulting in oil or chemicals entering the storm water system could impact salt marshes. Australia has been fortunate to date in not having suffered major incidents affecting salt marshes. In Botany Bay, where Sydney’s major oil refinery is situated, there have been a number of shipping incidents resulting in minor (by world standards) oil spills that have had adverse impacts on mangroves (Anink et al. 1985; Allaway 1992), although tidal conditions at the time of the incidents meant that oil did not extend into salt marshes. Australia has comprehensive contingency plans for responding to oil spills (Australian Maritime Safety Authority [AMSA] 1993) that recognize the importance and sensitivity of salt marsh. Booms, skimmers, other equipment, and chemicals are available for rapid deployment in the event of any spill. 378

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Shipping operations present other potential hazards to salt marshes, including the possibility of introduction of exotic species in ballast water. Although many ballast water introductions have been recorded in Australian waters, no species that might have a direct impact on salt marshes has been reported so far. Australia is active both nationally and internationally in moving to reduce to overall environmental impacts of shipping (Australia and New Zealand Environment and Conservation Council [ANZECC] 1996). HEAVY METALS

Australia’s major industrial centers are located close to estuaries, and as a result, there have been considerable increases in the amounts of bioavailable heavy metals in estuaries. (This has arisen from atmospheric fallout, runoff, and, in some cases, deliberate discharge and dumping.) However, while heavy metals in sediments have been documented (Chenhall et al. 2004), and uptake by salt marsh plants has been described (Chenhall, Yassini, and Jones 1992), the ecosystem consequences have not been explored. The various state and national environmental protection agencies have established guideline concentrations for a range of pollutants in estuarine and marine waters (ANZECC and Agricultural and Resource Management Council of Australia and New Zealand [ARMCANZ] 2000), largely based on human health considerations, with little knowledge about the responses of native biota. Chenhall et al. (2004) showed that surface sediments in Lake Illawarra had higher levels of heavy metals, reflecting anthropogenic inputs. However, under the ANZECC and ARMCANZ (2000) guidelines, most sites were classified as low risk, although there were some “hot spots” with much higher values. In Lake Macquarie (another large coastal lagoon in NSW), zinc and lead levels very much higher than those in Lake Illawarra have been recorded at Cockle Creek (Roy and Crawford 1984; Batley 1987), associated with smelter operation. Similar stories could be told for other long-established

industrial regions, but many of Australia’s estuaries do not have heavy industry. INVASIVE SPECIES

The salt marsh flora of Australia contains a large number of introduced species, although probably not as high a percentage as in New Zealand. Many of these are found in the upper salt marsh fringe, particularly on dry sandy coasts (Bridgewater and Kaeshagen 1979); and although there have been no detailed studies, they do not appear to pose a serious threat to the integrity of the ecosystem. A few more widespread species give rise to more serious concern. As in New Zealand, Spartina anglica is of major concern in Victoria and Tasmania. The history of the introduction of Spartina to Australia was documented by Boston (1981). In Tasmania, the original plantings were for mudflat stabilization (Phillips 1975; Pringle 1982) and from that perspective were a great success. In Victoria, a particular concern is that at some localities it has spread onto mudflats below the previous seaward limit of mangroves. Control of Spartina is now actively practiced, and the basis for control measures are discussed in Rash, Williamson, and Taylor (1996) and Kriwoken and Hedge (2000). The primary control technique in Tasmania is application of herbicide (Fusilade®), supplemented by physical removal and smothering (Tasmanian Department of Primary Industries, Water, and Environment 2002). In southern Queensland and continuing to spread southward into NSW, a major salt marsh weed is groundsel bush (Baccharis halimifolia; Natural Resources, Mines, and Water 2006). An added incentive for the control of this species is its importance as a trigger for hay fever. Juncus acutus can displace the native Juncus kraussii and form dense monospecific stands in upper and mid–salt marshes (Paul and Young 2006). Pampas grass, Cortaderia selloana, is both salt-tolerant and recovers vigorously after fire (personal observation). In NSW, it occurs in a number of salt marshes and has considerable potential for further spread.

A particular problem for the control of weeds in salt marshes is that most herbicides are not routinely approved for use in the intertidal environment, and the research necessary to ascertain toxicity of herbicides (and any accompanying surfactants) to intertidal biota and to the sensitivity of native plants to herbicide application would be expensive and is unlikely, given the very limited sales market, to be carried out by manufacturers. CLIMATE CHANGE

Except where sea cliffs occur, the Australian coast will be vulnerable to a greenhouse-induced rise in sea level (Short 1988), and salt marshes are likely to be particularly sensitive (Vanderaee 1988). As elsewhere, “coastal squeeze” is likely to occur if the seaward edge of marshes retreats, but the landward boundary is constrained by either natural topography or artificial structures. The widespread spread of mangroves into salt marshes in southeastern Australia (Saintilan and Williams 1999) may in part be an early indication of sea level rise. Sea-level rise will be not the only consequence of the greenhouse effect. Increased carbon dioxide and temperature are likely to alter the performance of individual species, changing the composition of assemblages. Changes in rainfall (both increases and decreases) may have impacts on the composition of upper marsh communities, and changes to the frequency and intensity of cyclones could result in extreme disturbance of some salt marshes in northern Australia. Coastal squeeze will be a particular problem on the developed coastline of southern Australia. In the north, there are long stretches of coast where landward shift may be possible, but the composition of the marshes may still change in response to changes in temperature, carbon dioxide concentration, and freshwater inputs. FIRE

Australia is a notoriously fire-prone continent. Salt marsh would not, however, be generally considered likely to burn. Nevertheless, upper salt marsh communities dominated by Juncus kraussii or

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Baumea juncea can carry hot fires. Documentation of fires in salt marshes is poor, and recovery has been little studied, although, anecdotally, vegetative regeneration is slow (as after the 1994 fires that burned both mangroves and salt marsh on the Hawkesbury River; personal observation). In tropical northern Queensland, fire is used as a management tool in Sporobolus virginicus grasslands to promote “green pick” for grazing cattle (Anning 1980). Kirkpatrick and Harris (1999) suggest that fire in Tasmanian salt marshes may have caused a decline in the shrubby chenopod Sclerostegia arbuscula. MOSQUITO CONTROL

Mosquito control is becoming an increasingly important issue in the management of Australian salt marshes. A number of mosquito species breed in salt marshes, including Ochlerotatus vigilax, Anopheles hillii, and Culex sitiens, which can act as vectors for a number of arboviruses causing debilitating diseases of humans (Webb and Russell 2006). Currently, the major diseases transmitted in this way are Ross River fever and Barmah Forest disease, and under global warming, their incidence is expected to increase. There is also the potential for malaria to become established. Given the increasing human population living within close proximity to estuaries, there is great pressure on local councils to implement control measures. Although pesticides (Webb and Russell 2006) are still widely used, there is public concern about the possible consequences, and alternative nonchemical approaches have been developed. Runneling is the creation of shallow drainage channels to increase tidal flushing and remove standing water, which is the mosquito larval habitat. Runneling has been widely practiced in Queensland (Dale 1994; Jaensch 2005) and has also been instituted in southwestern Australia (Latchford 1998). Studies to date indicate that the ecological impact of runneling on nontarget species is slight, although whether this would be true if mosquito control in more species-rich higher latitude marshes were required is not known. 380

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SALT PRODUCTION

Australia is one of the world’s leading producers of solar salt. The majority of production is in northwestern Australia, but plants have also been developed in southwestern Australia, Victoria Queensland, and South Australia. Evaporation ponds have been constructed on high-level salt flats and on salt marshes (and this has been a major factor in the decline of South Australian salt marshes; Edyvane 1996). Further expansion of the industry could threaten the persistence of salt marsh at some sites.

OFF-ROAD VEHICLES

A major threat to many Australian salt marshes is from off-road recreational vehicles (BMX bicycles, trail bikes, and four-wheel drive vehicles). There is a variety of consequences of passage of vehicles across marshes, including crushing of vegetation and alteration to drainage patterns, with consequent changes to habitats. Examples of damage are illustrated in Jaensch (2005) and Kelleway (2005). The scale of the problem is increasing. Kelleway (2005) has shown that in 1966, the proportion of salt marsh affected by track damage in the Georges River (NSW) was 1.5 percent, but by 1998, this figure had increased to 23.2 percent, notwithstanding that some of the sites were in a national park. Tracks may persist for many years, even when access is prevented.

AQUACULTURE

Aquaculture is a major growth industry internationally. There are a range of possible impacts on salt marshes, directly from loss of sites for construction of ponds, and indirectly from changed water quality from effluent. A number of early aquaculture facilities in Australia involved loss of salt marsh (e.g., construction of prawn ponds on Micalo Island in NSW; personal observation), but the protection now afforded salt marsh through the planning systems makes it unlikely that similar facilities would be approved.

LIVESTOCK GRAZING

Livestock have access to many Australian salt marshes; particularly in the tropical north, salt marsh is regarded as an important resource for rangeland cattle (Anning 1980). In some locations, sheep have access, but there is no practice of intensive grazing by sheep, as was the case in much of northern Europe. Indeed, one attempt to intensively graze sheep, on the Towra Point salt marsh in Botany Bay, was a complete failure, resulting in the burial of carcasses on the site (Holt 1972). In Europe, grazing profoundly changes the composition of salt marsh vegetation. There is little evidence of such impacts in Australia, and in species-poor tropical marshes, there are too few species to ring the changes. Howarth (2002) has made the interesting suggestion that past cattle grazing may have been a factor limiting the invasion of salt marsh by mangroves. Trampling damage is visible in the marshes on which cattle grazing occur.

CONCLUSION AND RECOMMENDATIONS Few quantitative data exist that directly link anthropogenic threats to Australasian salt marshes. Nevertheless, it is certain that large salt marsh areas have been drained and cleared and that many remaining marshes are under threats from an array of anthropogenic stressors. A lack of quantitative distribution data, such as species-specific biomass/cover values with standard errors, implies that it is difficult to evaluate past and future changes (Gurevitch and Hedges 1999; Osenberg et al. 1999). A further lack of process-oriented manipulative experiments also implies that we have poor knowledge about the processes that control present distribution patterns. These shortcomings cause conservation and restoration effort to some extent to be based on anecdotal and personal experiences and the gray literature (local reports). We recommend that managers combine remote sensing techniques with ground truth surveys (species identification, stem length, biomass, with standard errors) to establish

baseline data on the extent, zones, and borders of salt marshes. Preferably, data should be added on salt marsh fauna, sediment characteristics, and degree of pollution/anthropogenic impacts. Such data should be digitized and entered into a standardized database to allow for easy cross-regional comparisons and for efficient documentation of future large-scale changes (many of these suggestions are currently being implemented in regional councils in New Zealand). It is equally important that research institutions support baseline monitoring data with process-oriented manipulative experiments that test for effects of anthropogenic stressors on salt marsh plant health, community patterns, and ecosystem functioning. For example, the most abundant and ubiquitous plants and animals should be manipulated (inclusion/exclusions) to uncover mechanisms of zonation patterns and assess community impacts, and multiple stressors should be manipulated simultaneously using factorial design to elucidate relative importance of stressors as well as to test for interaction effects (e.g., Bertness 1984, 1985; Bertness and Yeh 1994; Silliman and Zieman 2001; Bertness and Ewanchuk 2002; Silliman and Bertness 2004). Lastly, results obtained from these tasks, should be published in peer-reviewed journals to ensure high-quality data and open access for researchers, and to avoid valuable information remaining hidden in local gray literature. The threats facing Australasian salt marshes are similar to those operating throughout the world (Adam 2002). Increased awareness of the ecological values of salt marshes has led to the introduction of measures to prevent actual loss of sites. These are likely to be effective in some circumstances, but the controls are rarely absolute and have escape clauses to permit development where the economic value of a project is perceived to be greater than the loss of environmental value, or where the national interest is thought to be best served by development. Further losses for major infrastructure, such as roads or ports, are likely to occur. However, any such development will probably

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include mitigation measures, such as rehabilitation of other damaged marshes or creation of new marshes. Research on restoration techniques is in its infancy in Australasia, but it will become increasingly necessary as the continuing population increase causes growing stress on coastal salt marshes. This highlights the need for strong management. Part of the solution will involve development planning and public education, but control of insect populations is likely to be called for (particularly in Australia). Given political reality, such calls will be heeded, but hopefully the methods used will minimize collateral damage to salt marsh and estuarine ecosystems. At the site-specific level, there is likely to be a tyranny of small decisions, causing attrition of the salt marsh resource. A boat ramp here, a coastal cycle track there are likely to be seen as socially desirable and having minimal environmental impact, although the cumulative impact of many such decisions could be considerable. Decisions made by individuals to carry out activities such as illegal dumping or using off-road vehicles in salt marshes could also have large cumulative impacts. Given increased visitation to northern Australia, damage from off-road vehicles could increase stress on large tropical salt flats and salt marshes significantly. Addressing these individually minor but cumulatively large impacts, both approved and illegal, will require heightened public awareness and action. Management of changes to hydrology and of pollution, both point source and diffuse, will require action at the catchments scale. The major challenge will be to address the continuing consequences of historic decisions and actions. The effects of human modification of the atmosphere will be felt by salt marshes globally. The affects include global warming, changes to patterns of storminess, rainfall and fire, increases in atmospheric carbon dioxide concentration, and increase in UV radiation due to a reduction in upper atmospheric ozone. The consequences are difficult to predict with certainty due to complex interaction between 382

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the various affects. In southern Australia and many estuaries in New Zealand, sea-level rise is likely to create coastal squeeze, resulting in salt marsh loss. In Australian tropical north and remote low-populated New Zealand estuaries, there will be greater opportunities for salt mash area to be sustained by landward retreat. Changes in temperature, rainfall, and storm occurrence and an increase in carbon dioxide will give rise to species-specific response, so that future salt marsh communities may have different compositions from those of today. Particularly in southern Australia and northern New Zealand, there is likely to be a continuance of the trend in the expansion of mangroves into salt marsh (Saintilan and Williams 1999). In short, salt marshes in New Zealand and Australia are, like salt marshes throughout the world, under continued pressure as coastal human populations and economies continue to increase. However, with heightened attention on monitoring, regulation, planning, protection, research, and restoration, it should be possible to ensure that these important sea–land transition ecosystems remain abundant and healthy. Acknowledgments. We greatly appreciate questionnaire replies from New Zealand regional councils: P. Grove, Environment Canterbury; T. Dwane, Northland; F. Maseyk, Horizon (Manawatu/Wanganui); M. Felsing, Waitako; S. Park, Environment Bay of Plenty; and B. Tikkisetty, Environment Southland. REFERENCES Adam, P. 1990. Saltmarsh Ecology. Cambridge: Cambridge University Press. ———. 1995. Saltmarsh. Pages 97–105 in L. P. Zann and P. Kailola (eds.), The State of Marine Environment Report for Australia. Canberra: Department of Education Science and Training, Technical Annex 3, State and Territory Issues. ———. 1997. Absence of creeks and pans in temperate Australian salt marshes. Mangroves and Salt Marshes 1: 239–241. ———. 2002. Saltmarshes in a time of change. Environmental Conservation 29: 39–61. Adam, P., N. C. Wilson, and B. Huntley. 1988. The phytosociology of coastal saltmarshes in New

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CONCLUSION

Salt Marshes under Global Siege Brian R. Silliman, Edwin D. Grosholz, and Mark D. Bertness Salt marshes are under siege from human disturbance on a global scale. Salt marsh coverage as well as the structure of these ecosystems continues to deteriorate drastically due to human-induced changes. The critical ecosystem services these systems support are likewise endangered. No longer can marshes be viewed in scientific, conservation, social, and political circles as one of the most resilient and resistant ecological communities. And no longer can they be championed as systems that can and should be used to buffer human impacts (e.g., absorption of nutrients in waste water and terrestrial runoff). These systems are in desperate need of protection from human influence. In this volume, we provide in-depth case summaries of how humans have modified salt marshes through both direct and indirect activities. Most ecologists and marsh managers realize that the largest historical human threat to salt marshes has been reclamation activities. In many cases, when reclamation threats are abated, salt marshes are thought to then be protected. This book dispels that myth and reveals that salt marshes are currently threatened by an impressive portfolio of human-generated threats, many of which are currently underestimated or even overlooked by coastal conservation managers. These threats include human-precipitated species invasions (Strong and Ayres, chap. 1; Grosholz et al., chap. 2; Byers, chap. 3; Meyerson et al., chap. 4; Keddy

et al., chap. 7; Bertness et al., chap. 8; Davy et al., chap. 16; Costa et al., chap. 17; Thomsen et al., chap. 18), small- and large-scale eutrophication and accompanying plant species declines (Henry and Jefferies, chap. 5; Bertness et al., chap. 8; Crain et al., chap. 9), runaway grazer effects that denude marsh substrate (Henry and Jefferies, chap. 5; Silliman et al., chap. 6; Keddy et al., chap. 7), climate change–induced effects including sea-level rise (Stevenson and Kearney, chap. 10), increasing air and sea surface temperatures, increasing CO2 concentrations (Mayor and Hicks, chap. 11), altered hydrologic regimes (Crain et al., chap. 9; Davy et al., chap. 16; Costa et al., chap. 17; Thomsen et al., chap. 18), and a wide range of pollutants including nutrients, synthetic hormones, metals, organics, pesticides, and so forth (Davy et al., chap. 16; Costa et al., chap. 17; Thomsen et al., chap. 18). We have provided a detailed mechanistic understanding of how humans generate these threats, how these threats impact marsh structure and function, and, importantly, clear recommendations to managers on what should be done to abate these threats. If these steps are not taken by conservation practioners, the long-term persistence of salt marshes and the services they provide is severely threatened. Specifically, without proper conservation action, we predict this key coastal community will become a nonsignificant habitat in less than one hundred years.

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THE STATE OF NORTH AMERICAN SALT MARSHES THE SPATIAL STATE OF OUR MARSHES

The state of our marshes is not good! Compared to the total amount of marsh on the North American coastline in precolonial times, we have lost a tremendous amount of this coastal key habitat. What remains is largely concentrated in high or increasing population areas in the United States or subject to intense geese grazing in Canada. Compared to four hundred years ago, only about 8 percent of marshes remain along the entire Pacific coast, 62 percent along the U.S. North Atlantic, 82 percent in the Gulf of Mexico, 88 percent in the U.S. South Atlantic, and 36 percent in maritime Canada and Hudson Bay (Bromberg and Silliman, chap. 13). Most of these losses are historical and primarily the result of reclamation on the Northeast and West coasts (Bromberg and Silliman, chap. 13; Callaway and Zedler, chap. 15), reduction in sediment loads reaching the Gulf of Mexico, and runaway goose grazing in Hudson Bay (Henry and Jefferies, chap. 5). Despite contemporary abatement of large-scale reclamation projects, the remaining salt marshes in North America are now threatened by a much broader range of human-induced forces. THE THREATS AND STATUS OF REMAINING MARSHES EAST AND GULF COAST SALT MARSHES

One of the most overlooked human impacts to East and Gulf coast marshes is humanprecipitated runaway grazing. The summaries of snail, nutria, and goose grazing in this volume clearly put to rest the sixty-year-old bottom-up paradigm of marsh ecology and conservation that consumers are unimportant in regulating marsh plant growth. In subarctic Canadian marshes, increasing populations of lesser snow geese that breed in coastal areas have severely damaged wetland vegetation, particularly in the Hudson Bay Lowland (Henry and Jefferies, chap. 5). This damage has led to exposure of peat and sediment to intense desiccation stress and conversion of North America’s 392

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largest salt marsh to an unvegetated, salty mudflat. The observed increases in bird numbers are tightly linked to their use of agricultural crops as a subsidized food source duration their long migration events. Recent experimental work in Georgia and Louisiana suggests a similar potential fate for southeastern marshes. These findings demonstrate that marsh grass growth is indirectly facilitated by blue crabs, which keep densities of potent snail grazers in check. Thus, the overharvesting of blue crabs may contribute to runaway grazer effects in southern U.S. marshes. When snail densities have increased locally, due to drought stress and/or predator depletion, concentrated fronts of snails have formed (more than one thousand individuals per square meter, and more than two hundred meters long and two meters wide) and acted as major contributing factors to the massive community die-off event of 2002–2004 (Silliman et al., chap. 6). Correlation work examining nutria and alligator population dynamics similarly suggests that human overhunting of alligators led to density increases in nutria in Gulf Coast estuaries, which in turn led to intensified grazing on salt marsh grasses. Although nutria numbers are now decreasing as alligator populations rebound in marine environments, this putative trophic cascade has powerful implications for marsh structure and function, especially while alligator populations remain threatened (Keddy et al., chap. 7). In New England salt marshes, seemingly the last bastion of bottom-up-only control along the entire northwestern Atlantic coastline, unpublished work clearly shows that sesarmid crab populations have exploded in Cape Cod, Massachusetts, marshes and are significant contributors, if not the ultimate causal agents, of widespread marsh plant die-off in these areas (M. D. Bertness et al., unpublished data). If conservation practitioners fail to effectively manage these potent grazers and base conservation strategies solely on the old bottom-up paradigm, as they have in the past, then our efforts to conserve New England marshes will likely be severely limited and could even fail.

Invasive species are also a problem in these marshes, but the intensity of these threats varies greatly regionally. Problems associated with the invasive plant, Phragmites australis, are most intense in the northeastern United States, where widespread development at the edges of small marshes has led to increased invasion succession via eutrophication (Meyerson et al., chap. 4; Bertness et al., chap. 8). Phragmites is beginning to invade mid-Atlantic, southeast, and Gulf coast marshes and more northern ones in Maine, but its success so far in those areas is mostly confined to brackish and freshwater marshes. In Maine, where salinities are lower, invasion into full-strength seawater marshes is likely inevitable where the woody buffer surrounding a marsh has been removed. In southern marshes, the picture is not so clear. Here, soil salinities are much higher, which will likely retard Phragmites invasion; native plants are much larger and more robust in comparison to their northern counterparts, which could make southern salt marshes more resilient to Phragmites takeover; and southern marshes are much more expansive, making adjacency effects due to woody buffer removal much less intense. Animal invasions into these marshes are much less well studied. However, European green crabs, Japanese shore crabs, porcelain crabs, and various snails and whelks are now being reported to be major components of the marsh food web at low elevations. How these animal invasions will impact marsh structure and function is not well known. Again, New England and the mid-Atlantic marshes are the center of these invasions. Expansion into adjacent regions, however, may occur. Thus, the potential to do so and the impacts on local marsh communities should be studied. Human modification of terrestrial runoff and hydrologic flow regimes are major issues in all regions, save for subarctic marshes in Canada. The impacts of these activities and remedies needed to abate them vary greatly from region to region. In New England, dams, bridges, and drainage ditches built years ago have altered the hydrology of most marsh

systems. Freshwater flow into these areas has been diminished significantly as well (Crain et al., chap. 9). Restoration of historic flow regimes, connectivity, and drainage patterns is now underway but should be increased in scale and intensity. In the mid-Atlantic, sedimentation and nutrients reaching estuaries and salt marshes have increased as woody buffers along waterways throughout the watershed have been removed for farming and development (Meyerson et al., chap. 4; Bertness et al, chap. 8). While the same pattern is true for Gulf Coast watersheds, the extent of woody buffer conversion is much less. In addition, increases in nutrient and sediment loads are not as intense. Indeed, due to channelization of Gulf rivers such as the Mississippi, which are no longer allowed to meander back and forth and spread their sediment across the delta, these marshes are deprived of life-lifting nitrogen and sediment that would otherwise offset the natural and human-induced (due to groundwater and oil drawdown) subsidence of the Gulf coastal shelf. As a result, Gulf Coast marshes are drowning at an alarming rate (Stevenson and Kearney, chap. 10; Bromberg and Silliman, chap. 13). In the Southeast, runoff is certainly a problem generating unwanted anoxic events in estuaries, but because of increased incidence of climate-induced droughts and increasing coastal populations, less freshwater is reaching these salt marshes. This decreased input of freshwater will likely increase overall stress by increasing marsh soil salinities, making marsh plants more susceptible to drought and associated runaway grazing by snails (Silliman et al., chap. 6). Managers who regulate coastal water discharge will thus have to strike a balance between the need to decrease nutrient input in coastal estuaries and the simultaneous need to increase stress-ameliorating, freshwater fluxes into shallow-water coastal ecosystems. Lastly, all marshes throughout this region will be exposed to the threats of human-induced climate change. However, each region is likely to respond differentially, and field experiments are needed to generate predictions of

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latitude-specific marsh responses. For example, warming may occur over the entire northwestern Atlantic coastal environment, but in New England salt marshes, this may result in relieving waterlogging stress and increasing plant productivity; while in southern marshes, warming could lead to increased evaporation and increased salt stress, decreasing plant growth. As for sea-level rise, there will be differential responses in marsh coverage along all coasts depending on: (1) local marsh accretion rates, (2) the slope of upland borders, (3) the presence of shoreline hardening and other coastal development, and (4) uplifting or subsiding rates of coastal land. Some marshes will be able to cope with the sea-level rise threat, while others will drown (Stephenson and Kearney, chap. 10). Most important, marshes without room to migrate landward in response to increasing sea levels will likely decrease in size and ultimately disappear. Local managers will have to gather all related information about their specific marshes and generate predictive models to best formulate effective measures to abate these climate change–induced threats. WEST COAST MARSHES

The threats to western salt marshes largely follow from the geomorphological limitations of salt marsh habitat and the extent of urbanization. Broadly speaking, marshes in the southern portions of the Pacific Coast including California and northern Mexico (Baja California) were historically limited by the spatial extent of appropriate geomorphology. Most salt marshes were limited to the estuarine areas at the confluence of river mouths and oceans. The smaller and more limited extent of rivers in southern areas results in increasingly smaller and more isolated marshes in the southern distribution (Callaway and Zedler, chap. 15). In addition, extensive urbanization in California and northern Baja California, including upland areas adjacent to salt marshes, has resulted in increasing habitat loss and degradation. Urbanization has brought with it bacterial and eutrophic inputs as the result of nearby human 394

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occupancy (Callaway and Zedler, chap. 15). In addition, the proximity of large cities, including San Francisco, San Jose, Los Angeles, San Diego, and Tijuana, to marshes in this region has carried with it a legacy of pesticides, fertilizers, metals, organics, and other pollutants that require removal and subject many marshes in this region to persistent toxic stress (Callaway and Zedler, chap. 15). The extensive loss and fragmentation of habitat in the southern Pacific region has also had several consequences for the health of salt marshes in this region. The lack of adjacent marshes results in a dearth of recruitment sources for plant seeds as well as fish and invertebrate propagules, thus slowing the rate recovery of salt marsh habitats following the creation or restoration of salt marsh habitat. Habitat loss and fragmentation as well as urbanization have also led to important endangered species issues in the southern Pacific region. Several obligate salt marsh and estuarine species in this region are now on the federal endangered species list, including the California clapper rail (Rallus longirostris obsoletus), light-footed clapper rail (Rallus longirostris levipes), salt marsh harvest mouse (Reithrodontomys raviventris), the halophyte “soft-bird’s beak” (Cordylanthus mollis ssp. mollis), and the tidewater goby (Eucyclogobius newberryi). The altered hydrologic regimes of southern Pacific estuaries have greatly constricted the natural opening and closing of bar-built estuaries that often close seasonally as outflow declines and sand bars occlude the openings. Diversions of surface water and extraction of groundwater from upland watersheds have contributed to the loss of outflow and the more extensive, earlier closing and later openings of these estuaries. The loss of ocean inputs can dramatically influence the salinity regimes, resulting in broad-scale fluctuation in plant and animal communities. Many of these threats in the southern Pacific region are of less concern for marshes in the northern Pacific region of Oregon, Washington, Canada, and Alaska, where estuaries were

historically much larger and more connected because of larger and more numerous river systems and more appropriate coastal geomorphology. These areas are also less urbanized, and generally the extent of marsh loss has been much less. In these regions, the extent of human impacts is more diffuse and less the result of urban inputs. However, the inputs from surrounding watersheds are generally much greater, since river systems here are larger and flow more consistently through the year; consequently, sediment inputs from deforested landscapes continue to pose a significant threat. In addition, northern salt marshes experience high levels of resource extraction, including commercial and sport fin and shell fishing, that can and may impact salt marsh ecosystem health. Invasive species, particularly invasive plants, continue to be a serious threat to maintaining ecosystem services in salt marshes throughout the entire Pacific region. Species in the genus Spartina have invaded estuaries throughout the West Coast and are in the process of dramatically changing these systems (Strong and Ayers, chap. 1; Grosholz et al., chap. 2) as they have in other continents (Strong and Ayres, chap. 1; Davy et al., chap. 16; Thomsen et al., chap. 18). The ongoing introduction of new species into these systems represents one of the primary concerns for salt marsh conservation in this region. Finally, like marshes in other regions of North America, West Coast marshes will clearly be subject to climate changes, including sealevel rise. But because of the greater, relative inputs from surrounding watersheds, northern Pacific marshes may be subject to greater changes in rainfall and subsequent hydrologic regimes. GLOBAL PERSPECTIVE

Salt marshes have attracted human settlement since the beginning of human history. Mesopotamian salt marshes were the cradle of civilization. In Europe, human manipulations of salt marshes began thousands of years ago as the ice shields receded. Today, one is hard-pressed to

find salt marshes in Europe that have not been extensively impacted by damming, drainage, grazing, or reclamation (Davy et al., chap. 16). For millennia, salt marshes were seen as land that needed to be reclaimed for human use. Europeans attacked this challenge head-on. Their success in this endeavor is best exemplified by the extent of Dutch marsh reclamation and establishment. Astonishingly, over 40 percent of the country is on former salt marshes and freshwater wetlands, and all the marshes now in the Netherlands were generated as a by-product of dam and dike projects (Davy et al., chap. 16). Thus, no pristine marshes remain. During the colonial era, the tradition of reclaiming the “wasted” resources of salt marshes followed Europeans around the world. In North America (Parts I through IV), South America (Costa et al., chap. 17), and Australia and New Zealand (Thomsen et al., chap. 18), human use of salt marshes began soon after European colonization. Prior to this era, indigenous people used marshes for hunting and gathering shells and grass thatch for domiciles. Indigenous people impacted marsh food webs, as when the Maori extinguished eleven species of flightless birds in New Zealand (Thomsen et al., chap. 18), but there was no tradition of destroying marshes. European colonists, however, harvested grasses and grazed cattle in marshes; filled marshes for harbors, towns, and farms; and manipulated marsh hydrology to increase plant productivity and make waterways more navigatable. In North America, these traditions have only subsided in the last twenty-five to one hundred years. Today, the starkest contrast between human impacts in the Old and New World salt marshes is not type but extent. In New England, most salt marshes have been lost to reclamation, and those that remain are severely reduced by human threats, such as invasive species and shoreline hardening. In the northwest Pacific and the southern and Gulf coasts of the United States, the situation is much different. Considerable salt marsh lands remain (Bromberg and Silliman, chap. 13), but they are threatened by land use

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(e.g., reduced freshwater and sediments, increased nutrients), climate change (elevated temperatures and sea-level rise), runaway grazing, and their interactions. South American marshes are also under siege by transplanted European land use traditions, but there is still time to save them, since human domination of these systems is far from complete.

OUR CONCLUSION AND OVERARCHING RECOMMENDATION Because ecologists and conservation biologists now widely recognize that human-generated threats to ecosystem structure and function are overlapping, salt marsh conservation practitioners must now think outside the box and use measures and models that deal with multiple, co-occurring, and potentially synergistic threats (The Nature Conservancy, for example, now uses this type of approach). This new approach to marsh conservation must also be regional in scale. Historically, there has been a scaling mismatch between local, or “one marsh at a time,” conservation measures and the reality of regional-scale, multimarsh distribution of human-induced impacts—the reality of which we have highlighted here. To save the marshes we have left, this old framework of marsh conservation and the outdated dogma of salt marsh ecology must be changed to reflect new trends in general conservation strategies and the new science that challenges our old way of thinking about these systems (e.g., incorporating topdown control and food web interactions). State and federal agencies can no longer act independently and at local scales. Here, we provide the first intellectual map highlighting variation in spatial distribution and intensity of anthropogenic threats and thus provide the first step needed to begin formation of multiregional and multithreat-based conservation strategies for North American salt marshes. This new management framework is especially important on the West Coast because of the high degree of marsh isolation. Conservation of existing habitat and restoration of degraded habitat must 396

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take into consideration the distribution of nearby marshes for sources of propagules, links for migratory species and endangered species, management of invasive species, and so forth. One of the most important and effective acts that conservation practioners can begin to do to ensure the long-term protection and persistence of salt marsh habitats is to champion the use of marine protected areas in marsh management. These protected areas must (1) include associated marine habitats, such as seagrass beds and oyster reefs; (2) incorporate extensive areas of undisturbed terrestrial border to buffer marshes from excessive eutrophication via runoff and allow for their landward migration as sea level rises; (3) account for the inclusion of positive interactions (Halpern et al. 2007) at all levels of biological association (e.g., between species—trophic cascades; and across ecosystems—nursery benefits); and (4) be large, numerous, and appropriately spaced (see Halpern et al. 2007 for discussion). Around the world, coral and rocky reef conservation practioners and scientists lead the field of marine conservation in this effort. Salt marsh conservationists and ecologists are far behind this work and thus should look to these fields for lessons learned and guidance when establishing marine protected areas for temperate coastal areas whose intertidal zone is dominated by salt marshes. Because of the conservation prestige associated with the designation of a site as a marine protected area, using this method as a means to preserve marshes will also raise public awareness as to the critical role marshes play in the ecology and economy of local human communities.

ECOLOGICAL AND CONSERVATION LESSONS LEARNED 1. Marshes are ideal systems to study and test ecological theory underlying human impacts because they are simple systems, easy to manipulate, and strongly influenced by humans. Thus, marsh systems can and should be used as arenas for testing theory

in conservation biology. The approaches used and highlighted in this book can easily be used as a model for studying human impacts in other coastal communities (e.g., mangroves, seagrasses, dune systems). Rocky shores have long been championed as model systems for ecology. In the case of human impacts, we argue that salt marshes provide a better model system for ecologists, because salt marshes are just as easily manipulated and are generally more impacted than rocky shores, in both the intensity and diversity of human threats. In the process of summarizing the impacts of human activities on marshes, we have highlighted numerous indirect mechanisms by which humans commonly generate ecosystem-level change in marsh structure and function. These mechanisms are likely to be at work in many other coastal communities and should be investigated when examining potential deleterious impacts of humans. For example, we have shown that humans are (1) precipitating runaway grazer effects via overconsumption of predators (terrapins, blue crabs, alligators) and subsidizing populations of potent herbivores (i.e., geese) via agricultural fertilization (i.e., a nontraditional trophic cascade); (2) generating cross-ecosystem or agency effects by removing woody buffers on the upper edges of marshes, which then result in increased nitrogen and freshwater flow into marshes, subsequent plant invasion, loss of native plant diversity, and overgrazing by native insects; (3) altering food web structure and ecosystem function and threatening native species through a continuing series of invasions by nonnative species; and (4) causing far-reaching (across thousands of kilometers) and cryptic impacts by increasing atmospheric carbon dioxide concentrations that then lead to warming and sea-level rise, reducing the amount of sediment reaching Gulf Coast marshes by channelizing normally freeranging rivers, and subsidizing populations

of migrating consumers in agricultural fields two thousand kilometers away. 2. Widespread theory dependency of the bottom-up paradigm in salt marsh ecology along with methodological traditions of using correlation techniques instead of experiments has led to underestimating and even overlooking major human-induced threats to salt marshes and misappropriation of conservation funds. Most poignantly, this has occurred in the dismissal of the roles of top-down forces in salt marsh dieoff and recovery from disturbance. A new paradigm, which incorporates both topdown and bottom-up forces, must be integrated into both salt marsh conservation and ecology and the education of young estuarine ecologists. 3. Currently, there are many more humaninduced threats to salt marshes than just the historical one of reclamation. These threats include human-precipitated species invasions; small- and large-scale eutrophication and accompanying plant species declines; runaway grazer effects that denude marsh substrate; climate change–induced effects including sea-level rise, increased air and sea surface temperatures, and rising carbon dioxide concentrations; altered hydrologic regimes, and a wide range of pollutants, including nutrients, synthetic hormones, metals, organics, and pesticides. Importantly, the extent, occurrence, and intensity of these threats vary geographically. Thus, when managers set out to begin conservation planning for their local marsh systems, each one of these potential threats should be mapped, addressed, and assessed with a regional perspective in mind. 4. Human threats to salt marshes do not occur in isolation but, rather, overlap in space and time. Therefore, both additive and synergistic interactions among these threats can amplify degradation of marsh ecosystems, as is the case with intense interactions

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between drought and runaway snail grazing effects in southern U.S. marshes, or with eutrophication and plant invasions in several regions, or with one invasive species facilitating another (e.g., green crabs facilitating a invasive clam species fifty years after its introduction). This means that managers must take into account the potential cascading and interactive threats generated by humans when planning conservation strategies for their systems. The future decline of coastal salt marshes throughout the world could thus be much greater than currently suggested by the single-stress models (e.g., sea-level-rise-only models) we now rely heavily on. Future studies of human impacts of marshes should therefore investigate the relative effects of as many stressors as possible so that cryptic synergisms and additive impacts can be identified and specifically addressed in marsh conservation measures and reserve design. 5. Because marshes are subject to multiple impacts operating on a variety of different spatial and temporal scales, there is also the likelihood that there may be nonlinear responses to these interacting stressors. Rapid transitions in salt marsh communities may be precipitated by comparatively small changes in tidal elevation as the result of global climate change. Changes of only tens of centimeters in tidal elevation can determine the transition from open, unvegetated mudflats to densely vegetated salt marsh. Extensive efforts to remove invasive cordgrass (Spartina alterniflora and hybrids) in western salt marshes have demonstrated that increased tidal elevation resulting from sedimentation following Spartina invasion

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can limit the ability of native plant restoration efforts. In some areas (Washington), eradication of invasive Spartina may also be followed by subsequent invasion of Zostera japonica depending also on small differences in tidal elevation. In another example of a nonlinear response generated by multiple human impacts, unleashed goose grazing in Hudson Bay has led to massive loss of one of the largest marshes in the world and, subsequently, turned on intense, wind-driven evaporative stress that is exacerbated by increasing climatic extremes. Once-extensive marshes are now unvegetated salt flats. Seedling establishment in these barren areas is impossible because of desiccation and hyperosmotic soil stress. A new alternate, salt flat state has quickly formed that can only be switched back (over hundreds if not thousands of years) by the exceedingly slow pace (centimeters per year) of rhizome expansion on the edge of the few plant culms remaining. Clearly, a climate-maintained salt flat would not have replaced the massive Hudson Bay salt marsh without the interactive, initiating effect of runaway goose grazing. Combined, these examples highlight that future management of salt marsh ecosystems may require recognizing and anticipating “tipping points,” where comparatively modest anthropogenically mediated changes interact to produce large and potentially stable-like shifts in the condition and health of salt marsh ecosystems. REFERENCE Halpern, B. S., B. R. Silliman, J. Olden, J. Bruno, and M. D. Bertness. 2007. Positive interactions fundamental to effective restoration and conservation. Frontiers in Ecology and the Environment 5: 153–160.

CONTRIBUTORS

Paul Adam University of New South Wales, Sydney, Australia D. A. Ayres University of California, Davis J. P. Bakker University of Groningen, The Netherlands Mark D. Bertness Brown University, Providence, Rhode Island James E. Byers University of New Hampshire, Durham J. C. Callaway University of San Francisco Jacoby Carter United States Geological Survey, Lafayette, Louisiana Randolph M. Chambers College of William & Mary, Williamsburg, Virginia Cesar S. B. Costa Fundação Universidade do Rio Grande, Brazil Caitlin Mullan Crain University of California, Santa Cruz A. J. Davy University of East Anglia, Norwich, United Kingdom Michelle Dionne National Oceanic and Atmospheric Administration Ocean Service, Silver Spring, Maryland

Jose M. Farina Pontificia Universidad Católica de Chile, Santiago, Chile M. E. Figueroa Universidad de Sevilla, Seville, Spain Keryn Bromberg Gedan Brown University, Providence, Rhode Island Laura Gough University of Texas, Arlington Edwin D. Grosholz University of California, Davis Hugh A. L. Henry University of Western Ontario, London, Ontario Caitlin E. Hicks University of Florida, Gainseville Christine Holdredge Brown University, Providence, Rhode Island Oscar O. Iribarne Universidad Nacional de Mar del Plata, Argentina Robert L. Jefferies University of Toronto, Ontario Michael S. Kearney University of Maryland, College Park Paul A. Keddy Southeastern Louisiana University, Hammond

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Lisa A. Levin Scripps Institution of Oceanography, La Jolla, California Jordan R. Mayor University of Florida, Gainseville

Jack Siegrist Southeastern Louisiana University, Hammond Brian R. Silliman University of Florida, Gainseville

Tiffany McFalls Southeastern Louisiana University, Hammond

J. Court Stevenson University of Maryland Center for Environmental Science, Cambridge

Laura A. Meyerson University of Rhode Island, Kingston

D. R. Strong University of California, Davis

Carlos Neira Scripps Institution of Oceanography, La Jolla, California

John Teal Teal Partners, Rochester, Massachusetts

J. Andy Nyman Louisiana State University, Baton Rouge David T. Osgood Albright College, Reading, Pennsylvania Susan Peterson Teal Partners, Rochester, Massachusetts Kristin Saltonstall University of Maryland Center for Environmental Science, Cambridge

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Mads S. Thomsen University of Aarhus, Roskilde, Denmark Anna C. Tyler University of California, Davis J. B. Zedler University of Wisconsin, Madison

INDEX

Boldface numbers refer to maps, figures, and tables aboveground biomass and CO2 levels, 210 and invertebrate density, 30 at Jamaica Bay, NY, 190 with nitrogen addition and phosphorus limitation, 91 and nutria, 47 and Phragmites, 65, 66, 72, 73, 142 and plant competition, 140 and rhizome stress, 197 Rhode River, 186, 187 and Scirpus, 96 and Scirpus olneyi, 212 and Spartina, 26–27, 30, 33, 37, 196–197 accretion on Cape Cod, 192 causes, 195, 268 deficits, 186, 192, 196, 198 and grazing, 119–120 historical trends, 180 measurement of, 196 in mid-Atlantic region, 185, 186, 188, 190, 191, 192–193 at Mont-Saint-Michel, 195 and Phragmites introductions, 67 in Scirpus marshes, 179 and sea-level rise, 173–174, 178, 268 and Spartina invasions, 27, 35, 36 and tidal asymmetry, 181–185 acid precipitation, 193 Adam, P., 373, 374 adaptive management, 276–277, 297, 303 advection, 43 aerial surveys, 119, 234 Africa, Spartina species in, 8 agricultural runoff, 65, 74, 290, 347, 369, 371, 377–378, 393 Agrostis stolonifera, 317, 369, 372 Agua Hediondo Lagoon, 290 Alabama, Phragmites australis in, 63 Alameda Island, 15 Aleutian Islands, 95 algal productivity, 288 Allee effect, 13–14, 18 Alleman, L. Y., 188 alligators diet, 122–123 ecological significance, 118–119 management strategies, 130

overhunting, 124 population, 123, 125, 129 predation and nutria population, 119, 126–128 trophic cascade hypothesis, 125–130 Althaea officinalis, 63 American lobster, 233 amphibians, 245, 246 Amphibola crenata, 366 Anas castanea, 375 Anderson, L. W. J., 44 angiosperms, 45 Anisfeld, S. C., 191 Anomalopterynginea spp., 365 Anser caerulescens caerulescens, 86, 87–93, 96 Antarctica, 176, 269 Apium prostratum, 365, 369 aquaculture, 43, 343, 345–346, 380 Arctic fox, 95 Ares, A., 73 Argentina, 339, 342, 344, 346, 348–349 Armeria maritima, 318 arthropods, 162, 278 Arthur, S. C., 124 Asian mytilid mussel, 46 Aster tripolium, 317, 321 Atlantic City, New Jersey, tide records, 189 Atlantic Coast, of North America, 35, 256–259, 392 Atlantic Coast, of South America, 339–340 Atriplex patula, 369 Atriplex portulacoides, 323 Atriplex prostrata, 372 Australasian coastal salt marshes data limitations, 381 management strategies, 381–382 overview, 361–362 research considerations, 362 See also Australia; New Zealand Australia climate, 373 invasive mussels, 44 mangrove vegetation, 174 salt marsh characteristics, 373–376 salt marsh loss and threats, 376–381 salt marsh management, 376 Spartina species, 10, 35

401

Avicennia germinans, 174 Avicennia marina, 174, 292, 364, 370 Baccharis halimifolia, 161, 379 Bahia Blanca Estuary, 341, 346, 347 Balanus glandula, 28 bald cypress, 120, 217 ballast water, 43, 51, 378 Balling, S. S., 162 Ballona Wetland, 290, 301 Baltic Sea, 313, 321, 325 Bando, Jun, 15 Barbour, M. G., 286 Barleeia subtenuis, 46 barnacle, 28 Barn Island Wildlife Management Area, 192 Barnstable, Massachusetts, 159 Barnstable Marsh, 172, 179, 190 Bart, D., 65 Batillaria attramentaria, 48–50, 52 Batiquitos Lagoon, 296 bat rays, 29 bats, 375 Baumann, R. H., 175, 184 Baumea juncea, 365, 380 Bay Champagne, 174 Bay of Fundy, 153, 193 Bay of Plenty, 367 Bédard, J., 95 Belknap, D. F., 192 belowground biomass and CO2 levels, 174, 210, 211 and invertebrate density, 30 measurement of, 65 and Phragmites, 65, 66 and sea-level rise, 268 and Spartina, 26–27, 30, 34, 35, 36, 37 Benoit, G., 191 benthic invertebrates, 28–31, 34–36, 37, 277 benthic microalgal photosynthesis, 27 Bertness, M. D., 65, 172, 192, 257, 268, 271 big cordgrass, 62 biocontrol, 72–73 biodiversity loss, 321–322 bioengineers, 73–74 bioeroders, 47 biogeochemistry, 156 biomass analysis under elevated CO2 conditions, 210–214 nutria/muskrat impact, 119, 121 Phragmites australis impact, 65–66, 142 San Francisco Bay, 26 Willapa Bay, 26–27 See also aboveground biomass; belowground biomass biotic feedbacks, 165 birds Australia, 375 ditching impact, 163 Europe, 61 New Zealand, 363, 366 Phragmites use and impact, 67 San Francisco Bay, 289 South America, 339, 340–341, 344, 345 bivalves, 31, 34, 35 black rush, 140, 158 black swan, 374, 375 Blackwater Estuary, 324 Blackwater National Wildlife Refuge, 117, 182, 184 Blackwater River, 198–199 Blossey, B., 63 blue crab, 108, 111, 117, 122, 277, 392 Blutaparon verniculare, 338

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boating, 290 Bodega Bay, California, 51 Bolboschoenus medianus, 365 border protection, 296–297 Boston, K. G., 379 Boston, Massachusetts salt marsh loss and urbanization, 139, 258 tide records, 189 Botany Bay, 378, 381 bottom-up control approach, 105, 125–126, 143, 397 Boundary Bay, 14 Bourn, W. S., 159, 161, 162 Braakman, 317 brackish marshes, 68, 120, 122, 125, 186 Brancaster West marsh, 319 Branta canadensis, 93 Branta canadensis moffitti, 31 Brazil eutrophication, 345 land reclamation, 341 metal pollution, 346 shore erosion, 348 shrimp farm impact, 343, 345–346 Spartina alterniflora, 35 tropical salt marshes, 338–339 ultraviolet radiation, 349 Breaux Act (1990), 235 Breitburg, D. L., 245 Bricker-Urso, S., 192 Brigantine National Wildlife Reserve, 190 Bromberg, K., 257, 268 Bromberg, K. D., 172 Brooks, M. J., 180 brown marsh, 234 Bucher, D. J., 373–374 Buchholz, J. W., 297 Buena Vista Lagoon, 296 Bulboschoenus, 377 Burdick, D. M., 157 Burke, D. G., 197–198 burning, 198 burrowing crab, 366 Bush River, 186 Byers, J. E., 43, 48, 50, 52 caging experiments, 107–108, 369 Cahoon, D. R., 291 CALFED Bay-Delta Program, 297 Calidris alpina, 35 California Humboldt Bay, 5, 11–12, 14, 63, 299–300 Phragmites australis, 63 salt marsh area, 286 See also San Francisco Bay, California California clapper rail, 394 California least tern, 290 Callaway, J. C., 291 Callinectes sapidus, 108, 111, 117, 122 caltrans, 296 Camargue, 194 Canada coastal marsh status, 257, 260–261 hunting regulations, 96 tidal restrictions, 154 wetland inventory, 254 Canada goose, 93 Canadian Maritimes region, 257, 260–261 canals, 116, 377 Cancer crabs, 50–51 Cape Barren Goose, 375 Cape Cod, 153, 172, 179, 182, 192, 244, 257 Cape Fear, 237

Cap Tourmente, 96 carbon, organic, 28, 174, 325 carbon cycling, 214–219 carbon-14 dating, 190 carbon dioxide (CO2), atmospheric from burning practices, 198 future research, 224 increasing levels, 73, 172–174, 208, 223 and litter decomposition, 218 and microbial biomass, 215–216 progressive nitrogen limitation, 219–223 and respiration rates, 216–218 salt marsh plants’ response to elevated, 210–215 SERC experiments, 208, 209–210 and soil carbon pools, 218–219 carbon sinks, 208, 218–219, 223–224 Carcinus maenas, 29–30, 51, 270–271 Carex aquatilis, 92 Carex litorosa, 364 Carex lyngbyei, 293 caribou, 93–95 Carlisle, B. K., 257 Carlson, P. R., 109 Carlton, J. T., 44, 47 Carpenter, S. R., 104, 125 Carpinteria Marsh, 293 Casagrande, R., 63, 73 Casco Bay, 193 cattle, 368, 381 See also livestock grazing Caulerpa taxifolia, 44 causeways, 368 Cereopsis novaehollandiae, 375 Cerithidea californica, 48–50, 52 Chambers, R. M., 63, 66, 68, 244 Chasmagnathus granulata, 143–144 Chasmagnathus granulatus, 339–340, 343, 346, 347 Chen caerulescens atlantica, 95–96 Chenhall, B. E., 378 Chenier Plain, 119 Chen rossii, 93 Chernobyl disaster, 327 Chesapeake Bay cryptogenic species, 44 marsh degradation, 185–186 marsh management strategies, 197–199 nutria damage, 117 Phragmites, 62, 67 SERC, 186, 208, 209–210 chestnut teal, 375 Chile, 339, 342, 350 China flood-dominated tidal velocity, 184–185 sediment accretion in Scirpus marshes, 179 Spartina species, 4, 10–11, 18 Chloridoideae, 5 Chmura, G. L., 193 Choptank River, 185–186 Chung, C. H., 18 Church, T. M., 188 Civilian Conservation Corps (CCC), 158 clams, 29, 34, 50–51 Clean Water Act, 279 climate change Australian salt marsh impact, 379 and dieback events, 243–244 Gulf and East coast salt marsh impact, 393–394 impact of, 320–321 New Zealand salt marsh impact, 372–373 South American salt marsh impact, 348–349 See also global warming closed populations, 44

Cnemiornis spp., 365 C:N ratios, 218, 220 CO2. See carbon dioxide (CO2), atmospheric coastal marsh loss Australia, 376–377 Canada, 260–261, 392 conservation, 262 Gulf of Mexico, 259–260, 392 Hudson Bay, 261 measurement, 256 model of, 262–263 New Zealand, 367 North America, 392 North Atlantic, 256–258, 392 Pacific, 260, 392 South Atlantic, 258–259, 392 statistics, 233, 254 coastal protection, 317–319 coastal squeeze, 9, 319–320, 330, 379 Cochran, J. K., 191 Cockle Creek, 378 Cogshall marsh, 144–145 Coles, S. L., 44 Collins, J. N., 182 Collins, L. M., 182 Colquhoun, D. J., 180 Columbia River, 286, 290 common reed. See Phragmites australis Connecticut marsh degradation, 191 restoration, 273 S. alterniflora dieback, 233 sea level rise, 178, 268 tidal restrictions, 153 conservation, 262, 327, 376, 396–398 control. See management strategies Coos Bay, 14, 260, 286, 289 Coprosma spp., 364 coral reefs, 45–46, 233, 245, 246, 254 Cordell, J. R., 34 cordgrass. See Spartina species Cordylanthus maritimus, 292 Cordylanthus mollis ssp. mollis, 394 Cornwell, J. C., 185 Corophium, 35 correlation studies, 106, 109, 240, 242, 243, 392, 397 Cortaderia selloana, 379 Costa, C. S. B., 341 Cottam, C., 159, 161, 162 Cotula coronopifolia, 340, 369, 372 Cox Island, 15 Coyote Hills Slough, 15, 16 C3 plants, 73, 174, 210, 212, 213, 219, 224 C4 plants, 174, 211, 212, 213, 224 crabs blue, 108, 111, 117, 122, 277, 392 burrowing, 366 Cancer, 50–51 European green, 29–30, 51, 270–271 of New Zealand, 366 and organochlorine contaminants, 347 and soil metal concentration, 346 of South America, 143–144, 339–340, 343, 346 Crassostrea gigas, 350–351 crawfish, 122 Crissy Field, 298 Crocodylus porosus, 375 Crooks, J. A., 46 crustaceans, 44–45 crustal readjustment, 190 cryptogenic species, 44 C sinks, 208, 218–219, 223–224

index

403

culm density, 66 culverts, 154 Curlew Pond, 188 Curtis, P. S., 174 cyanobacteria, 92 Cygnus atratus, 374, 375 Cyprinodon variegatus, 158 Daddario, J. J., 190 Dahl, T. E., 254 D’Alpaos, A., 197 dams, 116, 157, 185, 291, 297, 315, 370, 377 Danube River, 61 Dark, A. K., 186 Day, J. W., 194, 198 deer, 369 deforestation, 363, 369 Delaune, R. D., 175 Delaware coastal marsh loss, 257–258 dieback events, 234, 235 ditching, 161–162 greater snow goose wintering grounds, 96 landward marsh development, 269 marsh degradation, 188 Phragmites australis, 62, 66, 71 Delaware Bay, 63, 188, 273–278, 279 Denmark, embankment and land claim, 315–316 Denno, R. F., 68 de Roos, A. M., 86 Dethier, M. N., 35 detritus, 27–28, 30–31 Dickinson Bay, 185 diebacks causes, 238–246, 272 current situation, 234–235 definition, 237 Florida panhandle, 109, 233, 234, 244 future research, 247 in Jamaica Bay, NY, 237–238 in Louisiana, 235, 236, 237 management of, 247 on mid-Atlantic coast, 179 in New England, 258 in North Carolina, 236, 237 recovery from, 234 replanting, 247 research, 233–234 vs. small-scale mortality events, 238 in South Atlantic, 259 on southeastern and Gulf coasts, 109–112 by state, 236 stressors indicating, 196–197 Web sites, 235 Dike Island, 14 diking, 11, 152, 153, 274–275, 277, 289, 318 dissolved organic carbon (DOC), 174 dissolved organic nitrogen (DON), 69 Distichlis distichophylla, 374 Distichlis spicata, 25, 140, 158, 192, 209, 234, 340 ditching, 152, 153, 154, 158–165, 271 ditch plugging, 164 diversity, 70 Dollard marshes, 322 Donnelly, J. P., 178, 192 D’Onofrio, E. E., 348 drainage, 151, 160, 164, 182, 320, 322, 380 dredging, 275–276, 290 drought, 111, 241–242, 243–244, 245–246, 272 ducks, 67 Dumbauld, B. R., 34 dumping, 371 Dundee, H. A., 118

404

index

Dunlin, 35 Dunton, K. H., 244 Dusicyon grideus, 340 ebb-dominated tidal velocity, 181–184 ecosystem engineers, 5–7, 33, 45–47, 73–74, 152 Ecuador, 343 edaphic gradients, 151 eelgrass, 26, 232 Egeria densa, 350 elephants, 104 Elkhorn Slough, 42, 49, 260, 290–291, 298 El Niño, 67 El Niño/La Niña-Southern Oscillation (ENSO), 348 Elymus athericus, 321 Elytrigia atherica, 322, 323 embankment, 315–317 embayments, 42, 43, 44 endangered species, 394 engineering projects, 317–319 Environmental Protection Agency (EPA), 42, 51–52 eradication programs, 36–37, 44, 117, 372, 379 erosion in Delaware and Maryland, 188 and ditching, 161 during Holocene age, 177 in New Zealand, 369–370 in South America, 348, 351–352 in southern California, 291 Erwin, M., 188, 190, 192 Erysimum menziesii, 12 Essex, England, 320 Eucyclogobius newberryi, 394 European Commission, Habitats Directive, 327, 328–329 European green crab, 29–30, 51, 270–271 European salt marshes characteristics, 312–314 and climate change, 320–321 and coastal protection, 317–319 conservation priorities, 327 current status, 395–396 distribution, 313 embankment and land claim, 315–317 eutrophication and pollutant loading, 324–326 grazing impact, 321–322 human impact overview, 314–315 and invasive species, 323–324 loss and expansion, 193–195 oil spill impact, 326 and radionuclide releases, 326–327 and sea-level rise, 319–320 Spartina species, 8–9, 16 eutrophication Australia, 377–378 Europe, 61, 324–326 New England, 145, 146 New Zealand, 369 and PNL, 221 South America, 345–346 in U.S., 239–240 Ewanchuk, P. J., 65 exclusion experiments, 110, 111 exotic species. See invasive species extinctions, 363 Fagherazzi, S., 182 Fairbridge, R. W., 180 false water rat, 375 Famosa Slough, 301 farming, 65, 74, 290, 347, 369, 371, 377–378, 393 Farnsworth, E. J., 70 fecal matter, 91–92 Fell, P. E., 68

Fenn, K., 42 Ferina, N. F., 175 fertilization, 140, 144–145, 219, 220 Festuca arundinacea, 372 Festuca rubra, 318 Ficopomatus engimaticus, 45–46 fiddler crab, 162, 339 field experiments, 111, 128–130, 297, 321, 349, 393 filling, 289–290 filter feeding species, 50 fire and fire management, 121, 344–345, 379–380 fish and alligator population, 123 Australia, 376 ditching impact, 162–163 New Zealand, 366 nonnative species, 45 Phragmites australis impact, 67–68, 68 San Francisco Bay, 289 Spartina impact, 32 tidal restriction impact, 157 fisheries, 157, 233, 288 Fitzgerald, D. M., 182 Flint, R. F., 180 flood-dominated tidal velocities, 184 Florida, panhandle dieback, 109, 233, 234, 244 Florida Bay, 233 Florida Everglades, 118, 122–123 Floyd, T., 51 food webs, 30–31, 68, 108, 123 forb pannes, 161 Ford, M., 273 Ford, M. A., 120 Fort Lauderdale, Florida, 258 foxes, 95, 339, 340 France embankment and land claim, 316 Spartina species, 4, 8, 16 Franz, D., 190 Fraser Delta, 96 Fraser River, 14 free-air CO2 enrichment (FACE), 212 French, J., 197 freshwater marshes, 66, 89, 92, 154, 158, 393 freshwater releases, 112, 291, 342, 377 freshwater sedge, 92 Friedrichs, C. T., 179 Fundulus heteroclitus, 68, 158, 270, 271, 277, 278 Fundulus luciae, 68 fungal infections, 240–241, 272 Furbish, D. J., 182 Fusarium verticillioides, 240 Gabet, E. J., 182 Gabrey, S. W., 122, 123 Gaiero, D. M., 347 Gallagher, J. L., 71, 286 Ganong, W. F., 153 Gao, S., 184 garbage dumping, 371 garden experiments, 66 Gardner, T. A., 254 Garofalo, D., 188 gastropods, 366 Gauthier, G., 96 geese Arctic salt marsh degradation, 86, 87–93, 94, 106, 117–118 in Australia, 375 in Europe, 322 in New Zealand, 363, 365 Spartina grazing, 31 temperate salt marsh impact, 95–96

Gehrels, W. R., 180 Gemma gemma, 51 genetic surveys, 15 Georges River, 380 Georgia dieback events, 110–111, 234, 236, 239, 240, 241–242 salt marsh loss, 268 Georgia Coastal Research Council, 235 Germany embankment and land claim, 315–316 Spartina species, 8 Geukensia demissa, 29 Giblin, A. E., 156, 158 Giroux, J. -F., 95 glaciers, 173, 175–176 glasswort, 140 Glaux maritima, 318 global warming Australia, 379 ice rafting impact, 193 impacts of, 172–176 New Zealand, 372–373 Phragmites impact, 73 research recommendations, 224 See also carbon dioxide (CO2), atmospheric; sea-level rise golden mussel, 351 gold mining, 260 Goldstein, G., 73 Goldwasser, L., 48, 52 Goss-Custard, J. D., 35 Gosselink, J. G., 175 Gossypium hirsutum, 213 Gough, L., 120 government agencies, 198 Grace, J. B., 120 grapsid crab, 143–144 Gratton, C., 68 gray fox, 340, 343 Gray’s Harbor, Washington, 12, 14, 63, 290 grazers and grazing Arctic salt marshes, 87–95 East and Gulf coast marshes, 109–112, 392 management actions and recommendations, 96–97, 112 and plant sensitivity to flooding or salinity, 119 research considerations, 86, 104 temperate salt marshes, 95–96 trophic cascade hypothesis, 104–109 See also geese; livestock grazing; snails Great Britain. See United Kingdom greater snow goose, 95–96 Greater Thames Region, 195 Great Lakes region, Phragmites australis, 63 Green, M. O., 370 greenhouse gases. See carbon dioxide (CO2), atmospheric greenhouse studies, 66, 158, 174 Greenland, glacial melt, 173, 175–176, 195, 269 Grevstad, F. S., 299 gross national product (GNP), coastline economies as percentage of, 232 groundsel bush, 379 grubbing, 88, 89, 90, 96, 106 Guarnieri, R. A., 349 Gukensia demissa, 68 Gulf coast, 35, 63–64, 392–394 Gulf of Maine, 158, 177–178 Gulf of Mexico, 257, 259–260 Gunderson, L. H., 123 Guyana, 338 Haacks, M., 364, 369 habitat loss, 32, 254, 261–262, 293–297, 301, 394 Habitats Directive, 327, 328–329 Hacker, S. D., 35, 299

index

405

Hairston, N. G., 104 Halberd-leaved tearthumb, 63 hares, 322 Harris, S., 380 Harrison, R. W., 259 Hartig, E. K., 190 Hartman, J. M., 65 Haslett, S. K., 195 Hawaii, invasive plant species, 73 heavy metals, 324, 346, 370–371, 378–379 Hedge, P., 35, 379 Heijns, H., 174 Helice crassa, 366 Hemigrapsis sanguineus, 271 Hemigrapsus oregonensis, 51 Hensel, P. F., 194 herbicides, 8, 37, 72, 299, 325–326, 379 herbivores, 86, 94, 104, 117, 119, 197 Hilliard, R. W., 44 historical evidence and data alligator, muskrat, and nutria population, 123–125, 129 baseline establishment, 261–264 dieback events, 235–238 Phragmites distribution, 59 salt marsh use, 254–255, 267 sea-level rise, 348 South American salt marshes, 337 wetlands, 185 Holland Glade marsh, 188 Holocene, 6, 8, 175, 176–179, 180 Homarus americanus, 233 homeland security, 296–297 Hoozemans, E., 173 Hordeum marinum, 369 Howarth, R., 381 Howes, B. L., 268 Hudson Bay, 256 coastal marsh loss, 257, 261 coastal vegetation surveys, 97 lesser geese breeding colonies, 89 salt marsh characteristics, 90 salt marsh degradation, 86, 87–93, 94, 106, 117–118, 243 Humboldt Bay, California, 5, 11–12, 14, 63, 299–300 Hung, G. A., 193 Hunter, K. L., 68 Hunter-Cario, L., 15 Hunter Island, 191 hunting regulations, 96–97 Hurricane Katrina, 116, 130 hurricanes, 9, 123, 130, 184, 192, 260, 268, 348 Hutchings, P. A., 44 hybridization Phragmites australis, 72 Spartina alterniflora, 4, 11, 15, 16, 18, 25, 33, 293 hydrogen sulfides, 174, 196, 237, 241 hydrology, 149–152, 175, 268, 273, 286–287, 290–292 Hydromys chrysogaster, 375 hypersaline soils, 90, 261, 287, 296, 338, 340, 348, 373, 374

in Europe, 323–324 examples of, 44–51 in Gulf and East coast salt marshes, 393 impacts of, 37–38, 58, 269–271 in New Zealand, 363, 371–372 of Pacific Coast, 29–30, 292–293, 299–300, 395 policy recommendations, 51–53 in San Francisco Bay, 42 in South America, 350–351 See also Phragmites australis; Spartina alterniflora invertebrates in Australia, 375–376 and biomass, 30 ditching impact, 162 in Elkhorn Slough, Calif., 42 and introduced Phragmites, 68 and invasive Spartina, 28–31, 34–36, 37 of Pacific Coast, 288 in restored marshes, 277 in South America, 338, 339–340, 341 Ireland, Spartina species, 6, 8 Irish Sea, 327 isopods, 47 isostatic uplift, 317–318, 319 Iva frutescens, 140, 161

Iceland, 180, 313 ice rafting, 193 IJsselmeer, 317 Ilyanassa obsoleta, 29 impact assessment, 52 insects, 72–73, 104, 144, 289 Intergovernmental Panel on Climate Change, 269 international cooperation, 199, 298 intertidal marshes, 89, 93, 146, 152, 337 intertidal mud flats, 9, 18, 36, 320, 330 invasive species in Australia, 379 control of, 299–300, 302 in estuaries, 42–44

lab experiments, 111, 242 Lachnogrostis littoralis, 364 lagoons, 297–298, 373. See also specific lagoons Lake, R. W., 162 lake ecosystems, 125 Lake Illawarra, 378 Lake Macquarie, 378 Lama guanicoe, 343 Lambert, A. M., 63 land claims. See reclamation landfill, 341–342 LANDSAT imagery, 86 landward marshes, 269 Lanfredi, N. W., 348

406

index

Jacksonville, Florida, 258 Jaensch, R., 380 Jagaribe River, 342 Jamaica Bay, New York, 190–191, 237–238 James Bay, 86 Japanese littleneck clam, 50–51 Jaumea carnosa, 25 Jefferies, R., 106, 109 Jobbins, D. M., 160 Josselyn, M. N., 297 Juan Carlos I Dike, 318 Juncus acutus, 345, 372, 379 Juncus gerardi, 140, 158, 318, 372 Juncus kraussii, 364, 365, 366–367, 369, 379 Juncus maritimus, 364 Juncus roemerianus, 234 Kearney, M. S., 180, 181, 186, 188 Kelleway, J., 380 Kelley, J. T., 192 kelp forests, 233, 254 Kennish, M. J., 41 Khim, H. S., 46 Kim, G., 188 Kings Creek marsh, 185 Kirkpatrick, J. B., 380 Kirwan, M. L., 178, 197 Kitchell, J. F., 125 Koch, E., 197–198 Kollmorgen, W. M., 259 Konisky, R. A., 157 Kriwoken, L. K., 35, 379

La Pérouse Bay, Manitoba, 86 La Plata River, 343, 352 Lauwerszee Estuary, 317 Layia carnosa, 12 lead contamination, 347 League, M. T., 71 Leatherman, S. P., 175 Lee, R. W., 35 Leersia oryzoides, 63 Leonard, L. A., 67 Leopold, L. B., 182 Leptinella dioica, 364, 365 Leptocarpus similis, 364, 365, 366–367 Leptochelia dubia, 46 Leptospermum scoparium, 364 Lesser, C. R., 162 lesser snow goose, 86, 87–93, 96–97, 106, 261 levees, 116 light-footed clapper rail, 290, 294, 300, 394 light penetration, 27, 69–70 lignin content, 218 Limnoperna fortunei, 351 Limonium companyonis, 372 Linthurst, R. A., 237 litter accumulation, 67, 321, 322 litter decomposition, 66, 92, 216, 218 Little Beach, New Jersey, 190 Little Ice Age, 180 Littoraria irrorata (marsh periwinkle), 106–107, 110, 143, 186–187, 270–271 Littoraria littorea, 270–271 Littorina saxatilis, 42 livestock grazing in Australia, 381 in Europe, 321–322, 330 impacts of, 86, 90 in New Zealand, 368–370 in South America, 343–344 living shoreline projects, 197–198 Llewellyn, D. W., 120 lobsters, 233, 245 Loftus, W. F., 123 logging, 291 Long, S. P., 321 Long Island, marsh degradation, 191 Long Island Sound, 233, 245 Long Term Ecosystem Research (LTER), 186 Los Angeles Harbor, 296 Los Peñasquitos Lagoon, 292, 298 loss, of coastal marshes. See coastal marsh loss Lotze, H. K., 254 Louisiana coastal wetland loss, 116, 260 dieback events, 109, 110, 233, 235, 236, 237, 241, 242 nutria impact, 116–117, 119, 121–122 sediment loss via shipping channels, 9 See also alligators Lowe, A., 366 Lucke, J. B., 176, 181 Luque, C. J., 324 Lynch, J. C., 291 MacDonald, I. T., 370 Macdonald, K. B., 286 Machiasport, Maine, 180 Macoma petalum, 29 Maine marsh degradation, 192–193 marsh formation, 180 mosquito ditching, 159 open marsh water management, 164 Phragmites, 393 restoration, 273

sea level rise, 268–269 tidal restrictions, 154, 156 management strategies adaptive, 276–277, 297, 303 alligators, 130 Australasian salt marshes, 379, 381–382 Chesapeake Bay marshes, 197–199 conservation efforts, 262, 327, 376, 396–398 dieback, 247 in Europe, 319–320, 330 lesser snow goose, 96–97 living shoreline projects, 197–198 monitoring programs, 52, 70–71, 276–277, 301–302 mosquitos, 138, 158–165, 271–272, 380 New England salt marshes, 146 New Zealand salt marshes, 366–367, 372 nutria, 117, 197 Pacific Coast, 293–303 Phragmites australis, 70–73, 146, 276, 278 recommendations, 396–398 sediment subsidies, 198–199 South America, 351–352 Spartina alterniflora, 36–37, 299–300, 372 See also restoration mangroves, 174, 338, 342–343, 348, 364, 373, 379 Manhattan, coastal marsh loss, 257 manipulative experiments, 104, 106, 365 Maquoit Bay, 193 Marchant, C. J., 4 Marchant, M., 173 Mar Chiquita Lagoon, Argentina, 45–46, 342, 344, 345, 347 Marcoma petalum, 28 Maricle, B. R., 35 marinas, 290 marine protected areas, 396 mark-recapture studies, 129 Marsh, A. S., 174 Marshfield, Massachusetts, 159 marsh formation models, 178, 179–181, 197 marsh hay. See Spartina patens marsh mallow, 63 marsh periwinkle, 106–107, 110, 143, 186–187, 270–271 marsh pools, 271 Maryland coastal marsh loss, 257–258 marsh degradation, 185–186, 188 nutria eradication, 117 Phragmites australis, 62, 63, 71 salt marsh loss, 268 Mason, C. F., 325–326 Massachusetts coastal marsh loss, 255 dieback events, 235, 240 open marsh water management, 164 Phragmites australis, 63 restoration, 273 salt marsh loss, 268 sea level rise, 268–269 tidal restrictions, 153, 154 Massachusetts Bay, 257 Mattaponi River, 186 Mauna Loa, CO2 changes, 173 McCready’s Creek, 182 McIllhenny, E. A., 123, 124 McKee, K. L., 109, 241, 242 Mclary, M., 68 Meadows, R. E., 63, 66, 70 Mediterranean coast, 314 Megonigal, J. P., 186 Mendelssohn, I. A., 242 Menzies’ wallflower, 12 mesohaline marshes, 61, 63, 67, 185, 273 metal pollution, 242, 324, 346–347, 370–371, 378–379

index

407

methane, 174, 198, 217–218 metonic lunar tidal cycle, 244 Mexico, wetland status, 254 Meyerson, L. A., 64, 69, 70 Miami, Florida, 258 Micalo Island, 380 microalgal photosynthesis, 27 microbial biomass, 215–216 microbial respiration, 216 microorganisms, 45 microsatellite analysis, 60, 64 microtidal marshes, 182, 196 migratory wading birds, 375 Mimulus repens, 365, 369 Minchinton, T. E., 65, 67 mining, 314, 324, 346 Mission Bay, 46, 290 Mississippi Deltaic Plain, 119 Mississippi River, 9, 116 Mississippi River Delta, 6, 61, 175, 182, 259 moa, 363, 365 Mockhorn Island, 188 mollusks, 44–45 Monie Bay, 180, 181, 185–186 monitoring programs, 52, 70–71, 276–277, 301–302 Mont-Saint-Michel, 194–195, 316, 322, 325 Morris, J. T., 178 Morton, R. A., 175 Moser, M. E., 35 mosquito control, 138, 158–165, 271–272, 380 moths, 94 Mozder, T. J., 69 mudflats, 25–32, 35, 47, 109, 195, 238, 293, 295, 323, 338, 370, 372, 379 mud snail, 48–50 Mugu Lagoon, 291 multiple stressors, 244–246, 397–398 mummichogs, 68, 158, 270, 271, 277, 278 Murphy, F. J., 162 Murray, A. B., 178, 197 Murray-Darling River, 377 Musculista senhousia, 46 muskrat, 67, 116, 119–122, 123, 124, 129 mussels, 44, 46 Muzzi Marsh, 295–296 Mya arenaria, 29, 51 Myers, R. S., 120 Myliobatus californica, 29 Myocastor coypus. See nutria Myoporum laetum, 293 Mytilopsis sallei, 44 Nags Creek Marsh, 192 Nairn, H. J., 366 Nanticoke River, 185 Narrangansett Bay, 70, 141, 145, 191–192 narrow-leaved cattail, 63 National Estuarine Research Reserve (NERR), 294, 298 National Oceanic and Atmospheric Administration (NOAA), 286, 296 National Science Foundation (NSF), 294, 295 National Wetland Inventory (NWI), 62 natural experiments, 276–277 Natura 2000 network, 327, 330 Nature Conservancy, 396 Nauset Inlet, 182 Nauset Marsh, 192 Neckles, H. A., 158 Neiring, W. A., 192 nekton, 68, 157 Neophema chrysogaster, 375 NERR (National Estuarine Research Reserve), 294, 298 net ecosystem exchange (NEE), 214, 224

408

index

Netherlands coastal protection, 317 embankment and land claim, 315–316 salt marsh area, 316 Spartina species, 6, 8, 9 New Alameda Creek, 15, 16 Newell, S. Y., 241 New England coastal marsh loss, 257 current status of salt marshes, 395–396 mosquito ditching, 158–165 S. alterniflora dieback, 236 salt marsh threats, 392 shoreline development impact, 137–148 tidal restriction, 152–157 See also specific states New England Estuarine Research Society, 235 New Hampshire open marsh water management, 164 restoration, 273 tidal restrictions, 153, 154 New Jersey coastal marsh loss, 257 marsh degradation, 188, 190 marsh pools, 271 Phragmites australis, 62 New York coastal marsh loss, 257 marsh degradation, 190–191 New Zealand human colonization, 362–364 population, 363 salt marsh characteristics, 364–366, 374 salt marsh loss and threats, 367–373 salt marsh management, 366–367 Spartina species, 6, 9–10 Nicholls, R. J., 173, 174 Nile Delta, 185 nitrogen, 69, 91–92, 141, 144–145, 208 See also eutrophication NOAA (National Oceanic and Atmospheric Administration), 286, 296 nonnative species. See invasive species Noonburg, E. G., 50 North Atlantic coast, 256–258 North Carolina coastal marsh loss, 258 Phragmites australis, 62 S. alterniflora dieback, 236, 237 North Sea, 319, 320 Nova Scotia, 193, 260 NSF (National Science Foundation), 294, 295 nuclear reprocessing facilities, 326–327 Nummedal, D., 182 nutria, 49, 118 alligator predation, 122–123, 126–130 impacts of, 47, 119–122 in Louisiana, 116–117 management of, 116–117, 197 population, 124, 126–128, 129 Nutricola confusa, 51 Nutricola tantilla, 51 nutrient addition experiments, 145, 325 nutrient cycling, 68–69, 86, 324–325 Nuttallia obscurata, 50 NWI (National Wetland Inventory), 62 Nydick, K. R., 191 O’Connell, K. A., 34 Odiel marshes, 318, 324 Odum, E., 105 Oertel, G. F., 186 off-road vehicles, 380

oil spills, 326, 347, 378 oligohaline marshes, 61, 66, 118, 121, 185, 237 Ondatra zibethicus (muskrat), 67, 116, 119–122, 123, 124, 129 O’Neil, T., 124 Oosterschelde, 317 open marsh water management (OMWM), 164, 165 Ophicardelus costellaris, 366 opossum, 339 orange-bellied parrot, 375 Orcas Island, 14 Oregon salt marsh area, 286 wetland loss, 289 organochlorine contaminants (OC), 347 Orontium aquaticum, 217 Orson, R. A., 175, 176, 192 orthophotos, 159 Otago region, of New Zealand, 365–366 overgrazing. See grazers and grazing oxidation, of rhizosphere, 68, 74, 152 oysters, 9, 12, 45, 292–293 ozone, 349, 373 Pacific Coast, of North America, 257 characteristics, 286–287 coastal marsh loss, 260 human impacts, 289–293 management efforts, 293–303 salt marsh threats, 394–395 Pacific Coast, of South America, 340–341 Pacific Estuarine Research Laboratory (PERL), 294, 298 Pacific oyster, 350–351 Packett, R., 63, 66 Padilla Bay, 14 pampas grass, 379 Panageto, W., 175 Pandolfi, J. M., 254 Parapholis incurva, 292 Partridge, T. R., 365 Paspalum distichum, 343 Paspalum vaginatum, 343, 372 pathogens, 240 Patos Lagoon, 341, 342, 346, 348, 351 Patterson, C. S., 174 Pattison, R. R., 73 Paviour-Smith, K., 366 Payne, R., 63 Pb dating, 185, 188, 190, 191, 192, 193, 196 PCBs, 324, 347, 371 Pearl River Delta, 11 Pearse, J. S., 42 peat, 160, 161, 181, 268 peat core profiles, 59, 156 Peel-Harvey Estuary, 378 Peltier, W. R., 190 periwinkle snails, 106–107, 110, 143, 186–187, 270–271 PERL (Pacific Estuarine Research Laboratory), 294, 298 Perry, J. E., 179 Persson, L., 86 Peru, 346 pesticides, 347, 370–371, 380 Pestrong, R., 182 Philipp, K. R., 188 phosphates, 324 photosynthesis, 208, 210, 212, 213, 321, 326 Phragmites species, in Australia, 377 Phragmites australis CO2 elevation impact, 213–214 distribution, 59–64 in Europe, 323 genetic diversity, 59–61

growth, 58, 65, 142 hybridization, 72 impacts of, 28, 67–70, 269–270 invasiveness of, 64–66 and metonic lunar tidal cycle, 244 native vs. nonnative strains, 66, 70 in New England, 140–143, 146, 157, 393 vs. other wetland species, 66–67 overview, 57 progressive nitrogen limitation, 221 protection of native strains, 71–72 reproduction, 64 research needs, 74–75 restoration and management programs and strategies, 70–73, 146, 276, 278 and tidal restriction, 157 uses, 61 phytoplankton, 324 phytotoxins, 241 Plagianthus divaricatus, 364, 374 Plantago australis, 372 Plantago coronopus, 372 plant mortality, 238 PNL (progressive nitrogen limitation), 208, 219–223, 224 Poa cita, 364 polderland, 315–316 policy issues, 51–53, 267–268 political issues, 296–297 pollen dating, 193 pollution Australia, 378–379 Europe, 324–326 management efforts, 298 New Zealand, 370–371 South America, 345–347 U.S., 41–42, 239–241, 298 polychlorinated biphenyl (PCBs), 324, 347, 371 Polygonum arifolium, 63 Polygonum punctatum, 121 Polypogon monspeliensis, 292 Pont, D., 194 Poplar Island, 198 POPs (persistent organic pollutants), 370, 371 population statistics coastal, 232 New Zealand, 363 South America, 351 world, 172 Portnoy, J. W., 156, 158 ports, 43 Potamopyrgus estuarinus, 366 Pousa, J. L., 348 Powell, A. N., 291 precipitation, 286, 348 predators, 104–105, 108, 112, 117 See also alligators preservation, 235, 279, 301 Pringle, J. M., 43 progressive nitrogen limitation (PNL), 208, 219–223, 224 Prokelisia marginata, 299 protection, 71–72, 279, 317–319, 396 Protothaca staminea, 50 Prudence Island, Rhode Island, 144–145 public education, 352 Public Service Enterprise Group (PSEG), 273–278, 279 Puccinella stricta, 365 Puccinella walkeri, 365 Puccinellia spp., 369 Puccinellia distans, 372 Puccinellia fasciculata, 372 Puccinellia maritima, 317, 321, 322 Puccinellia phyrganodes, 92, 93 Puccinellia walkeri, 364

index

409

Puget Sound, Washington Asian mytilid mussel, 46 Phragmites australis, 63 Spartina alterniflora, 14, 292 Spartina anglica eradication, 36, 299 Spartina hybridization, 4 wetland area, 286 wetland loss, 289 Puget Sound Nearshore Ecosystem Restoration Project, 297 Pye, K., 195 quality ditching, 164 quasi-experiments, 165 rabbits, 339, 372 radiocarbon dating, 190 radioisotopic studies, 193, 196 radionuclide releases, 326–327 Rahmstorf, S., 178 railroads, 152, 260, 293 Rallus longirostris levipes, 290, 294, 295, 300, 394 Rallus longirostris obsoletus, 394 Rampino, M. E., 177 Ramsar Convention, 172, 254, 261, 376 Rangifer tarandus, 93–95 Ranwell, D. S., 16, 321 Rappahannock River, 63, 71 Rash, J. A. E., 379 ravinement surfaces, 177 Ray, G. L., 44 reclamation in Australia, 376–377 in Europe, 315–317, 330 Gulf Coast region, 259 in New England, 269, 395 in New Zealand, 367–368 Pacific Coast region, 260 San Francisco Bay, 11 in South America, 341–342 See also shoreline development impact Reclamation, California, 11 recreation, 290, 301 red-capped wren-spinetail, 345 Redfield, Albert, 171, 176, 179, 180, 190 Reed, D., 273 Reeder, T. G., 35 reefs, 45–46, 233, 245, 246, 254 refuges, wildlife, 88, 164 regional management efforts, 297 Rehoboth Bay, 188 reindeer, 94, 95 Reithrodontomys raviventris, 394 relative sea-level rise (RSLR), 185–186, 188, 190, 191, 192, 196, 197 remote sensing, 175, 235, 273, 381 Resh, V. H., 162 resource competition, 50 respiration rates, 26, 28, 92, 216–218 restoration definition, 272 Delaware Bay, 273–278, 279 in Europe, 319–320, 330 large project planning, 273–276 in New Zealand, 366–367 Pacific Coast, 25, 293–297, 300–303 of Phragmites-invaded marshes, 70–71 in South America, 351 Spartina alterniflora introduced for, 25 tidal, 157–158, 163, 164–165, 273 of vegetation, 272–273 retreat, managed, 320, 330 Rhizedra lutosa, 73 Rhizoctionia solani, 240

410

index

Rhode Island marsh degradation, 191–192 Phragmites australis, 63, 72 salt marsh loss, 268 Rhode River, 186, 187 Rhône Delta, 194 Rice, N. A., 51 rice cutgrass, 63 Riggs, S. R., 177 Rilling, G. C., 68 Rio Grand port, 341 river bulrush, 63 roads, 368, 381 Robertson, T. L., 68 Robinson, T. B., 44 Rockaway Beach, New York, 191 rodents, 339 Rogers, K., 174 Rolinski, S., 194 Roman, C. T., 192 Roman Empire, 314 root biomass, 211, 213, 219, 223 Rooth, J. E., 67, 185 root respiration, 216 Rossman, D. A., 118 Ross’s goose, 93, 261 RSLR (relative sea-level rise), 185–186, 188, 190, 191, 192, 196, 197 Rubin, M., 190 Ruiz, G. M., 44–45 Rumbstick Cove, 192 runneling, 380 Russia, overgrazing of reindeer, 95 Sacramento-San Joaquin delta, 63 Saenger, P., 373–374, 374 Sagittaria lancifolia, 121 Saintlan, N., 174 Salicornia ambigua, 349 Salicornia europaea, 140 Salicornia virginica, 12, 15, 25, 29, 287, 291 saline marshes, 122, 199, 287 salmon, 288, 297 Salmon River, 295 salt crusts, 296 salt flats, 338, 339, 342–343, 374, 398 salt marsh grass, 92 salt marsh harvest mouse, 394 Saltonstall, K., 59, 63, 64, 66 salt or spike grass (Distichlis spicata), 25, 140, 158, 192, 209, 234, 340 salt pans, 314, 316–317 salt production, 380 saltwater crocodile, 375 Samolus repens, 364, 365, 369 Sanders, J., 177 San Diego, California Famosa Slough, 301 tidal inlet restrictions, 290 San Diego Bay caltrans, 296 elevational distribution of plants, 288 Sphaeroma quoyanum, 47, 48 wetland loss, 260, 290 Sandy Hook, New Jersey, 188, 189 San Francisco Bay, California Asian mytilid mussel, 46 cryptogenic species, 44 ebb domination, 182 invasive species, 42 Phragmites australis, 63 rare species, 289 salt marsh characteristics, 11 sediment supply, 6

Spartina alterniflora, 25–32, 36, 37, 292, 293, 300 Spartina anglica, 14 Spartina densiflora, 12 Spartina foliosa, 15, 287 Spartina hybridization, 4, 16–18, 24 Spartina patens, 15 Sphaeroma quoyanum, 47 wetland area, 286 wetland loss, 260, 289 San Martín, C., 350 Sapelo Island, Georgia, 105, 106, 239, 241 Sarcocornetea fruticosa, 318, 329, 340 Sarcocornia fruticosa, 340 Sarcocornia perennis, 318, 339, 349 Sarcocornia quinqueflora, 364, 369, 377 satellite altimetry and imagery, 173, 175, 196, 198, 273 Savannah, Georgia, 258 Schoenoplectus, 377 Schoenoplectus americanus, 120, 121, 125, 174 Schoenoplectus maritimus, 317 Schoenoplectus pungens, 365, 367 Schoenoplectus tabernaemontani, 317 Schröeder, A., 86 Schwimmer, R. A., 188 sciaenids, 277 Scirpus species, 292 Scirpus californicus, 340 Scirpus fluviatilis, 63 Scirpus olneyi, 209, 210, 211, 212–213, 217, 218, 219, 220, 221, 222, 223 Scirpus pungens, 95–96 Sclerostegia arbuscula, 380 seabirds, 94, 363 Sea Grant, 51–52 seagrass, 364 sea-level fluctuations, 244 sea-level rise in Australia, 379 in Europe, 319–320 Gulf and East coast salt marsh impact, 393–394 historical evidence and data, 348 during Holocene, 176–178 and landward marsh development, 269 and marsh formation models, 178, 179–181, 197 marsh survival during, 195–199 in New Zealand, 373 and Phragmites, 73–74 rate, 268 in South America, 348 wetland loss due to, 173–174, 175 sedges, 209 sediment accretion. See accretion sedimentation rates, 33, 191, 195, 297, 299, 369–370 sediment deficits, 6, 9, 238, 258 sediment harvesting, 5–6 sediment subsidies, 198–199, 291 Seeliger, U., 341 selection regime modification (SRM), 43 Seliskar, D. M., 71, 286 Sellafield, 326–327 Selliera radicans, 365, 369 Seneca, E. D., 237 Sepetiba Bay, 346 SERC (Smithsonian Environmental Research Center), 186, 208, 209–210 setback, managed, 320 SETs (surface elevation tables), 175, 186, 188, 190, 196 Severn Estuary, 314 sewage, 345, 369 Shackleton, N. J., 176 Shaffer, G. P., 120 Shasta ground sloth, 59 sheep, 321, 343, 369, 381

sheepshead minnow, 158 ships and shipping, 9, 378 Shisler, J. K., 160 shoot pulling, 88–89 shorebirds, 32, 35, 49, 50, 302, 341 shoreline development impact in Australia, 376–377 in New England, 137–148, 269 in New Zealand, 368 North Atlantic, 256–257 See also reclamation shrimp farming, 343, 345–346 Silliman, B. R., 65, 106, 109, 110, 186, 241 Simenstad, C., 273 Siuslaw River, 14, 15 Skagit River, 290 skin cancer, 373 Slobodkin, L. S., 104 Smith, F. E., 104 Smith, W. G., 237 Smithsonian Environmental Research Center (SERC), 186, 208, 209–210 smooth cordgrass, 63 snails, 42, 44, 48–50, 52, 106–111, 186–187, 241, 392 snow geese, 243 soil acidification, 156, 241–242 soil anaerobosis, 241 soil carbon pools, 218–219 soil degradation, 86, 88–97 soil desiccation, 241–242 soil elevation, 47, 120, 156, 295–296 soil respiration rates, 92, 215, 216–218 soil salinity degraded salt marshes, 90 Pacific Northwest, 287 and Phragmites, 65, 66 and plant growth rate, 212 in southern California, 287 and tidal hydrology, 151, 273 South Africa exotic and cryptogenic species, 44 Spartina species, 8 South American salt marshes characteristics of, 337–341 current status, 396 eutrophication, 345–346 and fire, 344–345 freshwater diversion impact, 342 history, 337 invasive species impact, 350–351 landfill threat, 341–342 and livestock grazing, 343–344 and metal pollution, 346–347 and oil spills, 347 and organochlorine contaminants, 347 salt production impact, 342–343 Spartina alterniflora, 35 and ultraviolet radiation, 349 South Atlantic coast, 257, 258–259 South Bay Salt Ponds, 6, 297 South Carolina dieback events, 110–111, 234 marsh erosion, 175 Phragmites australis, 62 salt marsh loss, 268 sea level rise, 178 Spartina alterniflora, 6 Southern California Wetlands Recovery Project, 297 Spain Juan Carlos I Dike, 318 mining, 314, 324 Prestige disaster, 326 Spartina species, 8, 323

index

411

Spartina species in Australia and Tasmania, 10 in China, 10–11 cultivation, 3–4, 7–8 as ecosystem engineers, 5–7 in Europe, 8–9, 323 in New Zealand, 9–10 and nitrogen systems, 69 Pacific coast introductions, 11–15 in San Francisco Bay, California, 11 and sedimentation, 370 in South America, 340 species and hybrids, 4–5, 15–18 Spartina alterniflora and Avicennia germinans, 174 and benthic invertebrates, 28–29, 34–36, 37 benthic microalgal photosynthesis impact, 27 in China, 10–11, 184 CO2 elevation impact, 213 control efforts, 299–300 crab impact, 340 cultivation, 4, 5, 6, 7 Delaware Bay restoration program, 277 ditching impact, 158–159, 161, 271 eradication programs, 36–37, 398 in Europe, 8, 323 eutrophication impact, 345 fertilization experiment, 144–145 fish impact, 68 and growth of other invasive species, 29–30 Gukensia demissa impact, 68 hybridization, 4, 11, 15, 16, 18, 25, 33, 293 in New England, 140, 141 in New Zealand, 10, 372 organic matter and nutrient impact, 27–28 Pacific Coast, 11–14, 24–35, 292–293 and Phragmites australis, 63, 65 physical processes impact, 27 prediction of impacts, 32–34 range, 5 in San Francisco Bay, 11 snail impact, 106, 108, 110–111, 271 in South America, 338, 339, 345 vascular plant production impact, 25–27 and vertebrates, 31–32 See also diebacks Spartina alterniflora x Spartina foliosa, 4, 11, 16, 18, 25, 293, 300 Spartina anglica in Australia and Tasmania, 10, 379 eradication, 36–37 in Europe, 5, 8–9, 16, 323 hybridization, 4 introduced plants, 4 and invertebrate numbers, 35 in New Zealand, 6, 372 origins, 5 Pacific Coast, 14, 292 Puget Sound, 299 San Francisco Bay, 17 sheep grazing impact, 321 UV-B irradiation, 321 Spartina cynosuroides, 62 Spartina densiflora crab impact, 340 in Europe, 8, 323 eutrophication impact, 345 fire impact, 345 hybridization, 5, 15 introduced plants, 4 Pacific Coast, 11–12, 15, 292, 293 range, 5 San Francisco Bay, 17 in South America, 339, 345

412

index

Spartina foliosa extinction risk, 15 and hybrid invasion, 16–18, 25 and invertebrates, 31, 34 Pacific Coast, 5, 11, 12, 15, 287, 291 range, 5, 16 Tijuana Estuary restoration, 294–295 Spartina longispica, 5 Spartina x longispicula, 339 Spartina maritima, 4, 5, 8, 318, 323 Spartina neyrautii, 16 Spartina patens dieback events, 234, 241 ditching impact, 158, 161, 271 under elevated CO2 conditions, 211 in Europe, 8, 323 introduced plants, 4, 15 litter decomposition, 218 in New England, 138, 140 nutria protection impact, 120 Pacific Coast, 15, 292 progressive nitrogen limitation, 222 range, 5 San Francisco Bay, 17 at SERC marsh, 209 and soil salinity, 212 Spartina townsendii, 237 Spartina x townsendii, 9, 15–16, 323 Spartonoica maluroides, 345 Sphaeroma quoyanum, 47, 48 spikegrass, 140 Sporobolus virginicus, 374, 380 Spruce Creek, 155, 156 squeeze, coastal, 9, 319–320, 330, 379 St. Lawrence River, 95, 260 stable isotope enrichment experiments, 30–31 Staten Island, coastal marsh loss, 257, 259 Sterna antillarum browni, 290 Stevenson, J. C., 66, 67, 175, 181, 182, 185, 197–198 stoichiometry, 220 storm water discharge, 377 Stuiver, M., 190 Stumpf, R. P., 188 Suaeda australis, 377 Suaeda novae-zelandiae, 365, 369 subsidence, 9, 156–157, 165, 175, 196, 330 Suisun Marsh, 63 sulfides, 65, 68–69, 158, 174, 196, 197, 237, 241, 346 Superfund sites, 42 surface elevation tables (SETs), 175, 186, 188, 190, 196 Suriname, 338 suspension feeding species, 50 Swales, A., 370 Swamy, V. P., 273 Swanson, M. L., 294 Sweeney, C., 51 Sweetwater Marsh, 293 Sylvitski, J. P. M. synergism, 244–246, 397–398 Tasmania, 6, 10, 373, 374, 379, 380. See also Australasian coastal salt marshes Taxodium distichum, 120, 217 Taylor, S. J., 379 Teal, J., 105, 268, 274 Teal, J. M., 144 Terebrasabella heterouncinata, 44 Terrebonne Parish, 129 Tessier, M., 322 Tewksbury, L., 73 Texas, dieback events, 234 Thalassia testudinum, 232–233 Thannheiser, D., 364, 369

Thomsen, D., 366–367 tidal elevation, 6, 27, 32, 35, 36, 151, 398 tidal flushing, 151, 158, 178, 286, 294, 347, 380 tidal hydrology, 149–152, 273, 275, 286–287, 290–292, 297–299 tidal inlets, 290–291, 297–298, 317 tidal restoration, 157–158, 163, 164–165, 273 tidal restrictions, 150, 152–157, 272, 290 tidal velocities, 181–184 tide gates, 152, 154, 157, 273 tidewater goby, 394 Tijuana Estuary, 291, 294–295, 296, 298–299, 301 Tiling, G., 175 Tiner, R. W., 257–258 tipping points, 179, 192, 196, 198, 398 Tobin, M. J., 191 top-down control, 95–96, 103–112, 105, 125–126, 143, 242–243, 397 Tornqvist, T., 175 transplant experiments, 29, 140, 174 trematode parasites, 49–50 Trifolium fragiferum, 369 Triglochin maritima, 317 Triglochin striatum, 365 trophic cascades, 104–109, 117, 119, 125–130 trophic mode shift studies, 35 tropical salt marshes, 338–339, 376 tube worm (Ficopomatus engimaticus), 45–46 Turner, R. E., 175, 184, 196, 197 Turtle Cove Experimental Marsh, 121–122 turtlegrass, 232–233 Typha species, 69, 273, 292, 377 Typha angustifolia, 63, 340 Ucides cordatus, 346 ultraviolet-B (UV-B) radiation, 321, 349, 373 Umgiesser, G., 194 United Kingdom Blackwater Estuary, 324 embankment and land claim, 316 grazing impact, 321 managed retreat or setback, 210 marsh degradation, 195 Spartina species, 4, 6, 8, 9, 15–16, 35 Spartina townsendii, 237 urchins, 104 Urosalpinx cinerea, 29 Uruguay, 342 U.S. Army Corps of Engineers, 15, 25, 191 U.S. Fish and Wildlife Service, 71, 96 U.S. Geological Survey, 116, 154, 164, 235, 256 U.S. National Ocean Survey, 182 U.S. National Wildlife Refuges, 88, 164 Valiela, I., 144 Vancouver, Canada, 12 Van der Wal, D., 195 Varekamp, J. C., 180 Vasquez, E. A., 66 Venerupis philippinarum, 50–51 Venice Lagoon, 193–194, 198 Vienna, Maryland, 185 Vince, S. W., 144 Virginia dieback events, 234, 236 marsh degradation, 186, 188 Phragmites australis, 62 Virginia Coast Reserve, 186, 188 Wadden Sea, 314, 317, 322, 324, 325 “Wadden Sea squeeze,” 9 Wang, Y. P., 184 Waquoit Bay, 192

Ward, K. M., 291 Ward, L. G., 181 Warren, R. S., 192 Washington Gray’s Harbor, 12, 14, 63, 290 Phragmites australis, 63 salt marsh area, 286 Willapa Bay, 5, 12–14, 25–32, 36–37, 260, 293 See also Puget Sound, Washington Wasson, K., 42 waterbirds, 288 water diversion, 291, 342 waterfowl, 62, 198, 375 waterlogging, 142, 151, 152, 160, 161, 241, 394 water rat, 375 watershed utilization and management, 298, 368–370 water table, 142, 146, 151, 152, 156, 160, 161, 174, 196, 291 Watford, E. A., 377 Webster, P. J., 348 Weinstein, M. P., 274 Weis, J. S., 68 Wellfleet, Massachusetts, 244 Wellfleet Bay Wildlife Sanctuary, 179 West, R. J., 374 West coast. See Pacific Coast, of North America Western Canada geese, 31 Western Port Bay, 174 Westerschelde, 317 wetland inventories, 254, 261, 301 wetlands in Australia, 376 invasive species susceptibility, 24, 58 loss of, 116, 172, 173, 256 mitigated/created, 71 in New Zealand, 368 See also Phragmites australis; specific index entries Wharton Point, 192 White, T. C. R., 125 Whitlow, W. L., 51 wild rice, 63 Willapa Bay, Washington, 5, 12–14, 25–32, 36–37, 260, 293 Williams, J., 51 Williams, P. B., 294 Williams, R., 374 Williams, R. J., 374, 377 Williamson, R. C., 379 Wilson, J. B., 365 Wilsonia, 374 Windham, L., 69 Wolfe, J. L., 127 Wolf Glad marsh, 188 Wood, M. E., 192, 193 woody shrub, 140 Wraight, M. J., 369 Xeromys myoides, 375 Yang, S. L., 179, 185 Yangtze River, 184–185 York River, 155, 156, 186 Yuhas, C. E., 68 Zedler, J. B., 291, 297 Zhang, R., 184 Zieman, J. C., 186 Zipperer, V. T., 34 Zizania aquatica, 63 zonation, 140 Zostera japonica, 26, 31, 32, 398 Zostera marina, 32, 232 Zostera spp., 364

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