Forest Ecology [5 ed.] 1119476089, 9781119476085

Table of contents :
Cover
Title Page
Copyright Page
Dedication Page
Contents
Preface
Part 1 Forest Ecology and Landscape Ecosystems
CHAPTER 1 Concepts of Forest Ecology
Ecology
Landscape Ecosystems
LANDSCAPE ECOSYSTEM AND COMMUNITY
ECOSYSTEM STRUCTURE AND FUNCTION
Examples of Landscape Ecosystems
An Approach to the Study of Forest Ecology
Applicability to Forest Management
Suggested Readings
CHAPTER 2 Landscape Ecosystems at Multiple Scales
Overview of Spatial and Temporal Scales
Spatial Scales of Hierarchical Landscape Ecosystems
CLIMATIC CLASSIFICATION
PHYSIOGRAPHY
Vegetation Types and Biomes
Distinguishing and Mapping Landscape Ecosystems at Multiple Spatial Scales
REGIONAL LANDSCAPE ECOSYSTEMS
Regional Landscape Ecosystems of Michigan
Local Landscape Ecosystems
LOCAL LANDSCAPE ECOSYSTEMS IN UPPER MICHIGAN
Suggested Readings
Part 2 The Forest Tree
CHAPTER 3 Forest Tree Variation
Components of Phenotypic Variation
PLASTICITY OF THE PHENOTYPE
Sources of Variation
The Evolutionary Sequence
SEXUAL AND ASEXUAL SYSTEMS
Genetic Diversity of Woody Species
Genecology
PATTERNS OF GENECOLOGICAL DIFFERENTIATION
FACTORS ELICITING GENECOLOGICAL DIFFERENTIATION
EXAMPLES OF GENECOLOGICAL DIFFERENTIATION
FACTORS AFFECTING DIFFERENTIATION: GENE FLOW AND SELECTION PRESSURE
Ecological Considerations at the Species Level
NICHE
HYBRIDIZATION
POLYPLOIDY
THE FITNESS–FLEXIBILITY COMPROMISE
EPIGENETICS
Suggested Readings
CHAPTER 4 Regeneration Ecology
Regeneration
Sexual Reproduction
MATURATION AND THE ABILITY TO FLOWER
REPRODUCTIVE CYCLES
POLLINATION
PERIODICITY OF SEED CROPS
EFFECTS OF REPRODUCTION ON VEGETATIVE GROWTH
DISPERSAL
SEED BANK, DORMANCY, AND GERMINATION
Establishment Following Sexual Reproduction
POST-ESTABLISHMENT DEVELOPMENT
Vegetative Reproduction
Suggested Readings
CHAPTER 5 Tree Structure and Growth
Tree Form
ARCHITECTURAL MODELS
Roots
KINDS, FORMS, AND OCCURRENCE
FINE ROOT RELATIONS
HORIZONTAL AND VERTICAL ROOT DEVELOPMENT
PERIODICITY OF PRIMARY ROOT GROWTH
ROOT GRAFTING
SPECIALIZED ROOTS AND BUTTRESSES
Stems
XYLEM CELLS AND GROWTH RINGS
PERIODICITY AND CONTROL OF SECONDARY GROWTH
WINTER FREEZING AND WATER TRANSPORT
Water Deficits and Tree Growth
Suggested Readings
Part 3 The Physical Environment
CHAPTER 6 Light
Distribution of Light Reaching the Ecosphere
PLANT INTERCEPTION OF RADIATION
CANOPY STRUCTURE AND LEAF AREA
Light Quality Beneath the Forest Canopy
SUNFLECKS
Light and Growth of Trees
LIGHT AND SEEDLING SURVIVAL AND GROWTH
Light and Tree Morphology and Anatomy
LIGHT AND EPICORMIC SPROUTING
Photocontrol of Plant Response
Light and Ecosystem Change
Suggested Readings
CHAPTER 7 Temperature
Geographical Patterns Of Temperature
Temperatures at the Soil Surface
Temperature within the Forest
Temperature Variation with Local Topography
Temperature and Plant Growth
COLD INJURY TO PLANTS
DORMANCY
FROST HARDINESS AND COLD RESISTANCE
WINTER CHILLING AND GROWTH RESUMPTION
Natural Plant Distributions and Cold Hardiness
DECIDUOUSNESS AND TEMPERATURE
Suggested Readings
CHAPTER 8 Physiography
Concepts And Terms
Characteristics of Physiography and their Significance
PHYSIOGRAPHIC SETTING
SPECIFIC LANDFORMS
ELEVATION
FORM OF LANDFORMS
PARENT MATERIAL IN RELATION TO LANDFORM
POSITION OF LANDFORM IN THE LANDSCAPE
Multiple Roles of Physiography
Physiographic Diversity, Landscape Ecosystems, and Vegetation
MOUNTAINOUS PHYSIOGRAPHY
FLATLANDS
Physiography and Firebreaks
Microlandforms and Microtopography
TREE UPROOTING AND PIT-AND-MOUND MICROTOPOGRAPHY
MICROTOPOGRAPHY AND REGENERATION IN HARDWOOD SWAMPS
Suggested Readings
CHAPTER 9 Soil
Parent Material
Soil Formation
SOIL PROFILE DEVELOPMENT
Physical Properties of Soil
SOIL TEXTURE
SOIL STRUCTURE
SOIL COLOR
SOIL WATER
Chemical Properties of Soil
CLAY MINERALOGY
CATION EXCHANGE AND THE SUPPLY OF NUTRIENTS
SOIL ACIDITY
SOIL ORGANIC MATTER
Soil Classification
Landform, Soil, and Forest Vegetation: Landscape Relationships
Suggested Readings
CHAPTER 10 Fire
Fire and the Forest Tree
CAUSES
FIRE REGIME
FIRE ADAPTATIONS AND KEY CHARACTERISTICS
STRATEGIES OF SPECIES PERSISTENCE
Fire and the Forest Site
INDIRECT EFFECTS
DIRECT EFFECTS
ORGANIC MATTER AND EROSION
BENEFICIAL EFFECTS OF FIRE
Suggested Readings
CHAPTER 11 Site Quality and Ecosystem Evaluation and Classification
Direct Measurement Of Forest Productivity
Tree Height as a Measure of Site
SITE-INDEX CURVES
COMPARISONS BETWEEN SPECIES
ADVANTAGES AND LIMITATIONS
Vegetation as an Indicator of Site Quality
SPECIES GROUPS OF GROUND COVER
PLANT ASSOCIATIONS AND HABITAT TYPES IN THE WESTERN UNITED STATES
APPLICATIONS AND LIMITATIONS OF VEGETATION
Environmental Factors as a Measure of Site
CLIMATIC FACTORS
PHYSIOGRAPHIC LAND CLASSIFICATION
PHYSIOGRAPHIC AND SOIL FACTORS: SOIL-SITE STUDIES
SOIL SURVEYS
Multiple-Factor Methods of Site and Ecosystem Classification
ECOSYSTEM CLASSIFICATION AND MAPPING IN BADEN-WÜRTTEMBERG
APPLICATIONS OF MULTI-FACTOR METHODS IN THE UNITED STATES AND CANADA
Suggested Readings
Part 4 Forest Communities
CHAPTER 12 Animals in Forest Ecosystems
Plant Defense
INVESTMENT IN PLANT DEFENSE
PLANT DEFENSE AGAINST INSECTS
PLANT DEFENSE AGAINST MAMMALS
Roles of Animals in Plant Life History
POLLINATION
SEED DISPERSAL
GERMINATION AND ESTABLISHMENT
DECOMPOSITION, MINERAL CYCLING, AND SOIL IMPROVEMENT
DAMAGE AND DEATH
Influence of Livestock on Forest Ecosystems
Suggested Readings
CHAPTER 13 Forest Communities
Community Concept
GROUNDING COMMUNITIES
View from the Past: Community Concepts
SCHOOLS AND TERMINOLOGY
CONTINUUM CONCEPT
Community as a Landscape Ecosystem Property
Examples of Spatial Variation in Forest Communities
DISCRETE FOREST COMMUNITIES
MERGING FOREST COMMUNITIES
Competition and Niche Differentiation
Interactions Among Organisms
MUTUALISMS IN FOREST ECOSYSTEMS
COMPETITION
Understory Tolerance
CHARACTERISTICS OF UNDERSTORY-TOLERANT AND -INTOLERANT SPECIES
TOLERANCE RATINGS OF TREE SPECIES
EXAMPLES OF UNDERSTORY TOLERANCE IN FOREST ECOSYSTEMS
NATURE OF UNDERSTORY TOLERANCE
Suggested Readings
CHAPTER 14 Diversity in Forests
Concepts Of Biological And Ecosystem Diversity
The Value of Species Diversity
VALUE OF BIODIVERSITY
COMMON THREATS TO DIVERSITY
Measuring Diversity
LEVELS OF DIVERSITY
MEASUREMENT
DIVERSITY OF LANDSCAPE ECOSYSTEMS
EXAMPLES OF DIVERSITY
ECOSYSTEM DIVERSITY
Causes of Species Diversity
DIVERSITY AT CONTINENTAL AND SUBCONTINENTAL SCALES
DIVERSITY AT LOCAL SCALES
Focal Species In Conserving Diversity
FOUNDATION SPECIES
KEYSTONE SPECIES
ENDEMICS AND RARE AND ENDANGERED SPECIES
Diversity and the Functioning of Ecosystems
BIODIVERSITY–PRODUCTIVITY RELATIONSHIP
THE ROLE OF BIODIVERSITY IN ECOSYSTEM STABILITY
Forest Management and Diversity
EFFECTS OF TRADITIONAL FOREST MANAGEMENT ON DIVERSITY
PRESERVING DIVERSITY IN MANAGED FORESTS
ECOLOGICAL FORESTRY: INCORPORATING BIODIVERSITY INTO FOREST MANAGEMENT
Epilog: Conserving Ecosystem and Biological Diversity
Suggested Readings
Part 5 Forest Ecosystem Dynamics
CHAPTER 15 Long-Term Forest Ecosystem and Vegetation Change
Change Before the Pleistocene Age
Pleistocene Glaciations
Ecosystem and Vegetational Change Since the Last Glacial Maximum
EASTERN NORTH AMERICA
WESTERN NORTH AMERICA
Patterns of Tree Genera and Species Migrations
MIGRATION IRREGULARITIES AND DISTURBANCE
MIGRATION FROM GLACIAL MICROREFUGIA
Independent Migration and Similarity of Communities Through Time
Suggested Readings
CHAPTER 16 Disturbance
Concepts of Disturbance
DEFINING A DISTURBANCE
DISTURBANCE AS AN ECOSYSTEM PROCESS
SOURCEs OF DISTURBANCE
Major Disturbances in Forest Ecosystems
FIRE
WIND
FLOODWATER AND ICE STORMS
INSECTS AND DISEASE
CATASTROPHIC AND LOCAL LAND MOVEMENTS
LOGGING
LAND CLEARING
DISTURBANCE INTERACTIONS
Biotic Composition Changes
ELIMINATION OF SPECIES
ADDITION OF SPECIES
Suggested Readings
CHAPTER 17 Forest Succession
Basic Concepts of Succession
PRIMARY AND SECONDARY SUCCESSION
BIOLOGICAL LEGACIES
SUCCESSIONAL PATHWAYS, MECHANISMS, AND MODELS
AUTOGENIC AND ALLOGENIC SUCCESSION
HOW IS SUCCESSION DETERMINED?
Evolution of the Concept of Forest Succession
FORMAL ECOLOGICAL THEORY
How Does Succession Work?
CLEMENTSIAN SUCCESSION
STAGES OF SUCCESSION
SUCCESSIONAL CAUSES, MECHANISMS, AND MODELS
Change in Ecosystems
END POINT OF SUCCESSION?
Succession as an Ecosystem Process
Examples of Forest Succession
PRIMARY SUCCESSION ON RECENTLY DEGLACIATED TERRAIN
SUCCESSION FOLLOWING THE ERUPTION OF MOUNT ST. HELENS
SECONDARY SUCCESSION FOLLOWING FIRE IN PONDEROSAPINE FORESTS OF WESTERN MONTANA
SECONDARY SUCCESSION AND GAP DYNAMICS
FIRE AND OAK DOMINANCE—OAKS AT RISK
Suggested Readings
CHAPTER 18 Carbon Balance of Trees and Ecosystems
Carbon Balance of Trees
PHOTOSYNTHESIS, DARK RESPIRATION, AND LEAF C GAIN
LIGHT AND LEAF C GAIN
TEMPERATURE AND LEAF C GAIN
WATER AND LEAF C GAIN
SOIL NITROGEN AVAILABILITY AND LEAF C GAIN
CONSTRUCTION AND MAINTENANCE RESPIRATION
ALLOCATION TO STRUCTURE, STORAGE, AND DEFENSE
LIGHT AND C ALLOCATION
SOIL NITROGEN AVAILABILITY AND C ALLOCATION
Carbon Balance of Ecosystems
BIOMASS AND PRODUCTIVITY OF FOREST ECOSYSTEMS
MEASUREMENT OF BIOMASS AND PRODUCTIVITY
CLIMATE AND PRODUCTIVITY
SOIL PROPERTIES, FOREST BIOMASS, AND ANPP
BIOMASS ACCUMULATION DURING ECOSYSTEM DEVELOPMENT
Soil N Availability and Belowground Net Primary Productivity
Suggested Readings
CHAPTER 19 Nutrient Cycling
Nutrient Additions to Forest Ecosystems
MINERAL WEATHERING
ATMOSPHERIC DEPOSITION
BIOLOGICAL FIXATION OF NITROGEN
Nutrient Cycling within Forest Ecosystems
NUTRIENT TRANSPORT TO ROOTS
NUTRIENT UPTAKE AND ASSIMILATION BY ROOTS
ROOT ARCHITECTURE, MYCORRHIZAE, AND NUTRIENT ACQUISITION
PLANT LITTER AND THE RETURN OF NUTRIENTS TO FOREST FLOOR AND SOIL
NUTRIENTS IN THE FOREST FLOOR
ORGANIC MATTER DECOMPOSITION AND NUTRIENT MINERALIZATION
Nutrient Loss From Forest Ecosystems
NUTRIENT LEACHING FROM FOREST ECOSYSTEMS
DENITRIFICATION
The Cycling and Storage of Nutrients in Forest Ecosystems
NUTRIENT STORAGE IN BOREAL, TEMPERATE, AND TROPICAL FORESTS
THE NITROGEN AND CALCIUM CYCLE OF A TEMPERATE FOREST ECOSYSTEM
Ecosystem C Balance and the Retention and Loss of Nutrients
Forest Harvesting and Nutrient Loss
Suggested Readings
Part 6 Forests of the Future
CHAPTER 20 Climate Change and Forest Ecosystems
Climate Change Concepts
EFFECTS ON TEMPERATURE
EFFECTS ON PRECIPITATION
Climate Change Effects on the Forest Tree
TREE GROWTH AND MORTALITY
PHENOLOGY
REGENERATION
Climate Change Effects on Tree Species Distributions
OBSERVED RANGE SHIFTS
PROJECTED CHANGES IN TREE SPECIES DISTRIBUTIONS
PROJECTED CHANGES IN FOREST TYPE DISTRIBUTIONS
Climate Change Effects on Forest Disturbances
FIRES
INSECTS AND PATHOGENS
WIND
Climate Change Effects on Forest Carbon
CLIMATE CHANGE EFFECTS ON CARBON GAIN: PRIMARY PRODUCTIVITY
CLIMATE CHANGE EFFECTS ON CARBON LOSS: HETEROTROPHIC RESPIRATION
FEEDBACKs AMONG DISTURBANCE, CLIMATE CHANGE, AND CARBON IN FORESTS
Adapting to Climate Change Effects on Forests
ASSISTED MIGRATION
REFUGIA
FOREST CARBON MANAGEMENT
Suggested Readings
CHAPTER 21 Invasive Species in Forest Ecosystems
Concepts Of Invasive Species
DEFINITION OF INVASIVE SPECIES
CHARACTERISTIC TRAITS OF INVASIVE PLANT SPECIES
NON-PLANT INVASIVE SPECIES IN FORESTS
Impacts Of Invasive Species On Forests
IMPACTS OF INVASIVE PLANTS ON FORESTS
IMPACTS OF INVASIVE INSECTS AND PATHOGENS ON FORESTS
IMPACTS OF INVASIVE ANIMALS ON FORESTS
A Primer Of Invasive Species Management In Forests
EARLY INTERVENTION STRATEGIES
MANAGEMENT APPROACHES FOR ESTABLISHED INVASIVE SPECIES
NOVEL ECOSYSTEMS AND INVASIVE SPECIES
Suggested Readings
CHAPTER 22 Forest Landscape Ecology
Concepts Of Landscape Ecology
Forest Fragmentation And Connectivity
PATCHES IN FOREST ECOLOGY
FOREST FRAGMENTATION
Ecological Effects
CONNECTIVITY
Disturbances On Landscapes
EFFECTS OF HETEROGENEITY ON DISTURBANCES
EFFECTS OF DISTURBANCES ON HETEROGENEITY
HISTORICAL RANGE OF VARIABILITY
Interactions Of Landscape Patterns And Ecological Processes
LEAF AREA AND PRODUCTIVITY
FOREST CARBON DYNAMICS
NUTRIENT DYNAMICS
Suggested Readings
CHAPTER 23 Sustainability of Forest Ecosystems
Concepts of Sustainability
THE PREVALENCE OF HUMAN VALUES IN FOREST ECOLOGY
HISTORICAL PERSPECTIVE OF SUSTAINABILITY IN FORESTS
ECOSYSTEM SERVICES
TOWARD A DEFINITION OF SUSTAINABILITY
Where Do We Go From Here?
EPILOG: EARTH AS A METAPHOR FOR LIFE
Suggested Readings
References
Scientific Names of Trees and Shrubs
Index
EULA

Citation preview

Forest Ecology

Forest Ecology

Daniel M. Kashian Wayne State University, Detroit, USA

Donald R. Zak University of Michigan, Ann Arbor, USA

Burton V. Barnes (deceased) University of Michigan, Ann Arbor, USA

Stephen H. Spurr (deceased) University of Texas at Austin, Austin, USA

FIFTH EDITION

This fifth edition first published 2023 © 2023 John Wiley & Sons Ltd Edition History John Wiley & Sons Ltd. (4e, 1998); Ronald Press (3e, 1980; 2e, 1973; 1e, 1964) All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording, or otherwise, except as permitted by law. Advice on how to obtain permission to reuse material from this title is available at http://www.wiley.com/go/ permissions. The right of Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr to be identified as the authors of this work has been asserted in accordance with law. Registered Offices John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, USA John Wiley & Sons Ltd., The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK Editorial Office 9600 Garsington Road, Oxford, OX4 2DQ, UK For details of our global editorial offices, customer services, and more information about Wiley products, visit us at www.wiley.com. Wiley also publishes its books in a variety of electronic formats and by print-­on-­demand. Some content that appears in standard print versions of this book may not be available in other formats. Limit of Liability/Disclaimer of Warranty While the publisher and authors have used their best efforts in preparing this work, they make no representations or warranties with respect to the accuracy or completeness of the contents of this work and specifically disclaim all warranties, including without limitation any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives, written sales materials, or promotional statements for this work. The fact that an organization, website, or product is referred to in this work as a citation and/or potential source of further information does not mean that the publisher and authors endorse the information or services the organization, website, or product may provide or recommendations it may make. This work is sold with the understanding that the publisher is not engaged in rendering professional services. The advice and strategies contained herein may not be suitable for your situation. You should consult with a specialist where appropriate. Further, readers should be aware that websites listed in this work may have changed or disappeared between when this work was written and when it is read. Neither the publisher nor authors shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. Library of Congress Cataloging-­in-­Publication Data Names: Kashian, Daniel M., author. | Zak, Donald R., author. | Barnes, Burton Verne, 1930-2014, author. | Spurr, Stephen H., 1918-1990, author. Title: Forest ecology / Daniel M. Kashian, Wayne State University, Detroit, USA, Donald R. Zak, University of Michigan, Ann Arbor, USA, Burton V. Barnes (deceased), University of Michigan, Ann Arbor, USA, Stephen H. Spurr (deceased), University of Texas at Austin, Austin, USA. Description: Fifth edition. | Hoboken, NJ : Wiley, 2022. | Precedey by: Forest ecology / Burton V. Barnes ... [et al.]. 4th ed. c1998. | Includes bibliographical references and index. Identifiers: LCCN 2022037810 (print) | LCCN 2022037811 (ebook) | ISBN 9781119476085 (paperback) | ISBN 9781119476054 (adobe pdf) | ISBN 9781119476146 (epub) Subjects: LCSH: Forest ecology. Classification: LCC QK938.F6 F635 2022 (print) | LCC QK938.F6 (ebook) | DDC 577.3–dc23/eng/20220831 LC record available at https://lccn.loc.gov/2022037810 LC ebook record available at https://lccn.loc.gov/2022037811 Cover Design: Wiley Cover Image: © U.S. Forest Service Photo Set in 9.5/12pt STIXTwoText by Straive, Pondicherry, India

Dedication We dedicate the 5th edition of Forest Ecology to the co-­author of the 2nd and 3rd editions and the lead author of the 4th edition, Burton V. Barnes (1930–2014). It would be no easy task to find a more accomplished and humble leader in his field. He excelled at his science, was a truly beloved teacher, and helped to shape the world view of thousands of colleagues, friends, students, managers, and scientists alike, all with an unmatched humor and a love of the natural world.

Burt Barnes was world-­renowned as an expert in the ecology of North American aspens and the ecological classification of forest ecosystems. His professional training was in forest ecology, botany, and genetics, but he dabbled heavily in glacial geomorphology, soil science, phytogeography, and woody plant physiology. Perhaps his greatest love, however, was teaching, especially in the field, which drove his motivation for this textbook. Generations of students have been touched by his love for the art and science of teaching field ecology, which they will forever pass on to future generations. We, as authors of this edition, have been personally and professionally shaped by him as a mentor, colleague, and friend. His legacy is therefore unending. To him we say, in his own words, “Thanks for everything you have done—­and will do.”

Contents Prefacexxiii PA R T 1 Forest Ecology and Landscape Ecosystems 1 Concepts of Forest Ecology

3

Ecology, 4 Landscape Ecosystems, 4 Landscape Ecosystem and Community, 7 Ecosystem Structure and Function, 7 Examples of Landscape Ecosystems, 9 An Approach to The Study of Forest Ecology, 11 Applicability to Forest Management, 12 Suggested Readings, 13

2 Landscape Ecosystems at Multiple Scales

15

Overview of Spatial and Temporal Scales, 15 Spatial Scales of Hierarchical Landscape Ecosystems, 17 Climatic Classification, 18 Physiography, 20 Vegetation Types and Biomes, 20 Distinguishing and Mapping Landscape Ecosystems at Multiple Spatial Scales, 25 Regional Landscape Ecosystems, 26 Regional Landscape Ecosystems of Michigan, 28 Local Landscape Ecosystems, 30 Local Landscape Ecosystems in Upper Michigan, 30 Suggested Readings, 32

vii

viii  C o n t e n ts

PA R T 2 The Forest Tree 3 Forest Tree Variation

35

Components of Phenotypic Variation, 35 Plasticity of the Phenotype, 36 Sources of Variation, 37 The Evolutionary Sequence, 38 Sexual and Asexual Systems, 39 Genetic Diversity of Woody Species, 39 Genecology, 40 Patterns of Genecological Differentiation, 41 Genecological Categories, 42 Factors Eliciting Genecological Differentiation, 42 Growth Cessation, 43 Growth Resumption, 46 Examples of Genecological Differentiation, 46 Eastern North American Species, 47 Scots Pine, 49 Wide-Ranging Western North American Conifers, 50 Ponderosa pine, 50 Douglas-­fir, 52 Local Genecological Differentiation, 55 Factors Affecting Differentiation: Gene Flow and Selection Pressure, 56 Ecological Considerations at the Species Level, 57 Niche, 60 Hybridization, 60 Polyploidy, 63 The Fitness–Flexibility Compromise, 66 Epigenetics, 66 Suggested   Readings, 67

4 Regeneration Ecology Regeneration, 69 Sexual Reproduction, 71 Maturation and the Ability to Flower, 72 Increasing Seed Production, 73 Reproductive Cycles, 74 Pollination, 75

69

Contents  ix

Periodicity of Seed Crops, 76 Effects of Reproduction on Vegetative Growth, 78 Dispersal, 80 Seed Bank, Dormancy, and Germination, 82 Establishment Following Sexual Reproduction, 83 Post-­Establishment Development, 90 Vegetative Reproduction, 90 Suggested Readings, 94

5 Tree Structure and Growth

95

Tree Form, 95 Architectural Models, 97 Short and Long Shoots, 99 Patterns of Intermittent Growth, 100 Sylleptic and Proleptic Shoots, 102 Roots, 102 Kinds, Forms, and Occurrence, 103 Fine Root Relations, 105 Horizontal and Vertical Root Development, 107 Periodicity of Primary Root Growth, 108 Root Grafting, 108 Specialized Roots and Buttresses, 110 Stems, 110 Xylem Cells and Growth Rings, 111 Periodicity and Control of Secondary Growth, 112 Control of Earlywood and Latewood Formation, 114 Winter Freezing and Water Transport, 115 Water Deficits and Tree Growth, 116 Suggested Readings, 118

PA R T 3 The Physical Environment Forest Environment, 119 Site Factors, 120 Organization of Site Factors, 120 Interrelationships Among Site Factors, 121 Importance of Site in Forest Ecology, 122

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x  C o n t e n ts

6 Light

123

Distribution of Light Reaching the Ecosphere, 124 Plant Interception of Radiation, 125 Canopy Structure and Leaf Area, 125 Light Quality Beneath the Forest Canopy, 130 Sunflecks, 130 Light and Growth of Trees, 133 Light and Seedling Survival and Growth, 136 Light and Tree Morphology and Anatomy, 137 Light and Epicormic Sprouting, 139 Photocontrol of Plant Response, 139 Light and Ecosystem Change, 141 Suggested   Readings, 141

7 Temperature

143

Geographical Patterns of Temperature, 143 Temperatures at the Soil Surface, 145 Temperature within the Forest, 147 Temperature Variation with Local Topography, 148 Temperature and Plant Growth, 150 Cold Injury to Plants, 153 Dormancy, 155 Frost Hardiness and Cold Resistance, 156 Thermotropic Movements in Rhododendrons, 159 Winter Chilling and Growth Resumption, 161 Natural Plant Distributions and Cold Hardiness, 162 Deciduousness and Temperature, 164 Suggested Readings, 165

8 Physiography Concepts and Terms, 167 Characteristics of Physiography and their Significance, 168 Physiographic Setting, 169 Specific Landforms, 170 Elevation, 170 Form of Landforms, 170 Level Terrain, 170 Sloping Terrain, 171 Slope Characteristics, 171

167

Contents  xi

Position on slope, 172 Aspect, 172 Slope inclination, 173 Parent Material in Relation to Landform, 173 Position of Landform in the Landscape, 174 Multiple Roles of Physiography, 174 Physiographic Diversity, Landscape Ecosystems, and Vegetation, 176 Mountainous Physiography, 176 Mountainous Terrain of California and the Pacific Northwest, 177 Physiography and Forests of the Central Appalachians, 180 Flatlands, 183 The Great Plains, 184 Pine Savannas of the Western Great Lakes Region, 185 Till Plains of the Midwest, 185 Southeastern and Southern Coastal Plain, 186 Floodplains, 186 Physiography and Firebreaks, 190 Microlandforms and Microtopography, 191 Tree Uprooting and Pit-­and-­Mound Microtopography, 191 Microtopography and Regeneration in Hardwood Swamps, 193 Suggested Readings, 194

9 Soil

195

Parent Material, 195 Soil Formation, 197 Soil Profile Development, 197 Physical Properties of Soil, 200 Soil Texture, 200 Soil Structure, 201 Soil Color, 202 Soil Water, 202 Physical Properties of Water, 203 Soil Water Potential, 204 Chemical Properties of Soil, 206 Clay Mineralogy, 207 Cation Exchange and the Supply of Nutrients, 209 Soil Acidity, 210 Soil Organic Matter, 212 Soil Classification, 213 Landform, Soil, and Forest Vegetation: Landscape Relationships, 215 Suggested Readings, 217

xii  C o n t e n ts

10 Fire

219

Fire and the Forest Tree, 220 Causes, 220 Fire Regime, 220 Fire Types, Frequency, and Severity, 221 Fire Adaptations and Key Characteristics, 224 Strategies of Species Persistence, 227 Closed-­Cone Pines, 228 Fire and the Forest Site, 230 Indirect Effects, 230 Direct Effects, 232 Organic Matter and Erosion, 234 Beneficial Effects of Fire, 236 Suggested Readings, 236

11 Site Quality and Ecosystem Evaluation and Classification237 Direct Measurement of Forest Productivity, 238 Tree Height as a Measure of Site, 239 Site-­Index Curves, 240 Comparisons Between Species, 243 Advantages and Limitations, 243 Vegetation as an Indicator of Site Quality, 244 Species Groups of Ground Cover, 245 Indicator Plants of Coastal British Columbia, 246 Ecological Species Groups, 246 Plant Associations and Habitat Types in the Western United States, 247 Operational Site Classification Based on Vegetation, 247 Applications and Limitations of Vegetation, 250 Environmental Factors as a Measure of Site, 251 Climatic Factors, 251 Physiographic Land Classification, 252 Physiographic and Soil Factors: Soil-­Site Studies, 252 Soil Surveys, 256 Multiple-­Factor Methods of Site and Ecosystem Classification, 256 Ecosystem Classification and Mapping in Baden-­Württemberg, 257 Applications of Multi-­Factor Methods in the United States and Canada, 258 Ecosystem Classification and Mapping in Michigan, 258 Ecosystem Classification in the Southeastern United States, 261

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Contents  xiii

Ecosystem Classification in the Southwestern United States, 262 National Classification Systems in the United States, 262 Ecological Land Classification in Canada, 263 Hills’ physiographic approach, 263 Other approaches used in Canada, 264 Suggested Readings, 265

PA R T 4 Forest Communities 12 Animals in Forest Ecosystems

269

Plant Defense, 269 Investment in Plant Defense, 270 Plant Defense Against Insects, 272 Examples of Injury and Plant Defense, 273 Nutrition, 274 Plant Hybrid Zones as Reservoirs for Insect Diversity, 275 Plant Defense Against Mammals, 276 Roles of Animals in Plant Life History, 277 Pollination, 277 Seed Dispersal, 278 Fish and Reptiles, 278 Birds, 279 Mammals, 282 Germination and Establishment, 284 Decomposition, Mineral Cycling, and Soil Improvement, 285 Damage and Death, 286 Influence of Livestock on Forest Ecosystems, 288 Suggested Readings, 290

13 Forest Communities

291

Community Concept, 291 Grounding Communities, 292 Florida Keys, 292 Interior Alaska, 294 Southern Illinois, 294 View from the Past: Community Concepts, 296 Schools and Terminology, 296 Concepts of Clements and Gleason, 297

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xiv  C o n t e n ts

Phytosociology in Europe, 298 Continuum Concept, 298 Community as a Landscape Ecosystem Property, 299 Examples of Spatial Variation in Forest Communities, 300 Discrete Forest Communities, 300 Coastal California: Giant and Pygmy Forests, 301 Forest–Grassland Ecotone, 302 Alpine Tree Lines, 302 Merging Forest Communities, 303 Eastern Deciduous Forest—­Southern Appalachians, 303 New England, 304 Competition and Niche Differentiation, 305 Interactions Among Organisms, 307 Mutualisms in Forest Ecosystems, 307 Symbiotic Mutualisms—­Mycorrhizae, 307 Nonsymbiotic Mutualisms, 307 Competition, 308 Forest Community Structure and Composition, 308 Vertical Structure, 310 Stand Density, 312 Competition and Overstory Composition, 313 Competition in the Understory, 314 Understory Tolerance, 315 Characteristics of Understory-­Tolerant and -­Intolerant Species, 316 Tolerance Ratings of Tree Species, 317 Examples of Understory Tolerance in Forest Ecosystems, 319 Nature of Understory Tolerance, 319 Environmental Factors Relating to Understory Tolerance, 322 Physiological Processes Relating to Tolerance, 323 Suggested Readings, 325

14 Diversity in Forests

327

Concepts of Biological and Ecosystem Diversity, 327 The Value of Species Diversity, 328 Value of Biodiversity, 329 Common Threats to Diversity, 331 Measuring Diversity, 331 Levels of Diversity, 331 Measurement, 333 Inventory Diversity: Alpha Diversity, 333 Differentiation or Beta Diversity, 335

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Contents  xv

Diversity of Landscape Ecosystems, 336 Examples of Diversity, 337 Ground-­Cover Species Diversity in Northern Lower Michigan, 337 Ecosystem Groups, 337 Ecosystem Types, 340 Ecosystem Diversity, 342 Causes of Species Diversity, 343 Diversity at Continental and Subcontinental Scales, 343 Paleogeography and Continental Relationships, 343 Glaciation, 344 Latitude and Elevation, 344 Diversity at Local Scales, 346 Physiography and Soil, 346 Community Composition and Structure, 349 Disturbance and Succession, 349 Focal Species in Conserving Diversity, 351 Foundation Species, 352 Keystone Species, 352 Endemics and Rare and Endangered Species, 353 Diversity and the Functioning of Ecosystems, 354 Biodiversity–Productivity Relationship, 354 The Role of Biodiversity in Ecosystem Stability, 356 Forest Management and Diversity, 357 Effects of Traditional Forest Management on Diversity, 357 Preserving Diversity in Managed Forests, 358 Ecological Forestry: Incorporating Biodiversity into Forest Management, 359 Variable-­Retention Harvest System, 360 Designing a Variable-­Retention Harvest System, 361 How Well Does Variable Retention Conserve Biodiversity?, 361 Epilog: Conserving Ecosystem and Biological Diversity, 363 Suggested Readings, 364

PA R T 5 Forest Ecosystem Dynamics 15 Long-­Term Forest Ecosystem and Vegetation Change

367

Change Before the Pleistocene Age, 367 Pleistocene Glaciations, 368 Ecosystem and Vegetational Change Since the Last ­Glacial Maximum, 370

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xvi  C o n t e n ts

Eastern North America, 370 Overall Migration Sequence and Patterns, 371 Ecosystem Change in the Southern Appalachians, 373 Western North America, 373 Patterns of Tree Genera and Species Migrations, 375 Migration Irregularities and Disturbance, 377 Migration From Glacial Microrefugia, 377 Independent Migration and Similarity of Communities Through Time, 379 Suggested Readings, 380

16 Disturbance

381

Concepts of Disturbance, 382 Defining a Disturbance, 382 Disturbance as an Ecosystem Process, 384 Sources of Disturbance, 386 Major Disturbances in Forest Ecosystems, 387 Fire, 387 Roles of Fire in Forest Ecosystems, 387 Pines in New England and the Lake States, 389 Western Pines and Trembling Aspen, 390 Southern Pines, 393 Douglas-­Fir in the Pacific Northwest, 394 Giant Sequoia, 394 Fire History and Behavior, 395 Northern Lake States, 395 Boreal Forest and Taiga, 397 Northern Rocky Mountains, 398 Fire Suppression and Exclusion, 399 Wind, 400 Widespread and Local Effects, 401 Principles of Wind Damage, 401 Broadscale Disturbance by Hurricanes, 403 Gulf and Southern Atlantic Coasts, 403 New England—­1938 Hurricane, 404 Wave-­Regenerated Fir Species, 404 Floodwater and Ice Storms, 404 Insects and Disease, 406 Catastrophic and Local Land Movements, 407 Logging, 407 Land Clearing, 409 Disturbance Interactions, 409

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Contents  xvii

Biotic Composition Changes, 411 Elimination of Species, 411 Addition of Species, 412 Suggested Readings, 412

17 Forest Succession

413

Basic Concepts of Succession, 414 Primary and Secondary Succession, 414 Biological Legacies, 414 Successional Pathways, Mechanisms, and Models, 414 Autogenic and Allogenic Succession, 416 How is Succession Determined?, 416 Evolution of the Concept of Forest Succession, 417 Formal Ecological Theory, 417 How Does Succession Work?, 418 Clementsian Succession, 419 Stages of Succession, 420 Primary Succession, 420 Secondary Succession, 423 Successional Causes, Mechanisms, and Models, 427 Key Characteristics and Regeneration Strategies, 427 Availability and Arrival Sequence of Species, 428 Facilitation, Tolerance, and Inhibition, 428 Change in Ecosystems, 431 End Point of Succession?, 431 Succession as an Ecosystem Process, 432 Examples of Forest Succession, 434 Primary Succession on Recently Deglaciated Terrain, 434 Succession Following the Eruption of Mount St. Helens, 437 Secondary Succession Following Fire in Ponderosa Pine Forests of Western Montana, 439 Secondary Succession and Gap Dynamics, 439 Gap Specialists: American Beech and Sugar Maple, 444 Fire and Oak Dominance—­Oaks at Risk, 446 Suggested Readings, 447

18 Carbon Balance of Trees and Ecosystems

449

Carbon Balance of Trees, 450 Photosynthesis, Dark Respiration, and Leaf C Gain, 450 Light and Leaf C Gain, 452

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xviii  C o n t e n ts

Temperature and Leaf C Gain, 454 Water and Leaf C Gain, 455 Soil Nitrogen Availability and Leaf C Gain, 456 Construction and Maintenance Respiration, 456 Allocation to Structure, Storage, and Defense, 459 Light and C Allocation, 461 Soil Nitrogen Availability and C Allocation, 461 Carbon Balance of Ecosystems, 465 Biomass and Productivity of Forest Ecosystems, 466 Measurement of Biomass and Productivity, 469 Climate and Productivity, 473 Soil Properties, Forest Biomass, and ANPP, 475 Biomass Accumulation During Ecosystem Development, 477 Soil N Availability and Belowground Net Primary Productivity, 482 Suggested Readings, 485

19 Nutrient Cycling

487

Nutrient Additions to Forest Ecosystems, 488 Mineral Weathering, 488 Atmospheric Deposition, 490 Biological Fixation of Nitrogen, 493 Nutrient Cycling within Forest Ecosystems, 497 Nutrient Transport to Roots, 497 Nutrient Uptake and Assimilation by Roots, 498 Root Architecture, Mycorrhizae, and Nutrient Acquisition, 501 Root Architecture, 501 Mycorrhizae, 502 Plant Litter and the Return of Nutrients to Forest Floor and Soil, 504 Leaf and Root Litter Production, 505 Nutrient Retranslocation, 507 Nutrients in the Forest Floor, 509 Organic Matter Decomposition and Nutrient Mineralization, 512 Biochemical Constituents of Plant Litter, 513 Dynamics of Decomposition, 515 Nitrogen Immobilization and Mineralization, 519 Nitrogen Availability in Forest Ecosystems, 521 Nitrification, 522 Nutrient Loss From Forest Ecosystems, 523 Nutrient Leaching from Forest Ecosystems, 524

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Contents  xix

Denitrification, 525 The Cycling and Storage of Nutrients in Forest Ecosystems, 527 Nutrient Storage in Boreal, Temperate, and Tropical Forests, 527 The Nitrogen and Calcium Cycle of a Temperate Forest Ecosystem, 528 Ecosystem C Balance and the Retention and Loss of Nutrients, 530 Forest Harvesting and Nutrient Loss, 532 Suggested Readings, 534

PA R T 6 Forests of the Future 20 Climate Change and Forest Ecosystems

537

Climate Change Concepts, 537 Effects on Temperature, 540 Effects on Precipitation, 543 Climate Change Effects on the Forest Tree, 546 Tree Growth and Mortality, 546 Phenology, 548 Regeneration, 551 Climate Change Effects on Tree Species Distributions, 554 Observed Range Shifts, 554 Projected Changes in Tree Species Distributions, 556 Projected Changes in Forest Type Distributions, 559 Climate Change Effects on Forest Disturbances, 561 Fires, 563 Insects and Pathogens, 568 Wind, 570 Climate Change Effects on Forest Carbon, 571 Climate Change Effects on Carbon Gain: Primary Productivity, 572 Climate Change Effects on Carbon Loss: Heterotrophic Respiration, 573 Feedbacks Among Disturbance, Climate Change, and Carbon in Forests, 575 Fire, Carbon, and Climate Change in Forests of Yellowstone National Park, 577 Adapting to Climate Change Effects on Forests, 578 Assisted Migration, 579 Refugia, 581 Forest Carbon Management, 583 Suggested Readings, 585

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xx  C o n t e n ts

21 Invasive Species in Forest Ecosystems

587

Concepts of Invasive Species, 587 Definition of Invasive Species, 587 Characteristic Traits of Invasive Plant Species, 589 Non-­Plant Invasive Species in Forests, 594 Impacts of Invasive Species on Forests, 597 Impacts of Invasive Plants on Forests, 597 Competition, 598 Altered Fire Regimes, 598 Carbon and Nutrient Cycling, 599 Impacts of Invasive Insects and Pathogens on Forests, 600 Chestnut Blight, Dutch Elm Disease, and Forest Succession, 601 Impacts of Invasive Animals on Forests, 604 A Primer of Invasive Species Management in Forests, 605 Early Intervention Strategies, 607 Management Approaches for Established Invasive Species, 607 Novel Ecosystems and Invasive Species, 608 Suggested Readings, 610

22 Forest Landscape Ecology

611

Concepts of Landscape Ecology, 612 Forest Fragmentation and Connectivity, 615 Patches in Forest Ecology, 616 Forest Fragmentation, 617 Ecological Effects, 617 Connectivity, 620 Disturbances on Landscapes, 622 Effects of Heterogeneity on Disturbances, 622 Hurricanes in New England, 622 Landscape Pattern Effects on Disturbance Spread, 625 Effects of Disturbances on Heterogeneity, 625 Stand-­Replacing Wildfires in Yellowstone National Park, 627 Historical Range of Variability, 630 Interactions of Landscape Patterns and Ecological Processes, 632 Leaf Area and Productivity, 633 Forest Carbon Dynamics, 635 Nutrient Dynamics, 636 Suggested Readings, 638

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Contents  xxi

23 Sustainability of Forest Ecosystems

639

Concepts of Sustainability, 639 The Prevalence of Human Values in Forest Ecology, 639 Historical Perspective of Sustainability in Forests, 641 Ecosystem Services, 642 Toward a Definition of Sustainability, 644 Where Do We Go From Here?, 646 Epilog: Earth as a Metaphor for Life, 648 Suggested Readings, 650

References651 Scientific Names of Trees and Shrubs

731

Index739

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Preface F

orest Ecology deals with forest ecosystems—­spatial and volumetric segments of the ­Earth—­and their climate, landforms, soils, and biota. It is designed as a textbook for people interested in forest ecosystems—­either in the context of courses in forest ecology and environmental science or as an ecological reference for those in professional practice. This book is meant to provide basic ecological concepts and principles for field ecologists, foresters, naturalists, botanists, and others interested in the conservation and restoration of forest ecosystems. Ecology, in general, has undergone several sea changes since the appearance of the first edition in 1964, with enormous increases in public interest and scientific development of theory and research. Ecology and the issues associated with it have become part of our modern lexicon. The great number of advances in our ecological knowledge, as well as increased public interest, presents forest ecologists with both opportunities and challenges to sustainably manage ­ecosystems using our best understanding of ecosystem properties and processes. This book will hopefully be useful in that process by providing an understanding of the ecological relationships of individual trees and forest ecosystems. The book has six major subdivisions. “Forest Ecology and Landscape Ecosystems” introduces forests as whole ecosystems rather than tree communities, and at multiple scales. “The Forest Tree” considers the genetic variations among individual trees, the causes of diversity within and between species, regeneration ecology, and selected aspects of tree structure and function. “The Physical Environment” treats the physical factors of forest ecosystems that form the forest ­site—­the influences of light, temperature, physiography, soil, and fire on the individual forest plant and on plant communities. The concluding chapter in this part considers methods of evaluation and classification of the forest site and ecosystems. In Part 4, “Forest Communities,” we consider the forest community of trees and associated plants and animals that form a key structural component of forest ecosystems—­one part of the whole. We also consider the importance and measurement of diversity of species and ecosystems. In “Forest Ecosystem Dynamics,” we examine the functional relationships of the physical environment and the biota. We first examine changes in communities and ecosystems over tens of thousands of years. We then consider the extent to which disturbance, an ecosystem process, initiates change (termed succession) over shorter time scales of centuries. Chapters on carbon balance and nutrient cycling present a detailed consideration of the pattern in which carbon (i.e., energy) and plant nutrients flow within forest ecosystems and how natural and human-­induced disturbances alter these patterns. Finally, “Forests of the Future” explores the role of humans in the sustainability of forests. Here we emphasize two of the most pressing issues in forest ecology today, climate change and invasive species, and present a review of landscape ecology which has humans at its center. We end the book with a treatment of sustainability itself. In this edition, we have made great attempts to maintain the core organization and ­readability of previous editions while adding those areas most relevant to forest ecology that have developed over the last quarter-­century. We have retained the important focus on landscape ecosystems (rather than organisms and communities) that was developed in the fourth edition and have added critical new ecological concepts and research that have developed in genetics, diversity, climate change, invasive species, and sustainability. The ecological literature has only become more voluminous over the past 25 years, and as in previous editions our use of the literature was selective, rather than exhaustive. New references were most often chosen based on their accessibility to students

xxiii

xxiv  P r e fac e

and practitioners as understandable examples of important ecological concepts. At the same time, we have retained many older references that still provide excellent examples of fundamental ecological concepts, many of which would otherwise be lost in obscurity. As before, we have integrated woody plant physiology into multiple chapters, rather than developing it in a single chapter, and in this edition have done the same with climate because of its overriding influence on so many ecosystem processes. We have also further limited our treatment of forest ecology with examples from temperate and boreal forests, with special emphasis on North America and Europe. With a primary focus on the ecological principles of forests that form the basis for management, we have largely avoided specific treatments of forest management techniques and strategies throughout the book, although by necessity we have provided some examples of forest management in the chapters on diversity and invasive species. The revised edition would not have come to pass without the patience and dedication of many colleagues who helped in its preparation. We are especially indebted to the following for their reviews of one or more chapters: Dennis Albert, Brian Buma, Mark Dixon, Tom Dowling, Jennifer Fraterrigo, Stephen Handler, Donna Kashian, Doug Pearsall, David Rothstein, Madelyn Tucker, and Chris Webster. Other important contributions, either through provision of material, enlightening discussions, or enthusiastic encouragement, were made by Jonathon Adams, Virginia Laetz, and Dan Binkley. Victoria Meller was extremely supportive of D. M. Kashian and his time spent away from campus in the completion of this revision, as were the graduate students in the Kashian lab. Perhaps without knowing it, the graduate and undergraduate students in the 2018 and 2020 Terrestrial Ecology classes at Wayne State had an immense role in the thought processes and revisions made in the 5th edition. D. R. Zak was partially supported by the National Science Foundation and the Department of Energy during the preparation of this text. Daniel M. Kashian Detroit, Michigan, USA Donald R. Zak Ann Arbor, Michigan, USA

Forest Ecology and Landscape Ecosystems

PA R T 1

F

orest ecologists work to understand the dynamics, structure, and function of forest ecosystems. The first step in doing so is to approach ecosystems as geographic units or landscape ecosystems—­real, three-­dimensional, defined locations at the Earth’s surface. Landscape ecosystems include organisms, species, and communities, but also the air above and the soil below these living things such that organisms are but one important part of a larger landscape unit. Our perception of forest ecosystems in this book is that ecosystems are not simply extensions of forest communities, whereby ecosystems are conceived as organisms and their environment. Instead, units of whole ecosystems that integrate factors of climate, landforms, soils, co-­occurring biota, and all the interactions between them are the most appropriate units of study rather than their individual parts that too often claim our immediate attention. In understanding forest ecosystems, particularly in the context of an ecology course or in solving ecological problems, ecosystems need to be conceived at several spatial and temporal scales. It is very daunting to attempt to gain such understanding, and certain well-­known ecologists have suggested that the environmental complex is unknowable and inexpressible. However, conceiving landscapes as ecosystems and proceeding “from above,” that is, from the biggest to the smallest units to understand their spatial relationships, makes synthesis possible and manageable. Although local sites and stands with their familiar species appear a convenient starting point, we know that these are, in turn, affected by geological and climatic factors of higher levels within which they are embedded. Landscape ecosystems are nested within one another in a hierarchy of spatial sizes such that it is best to consider the big picture first—­the comprehensive view from the outside—­rather than the minutia of details as seen from inside of the ecosystem. Our perspective in understanding ecosystems involves establishing a framework for studying the components—­a framework of the spatial and temporal, hierarchical pattern of ecosystems of all sizes making up the ecosphere. Therefore, in the first part of this book, before immersing in the detail of organisms, sites, and their interrelationships, we wish to first examine landscape ecosystems, and then place them within a perspective of spatial levels and their processes at different

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

1

2  PA RT 1   Forest Ecology and Landscape Ecosystems

temporal scales. In Chapter 1, we present the basic concepts of forest ecology within a framework of landscape ecosystems. In Chapter 2, we present the concept of scale and consider the hierarchy of these landscape ecosystems. These fundamental concepts provide the basis for studying the variation, life history, and ecology of individual organisms that follow in Part 2.

Concepts of Forest Ecology

CHAPTER 1

A

forest is an ecological system dominated by trees and other woody vegetation. More than simply a stand of trees or a community of woody and herbaceous plants, a forest is a complex ecological system, or ecosystem, characterized by a layered structure of functional parts. Ecology is the study of ecological systems and their interacting abiotic and biotic components. Forest ecology, therefore, addresses the structure, composition, and function of forests. In forest ecology, we study forest organisms and their responses to physical factors of the environment across forested landscapes. Forests are widespread on land surfaces in humid climates outside of the polar regions. It is with forests in general, and with the temperate North American forest in particular, that this book is concerned. There are many ways to study forest ecosystems. Most simply, a forest may be considered in terms of the trees that give the forest its characteristic aboveground appearance or physiognomy. Thus, we think of a beech–sugar maple forest, a ponderosa pine forest, or of other forest types, for which the naming of the predominant trees alone serves to characterize the forest ecosystem. Forest types are often considered to be composed of forest stands, which are trees in a local setting possessing sufficient uniformity of species composition, age, spatial arrangement, or condition to be distinguishable from adjacent stands (Ford-­Robertson 1983). A broader concept of a forest may take into account the interrelationships that exist between forest trees and other organisms. Certain herbs and shrubs are commonly found in beech–sugar maple forests, and these may differ from those found in ponderosa pine or loblolly pine forests. Similar interrelationships may be demonstrated, for example, for birds, mammals, arthropods, mosses, fungi, and bacteria. Thus, part of the forest ecosystem is the assemblage of plants and animals living together in a biotic community. The forest community, then, is an aggregation of plants and animals living together and occupying a common area. It is thus a more organismally complex unit than the forest type. A third approach is to focus on geographic or landscape ecosystems. This approach is ­centered conceptually and in practice on whole ecosystems and not just their parts. When our primary focus is real live chunks of Earth space, that is, landscapes and waterscapes (oceans, lakes, rivers; hereafter included as parts of a landscape), we can effectively study their parts (e.g., organisms, soils, and landforms) while recognizing that each is but one part of a functioning whole. We emphasize this focus on ecosystems rather than on the individual organisms and species that are parts of them. In the past, the forest stand or the species has been the focus in natural resource fields such as forestry and wildlife. However, we are really managing whole forest ecosystems, despite their incredible complexity, because the diverse biota is inseparable from the physical environment that supports it. A consideration of the field of ecology from this viewpoint provides an overall perspective.

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

3

4  C h a p t e r 1   Concepts of Forest Ecology

ECOLOGY

Broader fields of scientific inquiry are difficult to limit and define, and ecology is one of the most indistinct. In 1866, Ernst Haeckel proposed the term oecology, from the Greek oikos meaning home or place to live, as the fourth field of biology dealing with environmental relationships of organisms. Thus, ecology literally means “the knowledge of home,” or “home wisdom.” Since its introduction, the term has been applied at one time or another to almost every aspect of scientific investigation involving the relationship of organisms to one another or to the environment (Rowe  1989). Haeckel’s organismal focus of ecology has since been redefined and expanded to include the physical aspects of the environment that provide life for those organisms (Hagen 1992; Golley 1993). Thus, Rowe (1989, p. 230) suggests: Ecology is, or should be, the study of ecological systems that are home to organisms at the surface of the earth. From this larger-­than-­life perspective, ecology’s concerns are with volumes of earth space, each consisting of an atmospheric layer lying on an earth/water layer with organisms sandwiched at the ­solar-­energized interfaces. These three-­dimensional air/organisms/earth systems are real ecosystems—­the true subjects of ecology.

This approach to ecology emphasizes whole ecosystems as well as organisms, both volumetric and having structure and function.

LANDSCAPE ECOSYSTEMS

The British botanist–ecologist Arthur Tansley (1935) introduced the term ecosystem, writing with an emphasis on “the whole ‘system,’ including not only the organism complex but also the whole complex of physical factors.” He also noted that from the point of view of the ecologist, ecosystems “are the basic units of nature on the face of the Earth.” Tansley was a biologist and vegetation ecologist, and so his idea of ecosystem was centered on organisms (species or communities) rather than geographic or landscape entities. With this bioecosystem approach, “ecosystem” derives its meaning from particular plant or animal organisms of interest, and an “abiotic” environment defined by the organisms as relevant or not is considered with lesser emphasis. In this approach, every organism defines its own ecosystem, nearly infinite in number and difficult to study and use as a basis for management and conservation. On the other hand, others (e.g., Rowe 1961a and Troll 1968, 1971) view ecosystems centered on geographic or landscape units (i.e., geoecosystems) of which organisms are but one important structural component (Rowe and Barnes 1994). We term these units landscape ecosystems in part to differentiate them from geology-­based units of study (e.g., Huggett 1995). Landscape ecosystems are geographic objects, with a defined place on the Earth. Landscape ecosystems have three dimensions (volume) just as organisms do, including landforms and biota at the Earth’s surface as well as the air above them and the soils below them (Figure 1.1). Other terms have been introduced to express the same idea, but are less commonly used, such as the ecotope (Troll 1963a, 1968) and the ecoterresa (Jenny 1980). This geographic/volumetric concept has been discussed and adapted by professional and academic ecological societies (Christensen et al. 1996), and is useful to field ecologists, naturalists, foresters and other land managers, and natural resource professionals. The concept is described in detail in Chapters 2 and 11. In addition to being geographic and volumetric, landscape ecosystems are hierarchical, extending downward from the largest ecosystem we know, the ecosphere (Cole 1958), through multiple levels of ecological organization (Figure 1.2). These levels include macrolevel units of continents and seas, each of which contains mesolevel units of regional ecosystems (major physiographic units and their included organisms), which in turn contain local ecosystems (Hills 1952), the smallest level of homogeneous environment with organisms enveloped in it. We therefore conceive the ecosphere and its landscapes as ecosystems, large and small, nested

Ecosystem boundaries Macroclimate Topoclimate

Biota

Landform Surface water Soils

Ground water

Bedrock

F I G U R E   1 . 1  The three-­dimensional, volumetric nature of a landscape ecosystem. Ecosystems comprise the atmosphere (macroclimate as well as the climate affected by surface relief), landforms and soils (underlain by ground water and bedrock), and the biota that provide a physical connection between the air and the Earth. Source: Bailey (2009) / Springer Nature.

Ecology

Organismal Biology

Cell

M

le

Organelle olecu

Cell Biology Molecular Biology and Biochemistry

F I G U R E   1 . 2  Objects of study from the most inclusive (ecosphere) to the least inclusive levels of organization (cell and organelle and molecule below it). Note that each higher level envelops the lower ones as parts of its whole. Some corresponding fields of study are also shown. Aggregates of organisms, such as populations and communities, are components of ecosystems at all scales. Like other components such as atmosphere, landform, and soil, they do not appear in this diagram of first-­order objects of study, but are shown in Figures 1.3 and 1.4. Source: Modified from Rowe (1961a).

6  C h a p t e r 1   Concepts of Forest Ecology

within one another in an ecological hierarchy, having processes at each level with their own spatial and temporal scales (see Chapter 2). Two landscape ecosystems in Figure  1.3 (Rowe  1984b) illustrate characteristic ecosystem ­differences in hilly or mountainous terrain. The two ecosystems are distinguished by different geomorphologies (convex upper slope versus concave lower slope, and high versus low topographic slope position) that mediate microclimate, soil water, and nutrient availability. The vertical dashed line is placed at an ecologically significant boundary that spatially separates the two ecosystems. Organisms in these ecosystems are sandwiched between the air and the Earth. Also shown in Figure 1.3 are the traditional fields of study in which individuals seek to understand each of the ecosystem components, although forest ecologists aim to understand the integrated effects of all of these components. In many parts of this book, we focus on organisms, species, and communities, but always remembering that they are parts of volumetric, hierarchical ecosystems. For studies of organisms in their immediate surroundings, a biological approach is often useful. Nevertheless, for management of ecosystems, studies of forest productivity, and the conservation and restoration of forest ecosystems, a landscape ecosystem approach is eminently practical and theoretically sound. Forest ecologists not only study (i) organisms of these systems and their aggregates as communities

HILLSIDE SLOPE Upper

Lower

Fields of Study Upper Air Layer

Meteorology Climatology

Lower Air Layer

Topoclimatology

Biota

Ecology Biology

Soil

Pedology

Landform

Geomorphology & Surficial Geology

Mineral Strata

Bedrock Geology

Water Table

Hydrology

F I G U R E   1 . 3  Structural profile illustrating landscape ecosystems of upper and lower slopes. Air–earth layers surround the organisms at the Earth’s energized surface. The vertical line is set at an ecologically significant topographic break, dividing the upper and lower slope ecosystems. Source: Rowe (1984b) / United States Department of Agriculture / Public Domain.

Landscape Ecosystems  7

and populations (see Chapters 3–5, 12, and 13), but also (ii) the functioning of local ecosystems that involves complex interactions among organisms and their supporting environment (Chapters 6–11, 13–14, and 16–19), and (iii) the spatial patterns of occurrence and interrelationships of entire forest ecosystems (Chapters 2, 18, 19, and 22).

LANDSCAPE ECOSYSTEM AND COMMUNITY The term ecosystem was introduced in 1935 by Tansley, but terms in various languages, such as taiga, heath, bald, Auenwald (river floodplain forest), pampas, prairie, chaparral, maquis, hammock, muskeg, and bog, have long been used to depict interactions among air, organisms, and soil. Very often the terms emphasize a distinctive plant community and may therefore imply that an ecosystem is simply an extension of a community. Tansley’s definition of ecosystem as “organism + environment” leads to this view, and general definitions such as “ecosystem = biotic community + environment” reinforce this view. Although indispensable for forest ecologists and managers, vegetation may not always coincide with climate–landform–soil-­based ecosystems because of disturbance and/or unknown historical factors. Thus, we emphasize the importance of geography and physiography within a regional climatic setting as the basis of understanding not only vegetation but whole ecosystem structure and function. This conceptual approach is not to say that communities are unimportant; they form the key ecosystem component whose response is essential in affecting ecosystem change and indicating the integrated effects of many site factors (Chapter 11). Communities and populations, however, fundamentally differ from entities such as ecosystems, organisms, and cells because they are aggregates of individuals but not functional systems. Entities such as ecosystem, organism, cell, and molecule (Figure  1.2) are “volumetric” levels of organization because they have structurally joined parts that form a functioning unit (Rowe 1961a, 1992c). By contrast, communities and populations are assemblages or aggregates of spatially separated trees and understory plants that have no necessary, physical, structural connections.

ECOSYSTEM STRUCTURE AND FUNCTION An ecosystem’s structure describes the spatial arrangement of its parts. In local forest ecosystems, multiple strata of atmosphere and soil (Figure 1.3), as well as surface relief, influence species composition and its patterns of occurrence. Plants provide a physical connection between the soil strata and the aboveground layers of air. The layered structure and related physical and chemical properties of soil are well studied and understood, but similar properties of the atmosphere are not (Rowe 1961a; Woodward 1987), particularly as they affect forest organisms at the Earth’s surface. Vegetation itself is also structured vertically, and the vertical structure and species composition of communities as well as species interactions provide insights critical in understanding properties and processes of ecosystems. In addition to species composition, vegetation structure is shaped by tree form and stem density, canopy characteristics, shrub and herbaceous plant abundance, and the amount and distribution of dead vegetation, among other characteristics. Thus, the physiognomy of vegetation varies markedly in different regional and local ecosystem types (Chapters 5 and 6). The horizontal spatial patterning of these structural components occurs at multiple spatial scales, but is most often described across broad scales. Landscapes are structured horizontally into mosaics of ecosystems that reflect differences in climate, geology, and physiography and their relative effects on vegetation. The natural communities of these systems reflect limiting factors of climate, soil water, nutrients, and disturbances, and in turn modify the physical factors. Different ecosystem mosaics characterize mountains, plains, and river valleys due to fundamental differences in their physical factors and the vegetation adapted thereto (Chapter  8). The diversity of landscape ecosystems can be assessed by understanding such mosaics (Lapin and Barnes 1995). Such an understanding provides a spatial ecosystem framework for programs in biodiversity that

8  C h a p t e r 1   Concepts of Forest Ecology

seek to conserve and manage the diversity of organisms and maintain and increase populations of rare and endangered species (Chapters 14 and 22). Local landscape ecosystems, besides having a structure of interconnected parts, are functional units characterized by many processes that define their properties. Ecosystem-­level processes are part of the entire system and are not restricted only to physical or biotic parts. These processes drive or mediate the flow of energy or matter and/or the cycling of materials in the system. Organic matter decomposition; cycling of water, nutrients, and carbon; and biomass accumulation are considered ecosystem-­level processes. Other processes are often more associated with plant species, populations, or communities, such as photosynthesis and respiration, reproduction, regeneration, mortality, and succession. Despite their basis in organisms, these are also ecosystem processes because they are mediated and regulated by characteristic ecosystem factors of temperature, water, nutrients, and disturbances (such as fire, windstorm, and flooding). Ecosystem function is often described with box-­and-­arrow diagrams, flow charts, and simulation modeling as ways of disentangling very complex systems (Figure 1.4). Landscape ecosystem structure and function are tightly coupled by the physical environment, the frequency and severity of disturbances that reset succession, and the life histories of the plants and animals that comprise the biotic community. There is increasing evidence that many aspects of an ecosystem’s function are linked to the diversity of its biota (Chapter 14). In turn, an

ABIOTIC ENVIROMMENT SUBSTANCES water, O2, CO2, minerals

OTHER RADIATION SPACE STRUCTURE pressure, heat, media (soil, water, fire, light air) and their motion, pH, etc. area, substratum

HUMANS as a superorganic factor

PRIMARY PRODUCERS (other green plants autotrophs*) pollination, distribution, selection by grazing

symbionts

DECOMPOSERS

DEAD organic substances

mineralizers H

CONSUMERS herbivores

scavengers

H

parasites

bacteriophages

soil humus

dead

living H

predators

OUTPUTS by respiration, erosion, oozing away of water are omitted

‘limit’ of the ecosystem

material or energy flow

Humans as a partner of the ecosystem

H

OTHER ECOSYSTEMS

other influences

*relations omitted

F I G U R E   1 . 4  A model of a landscape ecosystem detailing the flow of energy and matter among biota. The model also describes the interactions between the physical environment and the biota. Source: Reprinted from Ellenberg (1988) / Cambridge University Press.

Examples of Landscape Ecosystems  9

ecosystem’s biodiversity is strongly shaped by its physical factors such as climate, physiography, and soil, which provide the context within which organisms survive, adapt, and evolve. In addition, the functional aspects of an ecosystem are strongly affected by its geographical context and by its spatial position relative to its surrounding neighbors on a landscape. Adjacent ecosystems, especially in mountainous or hilly terrain, affect one another by the lateral exchanges of materials and energy. Water and snow, soil, organic matter, nutrients, and seeds are transported downhill; the effects on other ecosystems depend on the size, shape, and composition of the systems. Therefore, an understanding of the spatial pattern and configuration of landscape ecosystems can play an important role in the management of ecosystems and their biota. An excellent overview of function in terrestrial ecosystems is given by Chapin et al. (2012), its variation across landscapes considered in Lovett et al. (2005), and the history of the ecosystem concept itself is reviewed by Golley (1993).

EXAMPLES OF LANDSCAPE ECOSYSTEMS

There are many examples of landscape ecosystems, some of which are easily discerned in the field, whereas others must be carved out of geographic continua. For example, bogs in the glaciated terrain of northern and boreal regions are easily discernible ecosystem types. They exhibit distinctive physiography, soil, and vegetation, and they recur in the landscape. The distinct terraces of river floodplains distinguish local ecosystem types that differ in microclimate, soil, drainage, vegetation, and their dynamics. Mountains are characterized by gradients of elevation, aspect, and slope steepness along which strikingly different ecosystems can be purposefully delimited and mapped, although their boundaries may not be sharply demarcated. On old lake-­bed terrain, dry, sandy beach ridges are easily distinguished from adjacent depressions of swampy land. Differences in water drainage and oxygen availability in these adjacent systems result in major differences in their native tree-­and ground-­flora vegetation. Unseen but critical are the very different processes that also distinguish these ecosystems. Ecosystem components will change over time, giving rise to a sequence of different landscape or waterscape ecosystems on a given place on the Earth’s surface. Striking changes in the fossil record illustrate long-­term changes in physiography, soil, and biota (Chapter 15), but short-­term changes also occur. Shorter-­term changes are most obvious when natural or human-­caused disturbances affect vegetation. A site-­specific location may exhibit a range of different forest communities over 100 to 300 years, young and old, as disturbances or lack thereof affect the biota. Ecosystem structure and function at this site change over time. The landforms and associated pattern of parent material, soil, and climate also change over time, but more slowly than the suite of species available to recolonize the disturbed site. Thus, what we term the ecosystem site type or simply ecosystem type (land area supporting potentially equivalent ecosystems) can be distinguished and mapped regardless of the forest community currently present. One such example is illustrated in Figure 1.5 where deciduous forests occur on glaciated terrain in southern Michigan. In this setting, three local ecosystem types are distinguished by differences in physiography (outwash and moraine landforms); soil; drainage; and overstory, understory, and ground-­cover vegetation. The relatively fine-­textured, silty soil on the outwash plain (type 1) supports forest dominated by white oak, whereas the drier, coarse-­textured, sandy outwash soil (type 2) supports a black oak community. The fine-­textured, moist, clayey soil on the rolling moraine landform (type 3) supports a community dominated by northern red oak. The moraine formerly supported a beech–sugar maple forest. However, recurrent fires through the drier white and black oak ecosystems killed the fire-­sensitive mesic species of the adjacent moraine ­ecosystem and led to dominance of northern red oak. An occasional beech is still found, and red maple has invaded the shaded understory. The western portion of ecosystem type 1 was clear-­cut about 90 years ago, and a cover type markedly different from the adjacent old-­growth white oak forest has formed (Figure 1.5, type

10  C h a p t e r 1   Concepts of Forest Ecology

(a)

Ecosystem Type 1

T

3

2

1a-cut

1b-uncut

Black Oak

Northern Red Oak

American beech

T'

White Oak

Fine Outwash – Silt Coarse

(b)

and

sh – S

Outwa

End Moraine (Till)

Forest Cover Type 1a cut

1b uncut

2 Black Oak

3 Northern Red Oak

T

T'

1

2

3

Ecosystem Type Cutting Boundary

Type Boundary

Transect T-T'

F I G U R E   1 . 5  Three local landscape ecosystem types in glacial terrain that are differentiated by landform, parent material, soil, and vegetation: (a) Lateral transect from T to T′ showing vegetation and underlying geological parent material. (b) Top view showing distribution of forest trees, cover types, and ecosystem types. Following clear-­cutting of part of type 1, two forest cover types (1a—­early successional oak forest; 1b—­old-­growth white-­oak forest) are distinguished. The diagram illustrates that different forest cover types (1a and 1b) are not necessarily different ecosystem types. They represent two of many possible ecosystem derivatives (disturbed in 1a versus relatively undisturbed in 1b) of a given ecosystem type. See text for discussion.

An Approach to the Study of Forest Ecology  11

la). The overstory tree layer of the disturbed area is now dominated by white, black, and red oaks, white ash, black maple, American elm, black cherry, and sassafras. These species either sprouted from the base of cut trees, were already present in the white oak forest understory, or seeded in from adjacent communities following cutting. Today, we can recognize two (types 1a and 1b) of the many compositionally different forest communities that might occur at the site of ecosystem type 1; these easily could be mapped as two different forest cover types. Because the cut-­over area (type 1a) has physiography and soil like type 1b, its vegetation may gradually become similar in structure to those of ecosystem type 1b, providing that no major changes in climate, soil, or ­ecosystem processes (including changed browsing pressure by herbivores) were caused by clear-­ cutting. Thus, a given local ecosystem type, defined by relatively stable features of physiography and soil, may have a suite of disturbance-­induced cover types (Simpson et al. 1990). Therefore, recognizing forest cover types alone is not necessarily likely to provide a useful estimate of site potential for management or conservation. However, cover types are extremely useful in management planning for wildlife, timber, water, and recreational use of existing forest communities. It is often the conspicuous forest cover type that receives our immediate attention. However, the enormous complexity of geographic space and changes in its component ecosystems through time require that major attention be directed to atmospheric, geologic, physiographic, and soil properties of forest landscapes. In summary, for every landscape, a combination of factors should be used to distinguish the pattern of local and regional landscape and waterscape ecosystems that have similar ecological potential in the long run. Understanding ecological units at multiple spatial scales (Chapters 2, 11, and 22) is needed to provide the basis for monitoring ecosystem change over time and for ecosystem management.

AN APPROACH TO THE STUDY OF FOREST ECOLOGY

This book presents the scope of forest ecology as having at least two major parts. The first is analysis: studying the details of ecosystem components (atmosphere, organisms, and Earth). The second part is synthesis: understanding the broad ecological framework within which ecosystem components are integrated. We therefore emphasize a “big picture” approach even as we discuss much smaller details of the organic and inorganic parts of ecosystems, their interactions, and the structure and function of entire ecosystems. The explicit separation and analysis of the biotic and physical parts of ecosystems are not realistic or desirable from an ecological perspective, but may serve as reasonable subdivisions to provide both analysis and synthesis. The sequence of presentation is (i) understanding forest ecosystems at multiple spatial scales, (ii) the variation, life ­history, and structure of the forest tree, (iii) the physical environment of forests, (iv) forest communities, (v) forest ecosystem dynamics, and (vi) the outlook for and consideration of future forests. In Part 1 of this book, landscape ecosystems are the primary objects of interest, but we observe the importance of understanding spatial scale in examining them. Spatial scale influences the observations we make about ecosystems because ecological processes vary as spatial scale changes. Advances in hierarchy theory in ecology have helped us to understand ecosystems as volumetric segments of the ecosphere within which smaller volumetric units are nested. In Part 2, we consider that the forest tree largely owes its appearance, rate of growth, and size to the environment in which it has grown throughout its life. In other words, the phenotype, the individual as it appears in the forest, is the product of the environment acting on its genotype, its individual genetic constitution. An understanding of trees as parts of ecosystems depends, in part, upon the genetics of the trees themselves and the way their genetic heritage affects their response to the environment they live in. In addition, life history features of regeneration as well as anatomical and physiological aspects of forest trees in relation to environment are considered.

12  C h a p t e r 1   Concepts of Forest Ecology

We examine the physical factors of forest ecosystems in Part 3. The site is the sum total of environmental factors surrounding and available to the plant at a specific geographic place. These factors are primarily the atmospheric, physiographic, and soil components of the physical environment, but also include important influences of growing vegetation which itself affects the microclimate and soil. Climate includes various atmospheric factors that vary across hours, days, months, and years, such as solar radiation, air temperature, precipitation, humidity, wind, and carbon dioxide content. We provide special consideration to sunlight (solar radiation) and temperature, and we examine the relations and dynamics of water throughout the text in appropriate contexts. Physiological processes related to photosynthesis, respiration, and growth are included in Part 3 in chapters on light and temperature, but also in later chapters on carbon balance and nutrient cycling. Physiography—­comprising landforms and soil parent material—­—­plays a key role in affecting climate as well as the amount and rates at which radiation and moisture are received and distributed in forest ecosystems. Below the ground surface, the supply of soil moisture and nutrients, the physical structure of the soil, microbial communities in the soil, and the nature and decomposition pattern of organic matter, that is, factors pertaining to edaphic characteristics of soil, all affect the growth and development of plants. We also treat fire as a site factor. The study of the environmental factors and their effects on individual plants constitutes the field of autecology, and we summarize those aspects of forest autecology most pertinent to an understanding of forest ecosystems. Finally, we examine site quality and its evaluation, focusing on the degree to which individual site factors and combinations of them are used to estimate the productivity of forest ecosystems and determine management prescriptions for them. In Part 4, we consider forest communities and their plants and animals as integral parts of ecosystems. First, we examine the important roles of animals in affecting all phases of plant development, as well as their effects on forest communities and ecosystem processes. Forest communities are then treated, emphasizing their composition, occurrence, and the interactions (with emphasis on competition and mutualism) of their constituent individuals. Finally, we examine concepts of biological and ecological diversity, including their importance, measurement, and conservation. Forest ecosystems are ever changing, and we examine their dynamics in Part 5. We first consider change in ecosystems over hundreds to millions of years. Primary bases for this change come from geologic and physiographic studies of the land itself and determination of plant species and their communities as deduced from the paleoecological record. Disturbance and succession as ecosystem processes are treated next, followed by chapters on whole ecosystem functioning, including the carbon balance of trees and ecosystems and the dynamics of nutrients. Part 6 examines future forest ecosystems in considering their sustainability. We first examine the effects of climate change on forest ecosystems, including effects on individuals, populations, communities, and whole ecosystems. We also explore the potential for mitigation of climate change using forest management. We next consider the importance of invasive plants and animals for forest ecosystem structure and function. Finally, we place humans in an appropriate context of forest ecosystems in considering forest landscape ecology and the emerging concept of forest sustainability.

APPLICABILITY TO FOREST MANAGEMENT

The management of forests is, as it should be, a continuously evolving effort as new science, technology, and consideration of the role of humans in forest ecosystems develop. Management of forests may include decisions made at broad or local scales, but consideration is increasingly given to whole systems, their complexity and heterogeneity, as well as their long-­term sustainability and integrity (Franklin et al. 2018) rather than single-­focus use of their parts for recreation, wildlife, or

Applicability to Forest Management  13

timber. To ecologists, understanding ecosystems for their own sake has significant value in itself, but the principles of forest ecology are essential for their guidance in conservation and management practices. Spurr (1945) defined silviculture, or applied forest ecology, as the theory and practice of controlling forest establishment, composition, and growth. Modern texts suggest that current silviculture is most concerned with managing forests for future products and services in a manner that has social stability (Ashton and Kelty 2018). Nevertheless, we explicitly emphasize that silviculture is the theory and practice of controlling forest ecosystem composition, structure, function, and heterogeneity, often at multiple spatial and temporal scales. Even local management operations may have significant and often long-­term effects on surrounding landscapes far removed from the specific site of human activity. The management, conservation, and restoration of forested landscape ecosystems rely heavily on understanding the structure and function of forest ecosystems and the autecology of their organisms (i.e., forest ecology). More than ever, understanding how ecological systems work is critical for making sound decisions regarding human intervention in forests. Understanding the structure and function of whole ecosystems will also equip us to best deal with the ecological consequences of past human activities, such as deforestation and its wide range of impacts, habitat loss, and fragmentation. Human activities associated with industrialization, such as air pollution or the introduction of destructive new species, have resulted in local tree mortality and, in certain ecosystems, widespread death and decline of forest trees and other organisms. Increasing the carbon dioxide concentration in the atmosphere due to the burning of fossil fuels and deforestation are creating climatic changes and concomitant changes in forest ecosystems that are quickly increasing in their severity. Invasive plants and animals introduced into forests are well known for disrupting ecosystem structure and function on relatively short time scales. In all cases, one must understand the system in order to understand the extent to which it is disrupted. At the time of the last edition of this textbook over 20 years ago, a new paradigm in forestry had occurred that shifted the view of the forest from one of single commodities (timber, wildlife, water, and recreation) to forestry understood as sustaining ecosystems and their structure, diversity, heterogeneity, and function (Rowe 1994; Kohm and Franklin 1997). That paradigm emphasized maintaining the integrity of ecosystems across landscapes by sustaining their natural patterns and processes even when heavily manipulated by humans. This paradigm has been further refined over the past two decades to include an emphasis on considering or even incorporating natural disturbances, increasing ecosystem resistance and resilience, embracing structural complexity, and maintaining a range of ecosystem conditions across broad scales (Franklin et al. 2018). Whether intentionally or otherwise, this paradigm has the landscape ecosystem approach at its core because it focuses on the maintenance of landscapes as complete ecosystems. The only way to assure the sustained yields of forests, wildlife, and water, now and in the future, is to maintain the integrity of their processes and keep their biota in a healthy state. Such a practice demands an intimate familiarity with forest ecology as we have described it.

SUGGESTED    R E A D I N G S Bailey, R.G. (2009). Ecosystem Geography, 2e. (Chapters 1 and 11). New York: Springer-­Verlag 251 pp. + 1 map.

Golley, F.B. (1993). A History of the Ecosystem Concept in Ecology. New Haven, CT: Yale Univ. Press 254 pp.

Chapin, F.S. III, Matson, P.A., Vitousek, P., and Chapin, M.C. (2012). Principles of Terrestrial Ecosystem Ecology, 2e. New York: Springer 520 pp.

Franklin, J.F., Johnson, K.N., and Johnson, D.L. (2018). Ecological Forest Management. (Chapter  1). Long Grove, IL: Waveland Press 646 pp.

14  C h a p t e r 1   Concepts of Forest Ecology Kohm, K.A. and Franklin, J.F. (ed.) (1997). Creating a Forestry for the 21st Century. Washington, D.C: Island Press 475 pp. Rowe, J.S. (1961). The level-­of-­integration concept and ecology. Ecology 42: 420–427. Rowe, J.S. (1992). The ecosystem approach to forestland management. For. Chron. 68: 222–224.

Rowe, J.S. (1994). A new paradigm for forestry. For. Chron. 70: 565–568. Tansley, A.G. (1935). The use and abuse of vegetational concepts and terms. Ecology 16: 284–307.

Landscape Ecosystems at Multiple Scales

CHAPTER 2

F

orest ecologists often by default are also landscape ecologists—­people whose interests and ­practices necessarily encompass ecosystems of many different spatial scales. A landscape is a heterogeneous land area composed of a group of intermeshed ecosystems, each with interacting atmosphere, earth, and biota. Landscape ecology is the study of the spatial patterns of ­ecosystems, the causes of the pattern, and how the pattern affects ecological processes. Many of the core concepts of landscape ecosystems overlap with those of landscape ecology, but the domain of interest of the two fields is not identical. In this chapter, we examine the spatial occurrence and complexity of landscape ecosystems, whereas the discipline of landscape ecology is considered in Chapter 22.

OVERVIEW OF SPATIAL AND TEMPORAL SCALES

Forest ecologists ask questions, conduct field experiments, and resolve problems at different scales of space and time, and these scales directly influence the problems posed, the questions asked, and the techniques employed. Scale describes the dimensions of an object or process in space or time (Turner and Gardner 2015). Scale is often confused with biological or ecological levels of organization, such as cells, tissues, organs, organisms, and ecosystems (see Figure 1.2, Chapter 1), but the concepts are quite distinct. For example, organisms are obviously an important level of organization, but they may occur over very different spatial scales (e.g., 1–10 μm for bacteria to 84 m tall giant sequoia) with very different life spans (e.g., 24 hours for mayflies to nearly 5000 years for bristlecone pine). A given scale is defined by its grain (its finest spatial or temporal resolution) and its extent (the size of a study area or the length of time considered), and there is some trade-­off between the two characteristics. For example, forest ecologists might increase grain size (e.g., sample plot size) in a larger study area (e.g., the size of a county) because of their interest in ­capturing variation in species composition at a larger spatial scale. However, they might reduce grain size (using smaller plots or quadrats) if they are interested in describing species variation across a 1 ha woodlot. As a result, no single set of spatial and temporal scales is applicable to all the questions addressed by forest ecologists (Turner and Gardner 2015). Processes that occur at a given scale do so within a hierarchy. The process in question is provided context (influenced and constrained) by the processes at the hierarchical level above it, while the level below it provides its mechanisms. For example, it is vital in understanding the growth of trees in forested wetlands to examine those processes that occur at broader and finer spatial scales than those of an individual tree. Wetland tree growth is constrained by broader-­scale processes such as landscape position and soils, but the mechanisms that influence tree growth, such as tree physiology, occur at finer scales. A classical hierarchical model shown in Figures 2.1 and 2.2 suggests that many forest managers, ecologists, and naturalists would have an interest at a microscale (up to 100 ha and 500 years), which would encompass the size and life span of a typical forest in North America (Delcourt and Delcourt 1988). Disturbance events such as clear-­cuts, some wildfires, and wind events, as well as biotic responses such as gap-­phase replacement, competition, Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

15

16  C h a p t e r 2   Landscape Ecosystems at Multiple Scales F I G U R E 2.1  General spatial–temporal scales for the hierarchical characterization of disturbance regimes, ecosystem and biotic responses, and vegetation units. Quaternary studies at the macroscale refer to investigations of species migrations and extinctions in the time frame of 10 000–1 000 000 years in the Quaternary Period (see Chapter 15) on a continental scale. Source: After Delcourt and Delcourt (1988) / Springer Nature.

SCALE PARADIGM

TEMPORAL SCALE (yr)

109

megascale

106 macroscale (QUATERNARY STUDIES)

103

mesoscale (LANDSCAPE ECOLOGY)

microscale 100

100

104

108

SPATIAL SCALE

1012 (m2)

and productivity might occur at this spatiotemporal scale (Figure 2.2). Many landscape ecologists are concerned with the mesoscale (100–10 000 ha and 500–100 000 years), which encompasses longer-­­­term processes such as fire regimes, insect or pathogen outbreaks, and soil development, as well as biotic processes such as secondary succession. Seasonal patterns of precipitation and ­temperature and longer-­term (decades to centuries) weather trends and climatic fluctuations affect the distribution, growth, and interactions of plants and animals at microscale and mesoscale. Macroscale and megascale (Figures 2.1 and 2.2) extend in space and time to include not only events of ecosystem change and plant migration occurring at the level of physiographic regions (macroscale), but also plate tectonics and evolution of biota on the spatial scale of continents and the ecosphere (megascale). The disturbance regimes, ecological responses, and vegetation units that are of concern to forest ecologists may occur over a range of spatiotemporal scales rather than cleanly within one shown in the hierarchical framework in Figure 2.2. Disturbance events such as wildfire or insect and pathogen outbreaks may occur at a local site or over thousands of square kilometers. Wind may uproot a single tree in an old-­growth forest or may affect huge land areas for much longer intervals, such as the 150 000 ha (370 000 acres) of forest blown down in the Boundary Waters Canoe Area in northern Minnesota in 1999 (Mattson and Shriner  2001). Likewise, geomorphic processes such as soil creep, debris avalanches, stream transport, and deposition of sediments affect vegetation at levels from individual plants to large forest stands and occur on areas extending from the size of sample plots (m2) and stream watersheds (ha) to mountain ranges (km2). Vegetation change or succession is another good example of an ecosystem process that may be examined at different spatial and temporal scales. For example, vegetation change may occur in just a few years in small tree-­fall gaps, as widespread secondary succession over decades or centuries, or as evolution over millennia on a continental spatial scale. Even the same disturbance type may occur at different scales among ecosystems. For example, subalpine lodgepole pine forests of Yellowstone National Park were historically burned by large fires (>5000 ha) every 100–300 years (Romme and Despain  1989), which spans both the microscale and mesoscale in Figure  2.1. By contrast, hemlock-­­­­northern hardwood forests of the Lake States were likely burned by very small fires

Spatial Scales of Hierarchical Landscape Ecosystems  17 ENVIRONMENTAL DISTURBANCE REGIMES 100

104

108

1012

ECOSYSTEM AND BIOTIC RESPONSES 100

104

108

1012

VEGETATIONAL PATTERNS 100

104

Temporal Scale (years)

108

1012 108

Global Terrestrial Vegetation Plate tectonics

106

Evolution of the Biota

FM. Zone

108

Ecosystem Change Landforms Biota Speciation Soils Macro-andMigration Microclimate

Glacial– Interglacial Climatic Cycles Soil Development

Formation

Climatic Fluctuations

103

Fire Regime

104

*Examples: Wildfire, Wind Damage, clear-cut, Flood, Earthquake

108

Tree

1012

100

104

108

Type Subtype

Competition Productivity

Pathogen Outbreak *Disturbance Events

100

Stand

Secondary Succession Gap-phase Replacement

Human Activities

100

108

Shrub Herb

108

1012

100

104

108

108 1012

Spatial Scale (m2)

F I G U R E 2 . 2   Disturbance regimes, ecosystem and biotic responses, and vegetational units viewed in the context of four spatial–temporal scales shown in Figure 2.1. Source: Modified from Delcourt and Delcourt, 1988 / with permission of Elsevier.

(50 000 000

>62 500 km2

Ecoprovince

1:

10 000 000–50 000 000

2500–62 500 km2

Domain Division Province

Ecoregion

1:

2 000 000–10 000,000

100–2500 km

Section

Ecodistrict

1:

500 000–2 000 000

625–10 000 ha

Subsection

Ecosection

1:

100 000–500 000

25–625 ha

Landtype association

Ecoseries

1:

25 000–100 000

1.5–25 ha

Landtype and landtype phase

Ecotype

1:

5000–25 000

0.25–1.5 ha

Landtype phase

Eco-­element

1:

430 km h−1 were reported for the 1993  Kane tornado in northwestern Pennsylvania (Peterson and Pickett  1995), although only about 3% of tornadoes have winds that exceed 320 km h−1. Straight-­line winds in the form of downbursts during severe thunderstorms are common in the Lake States (Canham and Loucks  1984). A single thunderstorm may produce multiple downbursts as it moves across the landscape; this family of downbursts is known as a derecho. A derecho that occurred in northern Wisconsin in 1977 included 26 separate downbursts and created a damage path 20-­km wide and 200-­km long across the forested landscape (Frelich 2002). As opposed to wind events that affect large portions of a landscape, local wind gusts (90–­150 km h−1), associated with cyclonic storms that break off or uproot one or more trees, may affect only small areas (20–­100 m2) in the forest. These canopy gaps or exposed patches on the ground occur over vast areas in North America, and their accumulated ecological significance may be equal to or greater than that caused by large wind events (Bormann and Likens 1979). Notably, winds capable of causing forest disturbances occur over a broad range of velocities, and they encounter regional and local ecosystems with many different site conditions and vegetation at different times of the year. Despite the many gradients, the nature and amount of wind disturbance are somewhat predictable by regional ecosystem, specific landform–­soil drainage conditions, site exposure, and tree species and height. Principles of Wind Damage  Wind is turbulent near the Earth’s surface due to the nature of the ground over which the air is moving, with higher turbulence along rougher surfaces. Forests form a particularly rough surface such that wind over them is much more turbulent than over a field. Turbulence is organized into coherent gusts that move widely across the forest (Figure 16.8). Wind becomes unstable as it moves across a forest, and each gust consists of a rapid increase in wind speed and a downward movement of air into the canopy (see #5 in Figure 16.8). The strongest gusts exert a force on trees up to 10 times larger than does the mean wind velocity (Quine et al. 1995), making them the most consistent cause of windthrow in forests. Changes in vegetation height or composition induce additional turbulence and wind acceleration. Forest edges, roads, and other open areas that cause wind to deflect upward tend to increase damage to trees. In addition, the force of the wind on trees nearer to a forest edge is substantially greater than that on interior trees, and trees that grow on edges have significantly higher physical resistance to wind. By contrast, trees suddenly exposed by the formation of a new edge or the opening of a patch lack such a physical resistance and are more vulnerable to wind damage than those that grow on a forest edge. The edges between cutover and uncut areas tend to deflect wind currents, particularly at high elevations in mountainous areas, producing increased wind velocities where the deflected currents join together (Figure 16.9). Wind damage in mountainous forested country is often due to local acceleration of wind by either topography or forest edges (Gratkowski 1956; Alexander 1964).

402  C h a p t e r 1 6  Disturbance

z u(z)

4. Streamwise vortices

y

h x 5. Break up and production of smaller-scale turbulence

3. Instability 2. Transverse vortex

1. Kelvin–Helmholtz waves

F I G U R E  1 6 .8   Idealized formation of coherent gusts over a forest. (1) The rapid change of wind speed (u) at the top of the canopy (z = h) is unstable and leads to the emergence of Kelvin–­Helmholtz waves. (2) The waves become transformed into across-­wind vortices. (3) These vortices are unstable and begin to distort. (4) The distortion produces coherent gusts aligned in the direction of the wind. The gusts propagate across the forest and if strong enough, lead to wind damage. (5) Eventually the gusts become distorted and break up. Source: Finnigan and Brunet (1995) / Cambridge University Press. Reproduced with permission of the British Forestry Commission.

N Uncut

Cutover

Unaccelerated winds Accelerated winds

WIND

Turbulent area Uncut

F I G U R E  1 6 .9   Windthrow of standing trees following a harvest cut. A change in the direction of the cutting boundary that was parallel to the prevailing winds acts to funnel wind into standing timber, causing a pocket of blowdown. Routt National Forest, northwestern Colorado. Source: Alexander (1964) / with permission of Oxford University Press. Reprinted with the permission of the Society of American Foresters.

Major Disturbances in Forest Ecosystems  403

Topography has a strong influence on wind damage to forests. Wind v­elocity is often assumed to be at its highest near the tops of slopes and ridges, but this is the case only when the wind direction is perpendicular to the slope. When the wind direction is at a sharp angle to the slope, the wind velocity is highest at the midslope position, and at the valley floor when the wind is parallel to the slope (Frelich 2002). Local wind acceleration can also take place on gradual and smooth lee slopes during severe windstorms, in gaps and saddles of main ridges, and in narrow valleys or V-­ shaped openings in the forest that constrict the wind flow. Forests subject to hurricanes or other windstorms where wind direction is lateral and in the same direction often experience damage related to slope aspect (Chapter 22; Boose et al. 1994). Where damaging winds result from vertical downbursts, however, there may be little correspondence between wind damage and topography (Frelich and Lorimer 1991). Windthrow (uprooting of trees) and snapping is most likely where air currents are concentrated to form high wind velocities at a particular spot. Windthrow most often occurs in shallow-­ rooted species on shallow or poorly drained soils. Soil provides tree stability, and thus a tree whose root system is in contact with more soil mass is less likely to be uprooted by wind. Rooting depth, rather than lateral root spread, is strongly determined by soil conditions. Rooting depth is most often restricted by inadequate oxygen supply (poorly drained soils) or impenetrable layers (bedrock, ortstein, etc.). Tall trees with large crowns are especially susceptible to wind damage, and conifers are more vulnerable than deciduous trees in winter, when deciduous trees are leafless. It is possible to identify species at high risk for windthrow—­black spruce, white pine, and Norway spruce, for example—­and high-­risk sites.

Broadscale Disturbance by Hurricanes  Occasional severe storms, particularly h­urricanes, may destroy the overstory canopy across hundreds or thousands of hectares of forest, initiating secondary forest succession over large areas. Gulf and Southern Atlantic Coasts  More than 50 hurricanes have made landfall along the south-

eastern Louisiana coast since 1851 (Blake and Gibney 2011). Hurricane Katrina made landfall in southeast Louisiana and the Mississippi coastline on August 29, 2005, with hurricane-­force winds that extended at least 100 km inland. It was at the time the third most intense storm on record to make landfall in the United States, killing or severely damaging 320 million trees over 84 000 km2 (Chambers et al. 2007). Wind damage to forests in the region was spatially variable, however, with 60% severely disturbed, 35% moderately disturbed, and 7% only lightly disturbed in the Lower Pearl River Valley and its surrounding area (Wang and Xu 2009). Overall, post-­hurricane damage patterns were closely related to stand conditions and site characteristics (Chapman et  al.  2008; Kupfer et al. 2008). Forests closest to rivers and streams in the region were more severely disturbed than areas further away from bottomlands, related to the increased susceptibility of trees to wind on poorly drained sites. In addition, older stands with a higher percentage of hardwoods were more susceptible to wind damage. In general, wetland forests were more susceptible to damage than other forest types with the exception of those having high proportions of cypress and/or tupelo, which were resistant to wind damage in wetlands because of their buttressed roots. Oaks and sweetgum were more susceptible to strong winds because of their shallow rooting. On uplands, evergreen pines (especially longleaf pine) were more resistant to wind damage compared to oaks. Shorter trees with smaller diameters in the shrub/scrub ecosystems of the region were the most resistant to wind damage. Hurricane Andrew struck on August 24, 1992, in south Florida and severely affected ecosystems over an area 50-­km wide by 100-­km long (Loope et al. 1994; Pimm et al. 1994). From 25–­40% of overstory trees in pinelands were uprooted or snapped at heights of 1–­6 m; 2–­3 times as many pines were broken as were uprooted. The understory was relatively unaffected. Tall, large-­diameter hardwood trees of hammock ecosystems (small islands of trees within broad wetland areas) were

404  C h a p t e r 1 6  Disturbance

extensively damaged (20–­30% downed or large branches broken off), but forests dominated by bald cypress showed only modest damage. Mangroves suffered heavy damage, reaching 80–­95% trunk snapping and uprooting in the most vulnerable sites (Smith et al. 1994). Interestingly, small patches of surviving mangroves were saplings that had previously regenerated in gaps caused by lightning-­induced fires. These small gaps provided nuclei for recolonization of destroyed forests, given that the Florida mangroves only reproduce viable propagules on stems less than 1  m in height. A primary concern during post-­hurricane succession is the acceleration of introduced plant species spread that was already occurring before Hurricane Andrew. Propagule dispersal and facilitation of establishment and further spread in new canopy gaps of exotic tree species, such as the Chinese tallow, have proven to be problematic over the two decades since the event occurred (Conner et al. 2002; Henkel et al. 2016; Smith et al. 2020).

New England—­1938 Hurricane  Hurricanes also regularly and historically affect forests of the northern Atlantic coast and New England region. Seven major hurricanes along different tracks from 1620 to 1950 have inflicted light to severe damage (Chapter 22, Figure 22.7) at intervals of a few to more than a 100 years. A great hurricane in 1938 severely damaged more than 240 000 ha of central New England’s forests along a 100-­km swath from central Connecticut and Massachusetts to the northwest corner of Vermont (Foster and Boose  1992; Whitney  1994). Foster (1988b) found that damage to species and stands from this severe windstorm occurred quite predictably and specifically; damage was not heavy everywhere. Similar to patterns found after Hurricane Katrina, post-­hurricane damage was a mosaic of differentially damaged stands controlled by the physiography, wind direction, soil type, and nature of the pre-­hurricane vegetation, especially species composition, tree height, and density (Spurr 1956a; Foster 1988a,b; Foster and Boose 1992). A case study of disturbance across the New England landscape is presented in Chapter 22. Wave-­Regenerated Fir Species  Wave-­regeneration of subalpine fir species is an example of

specific broadscale, wind-­induced disturbance patterns. In high-­elevation balsam fir forests of the northeastern United States (Sprugel 1976; Sprugel and Bormann 1981) and fir forests of central Japan (Sato and Iwasa 1993), persistent wind (rather than wind velocity per se) and other factors maintain fir regeneration and other ecosystem processes over long time periods. The fir canopy is broken by numerous crescent-­shaped strips of standing dead trees, with mature forest on the leeward side of the dead trees and young, vigorously regenerating forest on the windward side (Figure  16.10). This “wave” of death and regeneration moves slowly through the forest, with seedlings replacing the dying trees at the wave’s leading edge. The waves move in the general direction of the prevailing wind at constant speeds of 0.37–­3.3 m per year. The effects of prevailing winds, including winter desiccation, summer cooling, increased rime deposition, and breakage of branches and roots, are presumably the causes of tree mortality of the dieback zone. Shallow, rocky soils of these ecosystems also contribute to death of the fir trees, as strong winds induce tree sway and cause root breakage on sharp rock surfaces, thereby reducing water conducting tissue and encouraging root disease fungi (Harrington 1986). Seedlings established under the canopy trees are released upon canopy tree death and eventually become canopy trees themselves in 60–­80 years (Figure  16.10). The wave-­regenerated forest is unusually orderly due to an omnipresent and highly predictable nature of the prevailing wind disturbance (Sprugel and Bormann 1981).

FLOODWATER AND ICE STORMS Floodwaters have shaped and rearranged physiographic features of streams of all sizes for tens or even hundreds of thousands of years (Chapter  8). Through inundation, sedimentation, and influence on local hydrology, floodwaters control the regeneration of riverine species and thus shape the distinctive pattern of forest vegetation associated with riverine landforms. Flood

Major Disturbances in Forest Ecosystems  405

Mature

Dying– regenerating

Young– intermediate

Mature

Wind

FIGURE 16.10  Diagrammatic cross section through a regeneration wave. Fir regeneration is initiated in and along the edges of the mature stand and released as the trees die. Source: Sprugel and Bormann (1981) / American Association for the Advancement of Science. Reprinted with permission from Science 211 : 390–­393, © 1981 by the American Association for the Advancement of Science.

regimes—­the long-­term duration, frequency, and temperature of floodwater—­have elicited g­enetic differentiation of river valley species. The many different physiological adaptations of river valley species to lack of oxygen for root metabolism enable particular species and races to dominate diverse riverine ecosystems (Gill 1970; McKnight et al. 1981; Hook 1984). Wind rapidly and visually impacts the overstory, but floodwaters slowly and pervasively change the forest site itself by bank cutting and soil deposition. Topographic sites available for establishment of vegetation are constantly created and destroyed by the flux of sedimentary deposits (Malanson 1993). Flooding also controls seedling regeneration by continuously and subtly eliminating upland invaders, favoring physiologically adapted plants and providing water and nutrients, and thus significantly determines the composition and dynamics of riverine vegetation. By altering ecosystem processes, flood control measures and human development of floodplains have changed the successional patterns associated with flood regimes. Many upland species are currently found in bottomland sites where flooding is excluded, emphasizing the enormous selective force of floodwaters on seedling establishment. Damage to overstory and understory trees alike by ice floes and woody debris moved downstream by floodwater can be considerable (Lindsey et al. 1961). Ice or glaze storms occur when rainfall freezes on contact with leaves, branches, or boles of trees, and affects tree growth when branches break under the weight of the accumulating ice. Ice is particularly damaging when ice-­coated branches are subsequently affected by wind. Damaging ice storms occur regularly throughout the eastern United States; Kashian and Barnes (2021), for example, noted ice storm damage in southeastern Michigan that occurred at least twice per decade over the last 60 years. As with many disturbances, some species are more susceptible to ice damage than others (Downs 1938; Whitney and Johnson 1984), particularly fast-­growing species such as black cherry, aspen, willow, and basswood. Oaks, sugar maple, white ash, and hickories are less likely to suffer ice damage, and conifers are particularly resistant to damage even with a constant presence of foliage. Severe ice storms may create canopy gaps, providing an opportunity for susceptible species such as black cherry to grow into the canopy. In addition to direct effects of ice storms on the overstory by severe branch or stem breakage, ice damage may indirectly affect forest tree species by providing entry for pathogens or weakening trees, so they are more subject to insect attack.

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INSECTS AND DISEASE

Photo courtesy of Colorado State Forest Service

Insects and disease-­causing pathogens that kill or reduce the growth of trees affect millions of hectares of forests each year in North America and are major regulators of forest ecosystems (Chapter 12; Romme et al. 1986; Haack and Byler 1993; Castello et al. 1995). Insects and pathogens create widespread disturbances and influence succession by killing or weakening trees singly or in patches. These disturbances reciprocally affect insect populations in many ways (Schowalter 1989). Unlike fire, wind, or even logging (see in the following text), insects and disease affect overstory trees directly without changing the understory, forest floor, or soil. Root fungi, bark beetles, and inner-­bark borers are the primary tree killers, tending to attack one particular species or genus of trees, and strongly control the rate and direction of succession (Franklin et al. 1987). Small gaps resulting from the death of overstory trees killed by pathogens or insects often favor shade-­tolerant species already established in the understory. Larger areas of overstory tree mortality have been shown to result in increases of understory plant diversity and productivity (Pec et al. 2015). Epidemic levels of insects or disease may occur, especially those due to bark beetles, and create a much larger effect on forested landscapes than endemic or background herbivory or disease. For example, the return interval for mountain pine beetle outbreaks in lodgepole pine forests of the northern Rocky Mountains is about 20–­40 years, with an average duration of about 6–­10 years (Cole and Amman 1980; Taylor et al. 2006; Alfaro et al. 2010). These components of the disturbance regime vary with climate and tree density, such that outbreak severity will vary based on these factors. An unprecedented outbreak of mountain pine beetle occurred in British Columbia between the mid-­1990s and about 2017. Peaking in 2005, the rate of spread and severity of the outbreak were the largest recorded in Canadian history, resulting in the loss of 58% of the province’s merchantable pine volume over >18 million hectares (Figure 16.11; Canadian Forest Service 2021). The size and duration of the outbreak have been attributed to a

F I G U R E  1 6 .1 1  Lodgepole pine trees killed by mountain pine beetles near Granby, Colorado. The red-­ needled trees in the photo are those recently killed by the beetle.

Major Disturbances in Forest Ecosystems  407

combination of factors, primarily the simultaneous maturation of lodgepole pine stands across the landscape (due to past logging and effective fire suppression) and warming winter temperatures. Other native insect species have a similar influence on forests. In some Colorado subalpine forests, the effects of spruce beetle outbreaks appear to be as great as those of fire (Veblen et  al.  1991b). Having coevolved with the natural disturbance agents that affect them, forest ecosystems are generally resilient to insect and disease outbreaks, but exotic or invasive insects and disease may have more drastic effects on forests. The effects of some exotic insects and ­diseases are described later in this chapter and in Chapter 21. Finally, insects and disease often interact with other natural disturbances, such as fire and wind; these disturbance interactions are also discussed in the following sections.

CATASTROPHIC AND LOCAL LAND MOVEMENTS Catastrophic events such as volcanism, earthquakes, and landslides create new ecosystems and shape species occurrence and distribution. Volcanic activity on several continents has provided new substrate for forest development, such as in the Cascades of the Pacific Northwest, the Andes in Chile (Veblen  1987), the Hawaiian Islands, and the Changbai mountains of northeastern China (Barnes et al. 1992). The 1980 eruption of Mt. St. Helens has provided abundant evidence of the overwhelming effects on existing vegetation and the rapidity of succession on newly formed substrates as a result of biological legacies of diverse kinds (Franklin et al. 1985; Franklin et al. 1988; Franklin 1990; Dale et al. 2005c; Del Moral and Chang 2015). Landslides in mountainous terrain provide new substrates for plant occupancy (Pabst and Spies 2001). For example, a 1960 earthquake in the southern Chilean lake district triggered volcanic activity and thousands of landslides that provided excellent regeneration sites for southern beeches (Nothofagus spp.) (Veblen 1987). Such broadscale disturbances occur frequently enough to influence forest composition and structure over long temporal scales. Landslides and rock avalanches initiated by prehistoric earthquakes also shaped landscapes and changed forests in mountains of the Puget Sound area of Washington State about 1000 years ago (Jacoby et al. 1992; Schuster et al. 1992). Catastrophic and local debris avalanches—­rapid downhill flows of soil, rock, and vegetation almost always preceded by heavy rains—­are major disruptive and land-­shaping forces in the Appalachian Mountains (Hack and Goodlett 1960; Eschner and Patric 1982; Jacobson et al. 1989).

LOGGING Lumbering was a major factor in the development of the United States, and major changes in forest composition followed it. Post-­logging fires, fueled by woody debris left after logging, also contributed significantly to change in forest composition (Whitney 1994). Aside from large-­scale clear-­cutting, logging as a disturbance may be similar to wind in that tree harvesting removes the overstory and releases the understory. Many understory hardwood species cut in logging-­related operations resprout to form vigorous and fast-­growing stems that then compete for overstory space. As a general rule, the competitive ability of post-­harvest regenerating trees will depend on the intensity and pattern of the cut. Light, partial cuts such as thinning and improvement cutting favor tolerant species, particularly those already established in the understory, and tend to accelerate forest succession rather than return it to an earlier stage (Figure  16.12). Moderately heavy partial cuts or cutting of small groups of trees to mimic the formation of canopy gaps (group selection cutting) favor mid-­tolerant species. Clear-­cutting or the creation of large gaps with a diameter of at least twice the height of the stand favors the invasion of pioneer species, particularly if mineral soil is exposed by the harvest operation. A forest manager can therefore greatly influence subsequent forest composition and the rate of succession by regulating the intensity and pattern of cutting.

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F I G U R E  1 6 .1 2  Partial cutting in this eastern white pine stand in Maine has favored advanced regeneration of tolerant red spruce. Source: U.S. Forest Service photo.

The differential effect of partial cutting versus clear-­cutting on forest succession is clear in the Douglas-­fir region of Washington and Oregon. Older stands that are cut are more or less even-­aged Douglas-­fir dating to forest fires that occurred 150–­500 years ago (Franklin et al. 1981). Clear-­cutting in such stands during or immediately after a seed year provides the reestablishment of even-­aged Douglas-­fir. Clear-­cut patches ranging from 1 to 15 ha in size will normally regenerate to Douglas-­fir within a few years with seed dispersed from surrounding uncut forest. On any site, the new forest that follows a clear-­cut is composed of pioneer species that range from shade intolerant to mid-­tolerant in their competitive ability in the understory in that region. On clear-­cuts on the lower and wetter sites in the Coastal Range, red alder colonizes prolifically as a pioneer species. Although large-­area clear-­cutting and planting of Douglas-­fir was a conventional forestry practice for several decades, an increasingly common view is that Douglas-­fir can be regenerated and grown under a partial overstory created with a variable retention harvest system or shelterwood (Chapter 14, Figures 14.16–­14.18). Logging also affects forest succession by the differential removal of one species over another—­thus changing the composition of the forest. Such effects are exemplified in the mixed-­ wood forests of Maine where logging has occurred nearly continuously for almost 150 years (Whitney 1994). Early logging focused on large white pines suitable for masts for wooden sailing ships and house construction, but later shifted to the large-­scale cutting of red spruce to build and supply sulfite and groundwood pulp mills. Removal of the red spruce from a community that included tolerant balsam fir, sugar maple, and beech created a mixed forest of these species. Balsam fir was more utilized as the supply of red spruce decreased. Partial cutting was soon replaced by clear-­ cutting, resulting in a rising water table that precluded forest in many of the wetter flats; pioneer species such as trembling aspen and paper birch colonized the drier sites. Thus, forest composition of central and northern Maine has been markedly shaped by successive waves of logging, each concentrating upon different species.

Major Disturbances in Forest Ecosystems  409

LAND CLEARING Many forests occur in humid regions that are suitable for the raising of agricultural crops. The great majority of the world’s population is found in forested regions, and thus much of the world’s forests have inevitably been cut and the land cleared for agriculture. The massive forest clearance that occurred in the eastern United States and its effects are described by Whitney (1994). In the tropics, forest land use for agriculture is often (but not always) transitory, with fields being allowed to revert to forest again after a few years. Even in the temperate zone, much land cleared and formerly used for farming has been found unsuitable for continued cropping or has been supplanted by bringing better lands into production. This land has been planted to forest or allowed to revert naturally to forest; many currently forested lands belonging to State and National Forests in the Upper Midwest have followed this process. Secondary forest succession following land abandonment is therefore an important process, occurring over many hundreds of thousands of hectares in well-­settled forest areas.

DISTURBANCE INTERACTIONS As described in the earlier text, most ecosystems are subjected to more than one type of disturbance. The result of those interactions may be cumulative (increasing impact) or negative (e.g., the second disturbance is less impactful or less likely to occur), or the disturbances may interact to produce a novel or unexpected result (Paine et  al.  1998). Interacting disturbances—­specifically that the effects of one disturbance may either change the likelihood or alter the effects of another—­ have long been recognized by forest managers and ecologists. For example, extensive blowdown of subalpine forest in 1939 led to an epidemic of spruce beetles in the 1940s that devastated 290 000 ha of the White River National Forest alone (Hinds et  al.  1965). The largest and/or most severe ­wildfires typically occur during periods of prolonged drought, as in the infamous Peshtigo Fire of Wisconsin in 1871, the massive western fires of 1910, and the 1988 fires in and around Yellowstone National Park, all of which burned through prior burn locations. Fire-­scorched or drought-­stressed trees are often sentinels to successful attack by bark beetles and other insects (Mattson and Haack  1987). Mudslides often follow fire in steep-­slope, chaparral-­dominated ecosystems of California and in other ecosystems of the western United States. The pattern and nature of multiple disturbances are characteristic of regional ecosystems with their distinctive climates (Mediterranean, northern continental, coastal/marine), physiographic features (mountains, plains), and soil/hydrology (excessively drained, poorly drained). Interacting disturbances have been studied extensively over the last two decades and may be characterized as two types of interactions (Buma 2015). Disturbances that alter ecosystem resistance, by changing the likelihood, intensity, severity, or size of the subsequent disturbance, are called linked disturbances (Simard et al. 2011). Linked disturbances occur when the first disturbance creates a legacy, such as reduced availability of host trees for an insect or altered distribution of fuels, that mechanistically affects the second. In a study of a spruce beetle outbreak in the 1940s in western Colorado, Kulakowski et al. (2016) found that young stands of Engelmann spruce and subalpine fir that regenerated from fires in the late-­nineteenth century were not susceptible to the spruce beetle outbreak. Thus, the location and severity of the spruce beetle outbreak were altered by legacies of the fires in the late 1800s (in this case, lack of mature, susceptible trees). Other examples of linked disturbances include those presented by Kulakowski and Jarvis (2013), Kulakowski et al. (2012), Simard et al. (2011), Lynch et al. (2006), Bigler et al. (2005), and Veblen et al. (1994). When disturbances alter the resilience of an ecosystem to a subsequent disturbance such that the interacting disturbance changes the likelihood or rate of recovery of the ecosystem, the interaction is the result of compounded disturbances. In a study of American beech dominance in coastal New England, Busby et al. (2008) found that intensive forest harvesting in the nineteenth

410  C h a p t e r 1 6  Disturbance

century favored regeneration of oaks and beech, but repeated hurricanes thereafter favored beech persistence in the understory over oaks. A severe hurricane in 1944 and heavy deer browsing released beech saplings (typically avoided by deer) while preventing further oak establishment. The reduced resilience of oaks as a result of compounded disturbances in this example has created long-­term changes in forest composition and diversity (Busby et al. 2008). Other examples of compounded disturbances include those presented by Andrus et al. (2021), Carlson et al. (2017), Kulakowski et al. (2013), and Girard et al. (2009). Interestingly, the same events can simultaneously result in both linked and compounded disturbance interactions (Buma 2015), as illustrated by two independent studies that examined the interaction of a severe wind event with subsequent fire in the subalpine forests of northwestern Colorado (Figure 16.13). Kulakowski and Veblen (2007) found that the changes in fuels resulting from the blowdown altered the severity (but not the size) of the fire that followed, suggesting that wind and fire in this case were linked disturbances. In a separate study in the same area, Buma and Wessman (2012) found that the blowdown affected resilience of the system by altering seed dispersal into the areas disturbed by wind and severe fires. In particular, wind-­damaged areas were too large to be revegetated by spruce and fir seeds dispersed from adjacent unburned forests, and

Initial blowdown event

Le

ga

cy

~5 years Altered spatial distribution of downed woody debris

Altered vertical fuel structure

Altered distribution of high-severity burn patches to due altered spatial fuel structure, Kulakowski and Veblen 2007

Low regeneration by seed dispersal species due to more homogenous burns, Buma and Wessman 2012

Low regeneration by serotinuous species due to altered fire characteristics, Buma and Wessman 2011

Burn temperature Burn duration

Me

ch a im nism pa ct and

Subsequent fire event

Linked interaction

Compound interactions

F I G U R E  1 6 .1 3  The interaction of a large wind event and subsequent fire in Colorado show that linked and compound interactions may occur from the same events. Burn severity was higher in the areas affected by wind (linked disturbance interaction), but the blowdown reduced post-­fire resilience by limiting seed dispersal into the large and severely burned patches from adjacent unburned areas and from serotinous cones within the area (a compound disturbance interaction). Source: Buma (2015) / John Wiley & Sons.

Biotic Composition Changes  411

the increased severity of the fires in the wind-­affected areas consumed the serotinous cones of lodgepole pine and the seeds therein. These mechanisms reduced the ability of the ecosystems to recover, suggesting a compounded disturbance interaction (Figure 16.13). An excellent review of disturbance interactions and their ecological implications is provided by Buma (2015).

BIOTIC COMPOSITION CHANGES

Secondary succession originates primarily from fire, windstorms, logging, and land clearing, but a second major kind of disturbance—­the addition or subtraction of plant or animal species—­will alter the trajectory of forest succession. These changes in the flora or fauna may create considerable disturbance to existing ecosystems and thus may well be considered as initiating secondary succession in a very real sense.

ELIMINATION OF SPECIES When a species is eliminated or greatly reduced in abundance, its place in an ecosystem is typically taken by other organisms. In many instances, species loss is the result of the introduction of exotic insects or pathogens to which native tree species have not adapted. There are numerous examples of species’ elimination or near-­elimination that have occurred over the last 100 years as a result of exotic insect or pathogen introduction. The American chestnut, once a major overstory dominant in eastern North America (Braun 1950), was virtually eliminated from the eastern deciduous forest by the 1940s by a pathogen introduced from Asia in 1904. American chestnut was replaced primarily by its major associates in New England, the mid-­Atlantic states, and the southern Appalachians (Keever 1953; Woods and Shanks 1959; Good 1968; Stephenson 1974). Dutch elm disease, caused by two related pathogens introduced from Europe in the early-­twentieth century, drastically reduced the abundance of mature American elm trees in swamp and river floodplain forests of the eastern and midwestern United States, replaced by a variety of different species depending on site conditions. In southeastern Michigan, American elm has not been eliminated from deciduous swamps but remains as an understory and subcanopy overstory species rather than a dominant overstory species (Barnes 1976). It will likely persist by seeds from young elm trees with an average life span that is drastically reduced. Emerald ash borer is an invasive insect introduced into the metropolitan Detroit area from Asia in the 1990s. Specific to the ash genus, the insect had spread to 36 states and five Canadian provinces by 2022, killing nearly all ash trees larger than about 1 in. in diameter once it reaches high population levels. Where ash was found in mixed stands, succession has mainly featured replacement by associates depending on the species and site conditions (Kashian and Witter 2011). However, pure stands of black ash, typically found on poorly drained soils, may be potentially converted to shrub wetlands or other cover types if ash mortality is high (Kolka et al. 2018; Palik et al. 2021). Like American elm, most ash species are likely to persist, but only as seedlings, saplings, and sprouts from top-­killed overstory trees and the seeds they produce (Kashian 2016). Effects on forests of invasive species such as chestnut blight, Dutch elm disease, and emerald ash borer are discussed in detail in Chapter 21. Reduction of tree species in a community is not always the result of the introduction of exotic species. For example, in the spruce–­fir forests of eastern Canada and the northeastern United States, spruce budworm epidemics have greatly reduced the abundance of overmature and mature balsam fir (together with lesser amounts of black and white spruces) over large areas. The budworm apparently has played a major role in these mixed softwood forests in suppressing the proportion of the very-­tolerant balsam fir compared to that of the somewhat less-­tolerant spruces (Bognounou et al. 2017). Notably, the response of balsam fir in this case is a local reduction in abundance rather than a widespread elimination because it has coevolved and adapted to the long-­ term presence of the native spruce budworm.

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In the animal portion of forest ecosystems, the virtual elimination of many predators from most forests in North America—­particularly wolves, bears, cougars, and lynx—­has played a role in increasing the populations of their prey: deer, rabbits, and other herbivores. This in turn results in greater browsing of the understory, including tree regeneration. Such disturbance precludes the regeneration of black cherry in areas of Pennsylvania (Marquis 1974, 1975, 1981; Tilghman 1989) and hemlock in many places in the northern United States (Frelich and Lorimer 1985; Peterson and Pickett 1995). Reintroduction of wolves into Yellowstone National Park in 1995 had significant effects on elk populations, both in their population size (reduced by predation) as well as their foraging behavior. As a result, browsing was reduced and recruitment to the overstory was increased for trembling aspen and various species of cottonwoods and willows (Ripple and Beschta 2003, 2012; Beschta and Ripple 2016).

ADDITION OF SPECIES The species occupying a given site may simply be those that have access to that site at the time of its availability rather than those best adapted to compete and grow on that site. Invasion of the site by better competitors often results in substantial changes in forest succession. The chestnut blight fungus, Dutch elm disease fungus, and emerald ash borer cited in the previous section are examples of accidental introductions that have greatly modified forests. Other organisms may be trees, other higher plants, fungi, bacteria, and animals of all kinds (Chapter 21). Important deliberate forest-­tree introductions in Europe include Douglas-­fir, eastern white pine, northern red oak, and black locust; in the northeastern and north-­central United States: Scots pine, Norway spruce, Norway maple, black pine, white mulberry, Chinese tree of heaven, Siberian elm, and European larch; and in temperate zones of the Southern Hemisphere: Monterey pine, patula pine, and slash pine. These and many others have become vigorous and often dominant species in the local flora. The effects of additions of such species to forests is the subject discussed in detail in Chapter 21.

SUGGESTED    R E A D I N G S Agee, J.K. (1993). Fire Ecology of Pacific Northwest ­Forests. Washington D.C.: Island Press 493  pp. ­Chapters 5 and 7–­12.

Paine, R.T., Tegner, M.J., Johnson, E.A., and E.A. (1998). Compounded perturbations yield ecological surprises. Ecosystems 1: 535–­545.

Covington, W.W. and Moore, M.M. (1994). Postsettlement changes in natural fire regimes and forest structure: ecological restoration of old-­growth ponderosa pine forests. J. Sustain. For. 2: 153–­181.

Pickett, S.T.A. and White, P.S. (1985). Natural Disturbance and Patch Dynamics. New York: Academic Press 472 pp.

Heinselman, M.L. (1973). Fire in the virgin forests of the boundary waters canoe area, Minnesota. Quat. Res. 3: 329–­382.

Sprugel, D.G. and Bormann, F.H. (1981). Natural disturbance and the steady state in high-­altitude balsam fir forests. Science 211: 390–­393. Whitney, G.G. (1994). From Coastal Wilderness to Fruited Plain. Cambridge: Cambridge Univ. Press 451 pp.

Forest Succession

CHAPTER 17

A

ll components of ecosystems change over time, including climate, landforms, soil, and biota, and we have discussed the basis of dynamic change in forest ecosystems in previous chapters. Change may be gradual or rapid depending on the particular disturbances that characterize regions and local sites (Chapter 16). Ecosystem development and change are most easily observed in vegetation, especially the forest trees, which may change dramatically over time following ­disturbances such as fires or windstorms. One manifestation of ecosystem change is forest succession, which is often characterized as the change in species composition or the replacement of the biota of a site by one of a different nature. It is notable that other aspects of ecosystem change, such as soil development, have also been documented (Olson  1958; Miles  1985; Matthews  1992), and changes involving biomass accumulation and nutrient cycling are described in Chapters 18 and 19. In this chapter, we focus upon changes in species composition. Importantly, forest succession is what happens in one place over an extended period of time, measured in tens to thousands of years. In other words, succession happens with space fixed and time changing (Rowe  1961b). It is characterized by a sequential change in the relative structure, kind, and abundance of the dominant species. We focus here on the intervals between disturbances and within a time period of the same order of magnitude as the life span of the longest-­lived organisms in the successional sequence (tens to thousands of years). Forest succession progresses in nearly infinite ways and is driven by many different factors, and many other processes occur simultaneously with it. However, succession can only have meaning in the context of a particular geographic framework defined by regional and local ecosystems. The heart of the matter is that location-­specific ecosystems undergo development, change, and succession. The way a given ecosystem changes when disturbed is displayed more or less visibly according to the resilience and inertia of its component parts. Our attention is quickly drawn to the visibly changing vegetation, but attempts to explain and predict its successional path separate from its geographic context are unlikely to be rewarding. Because vegetation is most easily observed, plant succession has historically been defined in the context of plant species or communities. Definitions include those of Grime (1977): “a progressive alteration in the structure and species composition of the vegetation” and Finegan (1984): “the directional change with time of the species composition and vegetation physiognomy of a single site.” Forest succession obviously deals with changes in biota, but the way we conceive its context (plant or ecosystem, part or whole) is critical in understanding the patterns and processes of succession over the complexity of regional and local ecosystems. Troll (1963b) was the first to emphasize and explicitly describe succession as a landscape process involving all ecosystem components, including climate, physiography, and soil as well as vegetation. Adopting vegetation as a primary focus of succession contributed to the development of an elaborate and often criticized concept, with literally thousands of published papers introducing additional terminology and controversy. A perceptive review by Pickett et al. (1987) clears up much

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

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414  C h a p t e r 1 7   Forest Succession

confusion, and a landscape ecosystem perspective provides the spatial framework for understanding how succession works. All successions have these key characteristics that affect their course and rate: (i) a sequence of concomitant environmental and vegetational change that may be characterized by stages for convenience and practical applications; (ii) disturbance regimes; and (iii) mechanisms (processes) occurring at different times in the sequence. Above all, the possible successional sequences of interacting organisms, mechanisms, and disturbances are controlled or mediated by the regional and local landscape ecosystems where succession takes place. In this chapter, we examine successional concepts, the evolution of the concept of forest succession, causes of ecosystem and forest change, and the mechanisms involved. The focus is of course on vegetation change but placed in an ecosystem context as far as possible. Examples of primary and secondary succession are presented to illustrate succession in diverse settings. For details beyond the scope of our treatment, the following books and reviews provide diverse viewpoints: Clements (1916, 1936); Tansley (1929, 1935); Drury and Nisbet (1973); Connell and Slatyer (1977); Golley (1977); Miles (1979, 1987); White (1979); West et al. (1981); Finegan (1984); Pickett et  al. (1987); Glenn-­Lewin et  al. (1992); Matthews (1992); Peet (1992); Worster (1994); Schmitz et al. (2006); Swanson et al. (2011); Walker and Wardle (2014); and Prach and Walker (2020).

BASIC CONCEPTS OF SUCCESSION

PRIMARY AND SECONDARY SUCCESSION Primary succession is ecosystem change that occurs on previously unvegetated terrain, uninitiated by a disturbance. For example, primary succession on recently deglaciated terrain in Alaska eventually supports spruce forest after first being colonized by communities of willow or alder. Secondary succession follows a disturbance to an existing forest that disrupts ecosystem processes and destroys or disrupts existing biota. The distinction between primary and secondary succession is often more arbitrary rather than real, but it is followed here as a matter of convenience in understanding the dynamics of forest ecosystems.

BIOLOGICAL LEGACIES Biological legacies that follow disturbances such as fires, floods, and windstorms distinguish the course of ecosystem change during secondary succession. Biological legacies are biologically ­generated elements of a pre-­disturbance ecosystem that persist in some form after a disturbance occurs (Franklin et  al.  2018). Examples of biological legacies include: (i) living organisms or ­portions of organisms that survive a disturbance, (ii) organic debris, and (iii) biotically derived patterns in soils and understories (Franklin 1995). Living legacies include intact plants and animals, rhizomes, and dormant spores and seeds. Important biotically derived structures include standing dead trees (snags) and fallen logs (coarse woody debris), large soil aggregates, and dense mats of fungal hyphae. These diverse biological legacies, remaining from pre-­disturbance ecosystems, strongly influence the paths and rates of forest recovery and ecosystem function by shaping ­post-­disturbance community structure and/or providing substrate for new colonizers. In fact, the types and abundances of biological legacies are among the most important ecological outcomes of disturbances (Franklin et al. 2018), although they differ sharply among different kinds of disturbances (Table 17.1).

SUCCESSIONAL PATHWAYS, MECHANISMS, AND MODELS Pickett et al. (1987) define the terms successional pathway, mechanism, and model. A ­successional pathway is the temporal pattern of vegetation change, which typically shows change in community

Basic Concepts of Succession  415

Table 17.1  Biological legacies generated by different types of severe disturbances. Disturbance type Biological legacy

Severe wildfire

Wind

Insect outbreaks

Clearcut harvesting

Volcanoes

Live trees

Few

Few or absent

Depends on forest composition

Few, confined to margins

Absent

Snags

Abundant

Few

Abundant

Abundant

Absent

Downed woody debris

Common

Abundant

Eventually abundant

Abundant

Absent

Uproots

Absent

Abundant

Absent

Few

Absent

Intact understory

Rare

Abundant

Abundant

Absent

Very little

Mineral seedbed

Abundant

Abundant with uproots

Absent

Absent if buried

Abundant

Source: Adapted from Franklin et al. (2018) and Swanson et al. (2011).

Trembling aspen

Fire severity (y)

Jack pine

Black spruce

50

Times since fire (x)

>250

F I G U R E  17.1  Alternative successional pathways for aspen–­conifer forests in the clay belt of northwestern Quebec, Canada. Forest composition and structure are strongly determined by the severity of the last fire (y-­axis) and the time elapsed since the last fire (x-­axis). High-­severity fires initially result in forests ranging from aspen-­dominated to a mixture of jack pine and black spruce. Low-­severity fires result in black spruce-­ dominated forests across a range of density. With time, forests generally converge toward open, multi-­aged black spruce forests. The rate and direction of forest succession on this landscape are strongly influenced by site factors, especially soil type. Source: From Franklin et al. (2018).

types and may describe the decrease of particular species populations. Figure 17.1 is an excellent example of successional pathways because the initial site conditions in conjunction with disturbance are major factors in determining ecosystem change. A successional mechanism is an interaction or process that contributes to successional change. Examples include general ecological processes such as dispersal, competition, and establishment, and physiological processes such

416  C h a p t e r 1 7   Forest Succession

as biomass allocation (Chapter 18) and nutrient uptake (Chapter 19) that affect plant form and reproductive ability. A successional model is a conceptual map that explains a successional pathway by identifying and specifying the relationship among the mechanisms and the various stages of the pathway.

AUTOGENIC AND ALLOGENIC SUCCESSION Successional change has traditionally been characterized as either autogenic (by endogenous factors) or allogenic (by exogenous factors) depending whether successive changes are brought about by the action of the plants themselves on the site or by external factors (Tansley 1929). Primary succession is an autogenic process in the classical view because it is plant-­driven, and secondary succession is allogenic because it is driven by periodic disturbances. The dichotomy between autogenic and allogenic factors is actually quite artificial: the causal factors don’t fit strictly into one category or the other (White 1979; Matthews 1992). Tansley (1929, p. 678) recognized this problem when he proposed the concept, cautioning that real successions show a mixture of the two classes of factors. Today, the ecosystem rather than the plant is or should be the focus, because change is elicited by a combination of factors both within the ecosystem and external to it. Allogenic factors (physical processes and disturbances such as fires and windstorms) may be initiated outside a given ecosystem, but their effects often strongly depend upon site conditions (e.g., topography, aspect, soil depth, drainage) and the associated vegetation within a given ecosystem. For example, the aspect, topography, soil properties, and vegetation of an ecosystem, and its position in relation to fire breaks, strongly affect fire frequency and severity. Windstorm effects are significantly related to soil depth, drainage, rooting ability, and the height of tree species of the affected ecosystem. Furthermore, autogenic factors of litter accumulation and plant-­induced changes in microclimate, soil water, and nutrients all affect the outcome of allogenic disturbances. In turn, autogenesis is almost always influenced by allogenic factors. Thus, allogenic and autogenic factors are closely interrelated.

HOW IS SUCCESSION DETERMINED? In either primary or secondary succession, ecosystem and vegetation change should ideally be tracked at one place over time such that an actual successional pathway for a given ecosystem might be determined. However, such monitoring is rarely possible due to the long life span of forest trees, and much of what we know about succession therefore comes from short-­term observations at different places. The typical approach used to characterize long-­term succession is to infer successional trends from stands of different ages, all of which are assumed to have developed on the same site and under the same disturbance conditions. This approach has come to be known as a chronosequence (Major 1951) or space-­for-­time substitution (Pickett 1988). If constructed carefully and replicated, chronosequences can be indispensable for inferring how forest ecosystem structure and function change over centuries (e.g., Kashian et  al.  2013). A chronosequence approach is not without its problems, however (Matthews  1992; Johnson and Miyanishi 2008). Chronosequences are generally most appropriate when successional pathways are likely to ­converge, biodiversity is low, and disturbance is infrequent and/or not severe (Walker et al. 2010). Chronosequences are less suitable when successional pathways are likely to diverge, or the ­ecosystems are species-­rich, highly disturbed, or arrested in time. The actual chronosequence of ecosystem and forest change is one that is site-­specific over time, and ideally this should be the process that is observed. Several examples of chronosequences are described later in this chapter.

Evolution of the Concept of Forest Succession  417

EVOLUTION OF THE CONCEPT OF FOREST SUCCESSION

The dynamic nature of forests was recognized by naturalists in North America and Europe as early as the eighteenth century in a form that approximates a plant-­based (bioecosystem) concept of forest succession (Spurr 1952a). Such ideas gained significant traction in the nineteenth century. In Europe, beginning with Hundeshagen in 1830, observed changes in forest composition were the subject of specific articles by professional foresters and botanists. Hundeshagen pointed out instances of spruce replacing beech and other hardwoods in Switzerland and Germany, and of spruce and other species taking the place of birch, aspen, and Scots pine. The first detailed North American report of species composition changes was apparently that of Dawson in 1847. Working in the Maritime Provinces of eastern Canada, Dawson recognized the effects of windthrow and fire in the forests found by European colonists and distinguished between successional trends in small clearings following cutting, following a single fire, following repeated fires, and as a result of agricultural use of land. As early as 1860, the writer Henry David Thoreau recognized that pine stands on upland soils in central New England were succeeded after logging by even-­aged hardwood stands that today constitute the principal forest type of the region. He named this trend “forest succession” and published an essay on “The Succession of Forest Trees.” His late natural history writings, published long after his death (Thoreau  1993), are remarkably filled with insights on ecological succession: pines are “pioneers” to oaks and “lusty oaken carrots” that are a “remarkable special provision for the succession of forest trees.” In articles published in 1875 and 1888, Douglas discussed at some length the concepts of forest succession and pioneer species, and presented an explanation of how short-­lived, light-­seeded pioneer species formed the first forest types on burned-­over pine land.

FORMAL ECOLOGICAL THEORY A general and formal theory of plant succession evolved slowly but was well established by the beginning of the twentieth century when several American ecologists systematized its study. The foundations of plant ecology as a study of vegetation dynamics and succession were initiated by Henry C. Cowles (1899, 1901), who analyzed succession on sand dunes of Lake ­Michigan, beginning with uncolonized sand and ending with a mature forest. Although dunes were formed and destroyed by wind, he found that “vegetation profoundly modifies the topography.” In 1911, Cowles noted that the original plants in any habitat give way “in a somewhat definite fashion to those that come after.” Furthermore, he recognized that changes in vegetation caused by climatic, topographic, and biotic changes occurred concurrently, at different rates, and at times in different directions. A contemporary, Frederic E. Clements, next developed an elaborate philosophical structure of plant succession (1916, 1949) that attempted to categorize and formalize all eventualities of plant community change. Specific examples of forest succession were documented by William S. Cooper, with his studies of Isle Royale in Michigan (1913) and the colonization by plants following glacial retreat in Glacier Bay, Alaska (1923, 1931, 1939). Clements, in particular, developed a complex nomenclature to describe plant succession in a meticulous, tidy organization of nature that has both facilitated and greatly complicated the efforts of his successors. Some of his terms have taken a permanent place in the vocabulary of ecologists (vegetation dynamics, biome), others have persisted but with broadened and changed meanings (climax), while many have been dropped from general usage (cyriodoche, consocies). The work of Clements is described in detail in the ­following text.

418  C h a p t e r 1 7   Forest Succession

HOW DOES SUCCESSION WORK?

Succession involves many factors and processes operating within a geographic framework at multiple scales (Table 17.2). Understanding the key characteristics of ecosystem change and what they might tell us about succession is critical for examining how succession works. Key characteristics determining actual vegetation change (with space fixed and time changing) include the following: The physical site conditions of the Earth’s regional and local ecosystems provide a geographic framework that determines the organisms involved and the rate and trajectory of succession. Disturbance factors and processes associated with particular regional and local ecosystems provide substrates for primary and secondary succession and periodically affect the rate and trajectory of succession. Colonizing organisms have an enormous range of life-­history traits (genetic adaptations of structural architecture, longevity, and physiological processes) and regeneration strategies that are tightly linked to site conditions and disturbance regimes. Changes in  site conditions and organisms occur reciprocally, with one affecting the other and vice versa. Organisms interact through competition and mutualisms. With these characteristics we can expect that: (i) successional pathways are likely to be unique, and no two pathways will be the same; (ii) different successional pathways will occur where regional and local ecosystems are also diverse; (iii) there are a great number of different and simultaneously operating mechanisms that control the rate and trajectory of succession; (iv) distinguishable ­patterns and stages of succession are evident in ecosystems not severely affected by humans; (v) species occurring early in succession modify the environment in ways that may limit their own regeneration and may favor or inhibit the regeneration and growth of species occurring late in

Table 17.2  Causes of succession: Hierarchical summary of factors and processes. Landscape ecosystem characteristics Regional climate, physiography, vegetation patterns Local climate, physiography, landform, soil, species distribution Site availability Disturbance (kind, periodicity, size, severity, dispersion, disturbance history of target site and adjacent ecosystems) Differential species dispersal and establishment Landscape position in relation to seed source availability Reproductive mode of species (sexual, asexual) and fecundity Dispersal (dispersal agents, landscape configuration) Propagule pool for establishment (time since land disturbance, land use treatment) Differential species performance (in regeneration and post-­establishment development) Genetic adaptations and plasticity (germination and establishment traits, understory tolerance, assimilation and growth rates, biomass allocation, plant defense) Competition and mutualism (presence, kind, amount) Environmental stress (climatic cycles, site history, prior occupants) Allelopathy (soil chemistry, neighboring species)

How Does Succession Work?  419

succession; and (vi) although succession proceeds at various rates, there is no end point of the process. With these key characteristics, what are the significant features of how succession works? We may summarize the general trends of how succession works as follows: 1. Following a disturbance that creates either (i) new land (unvegetated substrate for primary succession) or (ii) a disrupted forest with a vegetated site having remnants of the ­pre-­disturbance forest (with secondary succession following), substrate is available for recolonization or continued growth by organisms. 2. Species from surrounding areas or within the disturbed area itself occupy the disturbed site in varied patterns and temporal sequences. Where a disrupted forest undergoes secondary succession, new colonists may be dispersed into the site or are already present in surviving vegetation, the soil seed bank, or the vegetative propagule bank (Figures 4.2 and 4.13, Chapter 4). 3. The colonizing plants develop on the site and either become established or die. 4. Plants interact through competition and mutualisms and thereby affect further plant establishment and development over time. 5. Plants change site conditions, such as light, temperature, or soil properties, and processes which may or may not affect them as they continue to interact with other plants and animals. 6. All of the abovementioned processes are modified by multiple disturbances that are ecosystem-­specific in type, frequency, and severity. 7. Some species persist, thrive, and dominate at their time during succession, whereas others drop out or diminish in abundance. The nature and rate of change of vegetation depend on the spatial and temporal scales under consideration, the site conditions of regional and local ecosystems, species’ life-­history traits and regeneration strategies, disturbance regimes, chance effects, and other processes (Table 17.1). If it is known which species are present, which are in the vicinity of a particular land area, and the ecology of the species in the context of the various land-­vegetation patterns, then it is possible to make an educated guess as to what is likely to happen in the future (Rowe 1961b).

CLEMENTSIAN SUCCESSION It is instructive to compare the scenario just described with Clements’ model of succession, the first to describe the vegetative approach and a legacy for our current understanding and the basis for opposing ideas. Clements (1905; Weaver and Clements 1929) declared what he saw as the rules of succession in no uncertain terms. The basic processes were clear cut, and stages proceeded in a systematic, stepwise, directional, and predictable process (Figure 17.2). In this model, ecesis (germination, establishment, and growth) occurs immediately after plants reach the site (migration). Dominant plants change the environment (reaction) to favor or “facilitate” the dominance of the next community that can better compete at the site. The end point of a given succession is reached as the climax formation or association is attained. The plants that dominate at the climax stage modify their environment such that they perpetuate themselves, resist change, and therefore create the community of maximum stability. Clements also assumed that all primary successions in a climatic region eventually converge to the same climax community from multiple starting points (monoclimax). A critical flaw of Clements’ was that in emphasizing the influence of plants upon

420  C h a p t e r 1 7   Forest Succession DISTURBANCE MIGRATION ECESIS Biotic Reactions Favor New Migrants

BIOTIC REACTION

COMPETITION Biotic Reactions Favor Already Established Dominants

F I G U R E   17. 2  Classical (Clementsian) model of autogenic plant succession. Disturbance brings about nudation followed by arrival of propagules (migration) that germinate and grow (ecesis) and modify the site (biotic reaction) in competition with one another. Ecesis, reaction, and competition often occur simultaneously. Competition and reaction continue in cycles (left) that favor new immigrants and hence new communities; ultimately (lower right) the reactions lead to the climax or end stage. Source: From Christensen (1988b) / University of Washington Press.

CLIMAX

the habitat (i.e., biotic reaction in Figure 17.2); he saw the driving force of succession to be changes caused by internal factors (i.e., autogenic succession). It is therefore important to consider the stages of succession as a basis for examining insights, problems, and misconceptions about orderly succession.

STAGES OF SUCCESSION In describing plant succession, we begin with a hypothetical unvegetated substrate or disturbed forest and describe the successive plant communities that will occupy this site. Doing so requires assumptions that (i) the regional climate will remain unchanged, and (ii) disturbances such as severe windstorms, fires, or insect outbreaks are absent. Such assumptions are unrealistic given the hundreds or thousands of years involved in forest successions. However, these assumptions provide the view of succession as a more or less orderly and predictable sequence that depends upon the character of the physical site and associated biota. The assumptions also provide the framework to examine some problems of this perspective. The recognition of stages is a matter of convenience rather than of their actual occurrence. At best, stages are simply wave-­like replacements of species with similar ecologies, and at worst, they are arbitrary divisions in a continuum (Matthews 1992). It is notable that depending on site and species availability, plant communities vary in their rate of change as new species colonize the site and existing species either reproduce or disappear. The arbitrary classification of this ­continuum into stages, characterized by the dominance or presence of certain species and plant life-­forms, is certainly useful and a convenience worth maintaining.

Primary Succession  Unvegetated sites range from pure mineral material (rock or soil) to water; plant colonization and growth are most favored by mixtures of soil and water (i.e., moist, well-­drained mineral soil). Primary plant succession beginning with dry rock material (either as rock or as mineral soil) is termed a xerarch succession; that beginning with water is termed a hydrarch succession; while that beginning with moist, aerated soil is a mesarch succession. The major stages of primary succession are generally consistent, but many types of primary successions exist (Matthews 1992). Both the specific successions and the vegetational stages within each are arbitrarily chosen. In Table 17.3, a series of 10 stages is given for each representative type of primary succession. Notably, some stages are omitted under conditions where the next ­successional life-­form (tree, shrub, herb) is capable of directly colonizing an earlier vegetational

How Does Succession Work?  421

Table 17.3  Stages of primary succession. Stage 1

Xerarch Dry rock or soil

Mesarch Moist rock or soil

Hydrarch Water

2

Crustose lichens

(Usually omitted)

Submerged water plants

3

Foliose lichens and mosses

(Usually omitted)

Floating or partly floating plants

4

Mosses and annuals

Mostly annuals

Emergents

5

Perennial forbs and grasses

Perennial forbs and grasses

Sedges, sphagnum, and mat plants

6

Mixed herbaceous

Mixed herbaceous

Mixed herbaceous

7

Shrubs

Shrubs

Shrubs

8

Intolerant trees

Intolerant trees

Intolerant trees

9 10

Mid-­tolerant trees

Mid-­tolerant trees

Mid-­tolerant trees

Tolerant trees

Tolerant trees

Tolerant trees

type. In the Clementsian paradigm, the series of stages at a given site is termed a sere and each stage a seral stage. There exist several potential misconceptions in relating the possible stages to real-­world succession. First, succession does not necessarily begin with stage 1, proceeding neatly through each successive stage unidirectionally. Second, stages do not typically proceed discretely one after another in relay fashion, but instead exhibit considerable overlap between them. Third, there is no predetermined time period for each stage to begin and end; depending on the site, one stage may occupy the site for a long time and seemingly terminate succession until site conditions change. In sum, real-­world succession typically has stages that are far less uniform and rigid than those listed in Table 17.3. The various life-­forms and developmental stages in succession may be characteristic of more than one stage, and indeed, many persist through many stages. Some mosses, for instance, may colonize a site early in succession and persist until the later stages characterized by tolerant trees. Tree seedlings often establish themselves very early in succession along with annuals and grasses or beneath shrubs, particularly in secondary succession when tree seed sources surround the ­disturbed area. Thus, all plants of each stage do not necessarily appear and die out abruptly at an appointed time; rather they may overlap in various sequences according to the life-­history traits of each species. For example, in the Appalachian Mountains many clonal shrubs and alder, rhododendron, and mountain laurel thickets persist for 20–­40 years or more, preventing the establishment of tree seedlings within them (Niering and Goodwin 1974; Damman 1975; Vose et al. 1994). The actual species composition of the different stages depends upon those species that have access to the site in question, either in terms of the relative location of seed sources or their ability to be dispersed to the site (Clements  1905; Abrams et  al.  1985; McCune and Allen  1985; ­McClanahan 1986; Matthews 1992; Fastie 1995; Buma et al. 2017). Thus, the actual plant communities present on a given site depend upon the available plants as well as upon the site itself. Such plants may be already on the site as buried seeds and propagules (roots, rhizomes) or may be ­transported by wind, water, or animals from adjacent or distant ecosystems. The importance of species availability and dispersal in succession was clearly borne out in a study by Buma et  al. (2017), who resampled Cooper’s original permanent plots in Glacier Bay, Alaska after 100 years of succession. The plots were located by Cooper at particular distances from a retreating glacier to represent time since glaciation, and thus various stages along the ­successional sequence. The first 50 years of succession seemed to confirm the basic hypotheses of plant succession that plant coverage, in general, and willow coverage, particular to this location, would be higher in “older” plots. The second 50 years of succession, however, suggested that the spatial location of seed

422  C h a p t e r 1 7   Forest Succession

sources and species dispersal into the plots began to become more important than time since glaciation. Moreover, the plots were highly variable in canopy coverage, species composition, and soil development. In particular, predicted late-­successional dominant species (Sitka spruce) did not establish in plots where willow arrived early and dominated. These results highlight that, in some cases, early species arrival has an enormous influence on successional pathways (Buma et al. 2017). Egler (1954) described the dichotomy of two models, “relay floristics” and “initial floristic composition” (IFC), that contrast the strategies of species as they colonize old fields. Relay floristics (Figure 17.3a) refers to the appearance and disappearance of stages of species at a site. This model suggests that one group of plants favors or facilitates the establishment of the next

Weeds

Grassland

Shrubland

Forest

Shrubland

Forest

Species abundance

Crop

Abandonment

(a)

(b)

Weeds

Grassland

Species abundance

Crop

Abandonment

Time

Time F I G U R E  1 7 .3   Models of succession from studies of abandoned agricultural fields (old-­field succession). (a) Relay floristics—­colonization of successive groups of species occurs in sequential phases. (b) Initial floristic composition (IFC) —­all species establish at or shortly after initiation of succession but still dominate sequentially in stages. Source: Egler (1954) / Springer Nature. Vegetation science concepts I. Initial floristic composition, a factor in old-­field vegetation development. Vegetatio 4:414, © 1954 by Uitgeverij Dr. W. Juink, Den Haag. Reprinted with kind permission from Kluwer Academic Publishers.

How Does Succession Work?  423

group in the sequence by modifying the environment. Some attribute this rigid scheme to ­Clements, but, in fact, he acknowledged the less discrete nature of the stages as they appeared. In 1905, Clements observed that many migrants appeared into a new, denuded, or greatly modified habitat to then be sorted into three groups by the establishment process: (i) those that arrive but are unable to germinate or grow and soon die, (ii) those that grow normally under the conditions present, and (iii) those that arrive but remain dormant through one or more of the earlier stages to appear at a later stage. Matthews (1992, p. 294) also observed that the relay floristics model, as shown here, does not realistically represent the lack of discreteness of the species groups in primary succession. In contrast to the relay floristics model, the IFC model (Figure 17.3b) suggests that all species are present at or shortly after succession begins, as acknowledged by Clements, and then occur as separate stages over time. The IFC model incorporates the idea that succession is a function of species’ life-­history traits rather than simply facilitation by those species arriving earliest. Perhaps the most questionable assumption in Egler’s IFC model is that all or most species are present at the start of succession and additional species arrivals do not occur. For example, Finegan (1984) argued that shrubs and small trees (cherries, juniper), some pines, and red maple are early colonizers in old fields in the eastern United States, but oaks and hickories colonize significantly later and are not present at the start. Finegan also demonstrates many other situations where succession does and does not resemble the IFC model. Matthews (1992) examined the IFC model in studies of primary succession on recently deglaciated terrain in Alaska and found an appreciable lag between colonization by pioneer species and that of loose groupings of later colonizers. These differential patterns of establishment draw the IFC model into question, as more sequences resemble the relay floristic model without the discreteness shown in Egler’s diagram (Figure 17.3a). In reality, there simply lacks a model of forest succession that is universally applicable to all sites and situations.

Secondary Succession  To some degree, secondary succession mimics primary succession in its basic principles. Most forest ecosystems regenerate rapidly following disturbances, and so secondary succession is often described in terms of the redevelopment of forest tree populations. Ecologists working in even vastly different forest regions have produced similar conceptual models of post-­disturbance forest succession featuring four phases or stages (Table  17.4). In European beech forests, Watt (1947) described “upgrading” and “downgrading” periods that he divided into four phases, of which the gap phase has had the most application elsewhere. In general, we recognize two basic stages, known as aggrading (also upgrading or building) versus senescing (also downgrading, degenerating, or declining). Among the best-­known models in North America are those featuring stand development (Oliver  1981; Table  17.4) and populations (Peet and ­Christensen 1980a, b; Table 17.4). The stand development model was developed in New England and the Pacific Northwest (Figure  17.4a). The population model applies particularly to the

Table 17.4  Stages of secondary succession. Forest maturation model (Franklin et al. 2018)

Stand development model (Oliver 1981)

Population model (Peet and Christensen 1980a, b)

Disturbance and legacy creation Preforest

Beechwood model (Watt 1947) Gap

Stand initiation

Establishment

Young forest

Stem exclusion

Thinning

Mature forest

Understory reinitiation

Transition

Old forest

Old growth

Steady state

Upgrading Bare

Forest canopy closure Oxalis Downgrading Rubus

(a) Species A Dominates

Species B Dominates

Species B, C, & D Dominate B C D D C

A

C D

A Time Since Disturbance

Disturbance Stand Initiation Stage

Stem Exclusion Stage

Understory Reinitiation Stage

Old Growth Stage

B B B

C C

C

A C C

A B C

C B

B

C

B

B

C C D

B

D

D

C

D

C

D

D

C

C

A

C D

C

C B

BB

D

D D C

C D

C

C

D

D

C

D

C

C

C

(b) Stand Initiation Stage

Stem Exclusion Stage

Understory Reinitiation Stage A

A

A

A

A

Disturbance

Accelerated Succession Stage

A

A

B

C

B

C

C

A A AA A A A

B

A

C BB C B C

A

C

A

B

C

C B B C B C B

F I G U R E  1 7 .4  Stages of stand development following major disturbance. (a) All trees forming the forest are already present in the stand or colonize soon after disturbance. However, the dominant overstory tree species changes as stem number decreases and vertical stratification of the species progresses. Height attained and duration of each stand varies with species, site conditions, and disturbances. Barring intervening disturbances, the “old growth” stage may be reached in less than 200 to over 500 years. Source: After Oliver (1981) / with permission of Elsevier. (b) Alternative diagram of disturbance-­mediated accelerated succession in a pioneer forest community. Species A is a pioneer tree, whereas B and C are later successional species. Disturbances may include logging, windthrow, ice storm, fire, and insect/disease epidemic. The old-­growth stage of species B and C is not shown. Source: After Abrams and Scott 1989. Part (a) is reprinted from Oliver, C. D. 1981. Forest development in North America following major disturbances. Forest Ecology and Management 3(3): 156, © 1981 with kind permission of Elsevier Science-­NL, Sara Burgerhartstrast 25, 1055 KV Amsterdam, The Netherlands. Part (b) after Abrams and Scott 1989. Reprinted with the permission of the Society of American Foresters.

How Does Succession Work?  425

Piedmont Plateau in the southeastern United States, where abandoned agricultural land reverted to pine-­mixed hardwood forest. Both models also apply widely to forests in other regions with the following characteristics: (i) developing following a major disturbance, (ii) having single or several age classes, and (iii) having stems that regenerate during a relatively short period following disturbance. These models are described in the following text. By contrast, Watt’s beechwood example was developed for forests on mesic sites and is most applicable in forest regions where small gaps are characteristic. Franklin et al. (2018) developed a model we term the forest maturation model that focuses heavily on the importance for biodiversity of early-­successional forests, old forests, and biological legacies (Table 17.4). It was developed for forests that experience infrequent, high-­ severity disturbances such as those in the Pacific Northwest but is broadly applicable to forests disturbed more frequently and/or less severely. All models of secondary succession are initiated by a disturbance that occurs at fine to broad scales. Such disturbances create marked structural changes in forests by killing trees, but their ecological function is to free resources such as light, nutrients, and soil moisture as well as to create a suite of biological legacies that will direct the rate and course of succession and forest redevelopment. Described as the disturbance and legacy creation event by Franklin et al. (2018), such an event is transitional between the undisturbed forest and the beginning of secondary succession (Table 17.4). In the stand initiation or establishment stage (also called reorganization), plants regenerate the area, and new individuals and species continue to appear for several years. Regenerating plants may develop from (i) newly dispersed or buried seeds (e.g., species A in Figure 17.4a), (ii) sprouts from stems and roots remaining from disturbed plants, and (iii) advanced ­regeneration (species B, C, D)—­saplings and seedlings of the previous forest understory that accelerate growth when released by disturbance. Stand initiation may last from 5 to 100 years depending on the type and severity of disturbance, species’ regeneration strategies, weather, herbivory, and many other factors (Oliver and Larson 1996). During this period, invasion continues until resources (particularly growing space) become limiting, and plant species diversity is often maximized. Stand initiation is analogous to the preforest stage (Franklin et al. 2018), which precedes the redevelopment of a closed-­canopy forest. At this point in succession, diversity of both plants and animals is relatively high, biological legacies are common, habitat niches and food sources are numerous, and food webs and trophic relationships are complex relative to later points in succession. Trees are typically present and are developing in the new forest, but do not yet dominate the ecosystem. Incorporating such a non-­forest stage into a conceptual model of forest succession is critical because the lack of trees results in an open, structurally complex stage that is often the most species-­rich of the successional process (Swanson et al. 2011). The ecological importance of early-­successional forests is often overlooked (Swanson et  al.  2011; King and Schlossberg 2014), in part because such forests are perceived as recently disturbed ecosystems in the midst of a “recovery” to a more recognizable, familiar forested condition. The preforest or stand initiation stage is followed by another important transitional stage, the forest canopy closure event, which creates sudden and dramatic changes in environmental conditions (Franklin et al. 2018). This event represents the reassertion of strong dominance of a site by trees, although the degree of dominance (and the speed at which canopy closure occurs) varies among forest types depending on the density of the canopy. The most obvious environmental changes that occur with canopy closure include the limitation of light at the forest floor (Chapter 6) and the moderation of the microclimate (Chapter 7). However, canopy closure also has significant influence on other ecosystem factors, such as competition and diversity. Ecosystem processes such as biomass accumulation and carbon storage are also affected, because canopy leaf area approaches its maximum at about the time of canopy closure, thus determining the rate of photosynthesis, carbon allocation, and growth (Chapter 18). For example, in lodgepole pine forests in Yellowstone National Park, Kashian et  al. (2013) found that tree density strongly controlled

426  C h a p t e r 1 7   Forest Succession

carbon storage only until canopy closure occurred, after which tree size and stand age became more important because leaf area had been maximized. Once canopy closure occurs, the processes that occur in the next stages are generally related to competition or biomass accumulation. In the stem exclusion or thinning stage (also called aggradation or young forest), virtually all growing space is utilized, crown closure occurs, overtopped seedlings die, and tree regeneration is limited or ceases (Figure 17.4a), although herbs and shrubs are typically still present in the understory. In addition, more vigorous trees grow larger and capture the space from weaker trees, and severe self-­thinning of the initial cohort takes place (Peet and Christensen 1987; Figure 13.11, Chapter 13). Figure 17.4a illustrates that the pioneer species A has been excluded and other species (B, C, D) assort themselves vertically depending on growth rate and biomass allocation. Many examples of development in single-­and multiple-­cohort stands are described in detail by Oliver and Larson (1996). In general, these competitive effects result in a forest with relatively uniform structure. Biological legacies (mostly snags and fallen logs) remain from the pre-­disturbance forest, and many trees killed due to thinning are also present. Plant species diversity is often minimized in this stage. In the understory reinitiation, transition, or mature forest stage (Figure  17.4a), the overstory begins to thin out or “break up” due to insects, pathogens, or other disturbances, thereby increasing the light reaching the understory. Gaps may begin to form in the canopy, releasing some of the smaller trees previously suppressed in the understory or subcanopy layers. Most overstory trees have maximized their height growth and crown spread. Tree regeneration begins, and individuals of advanced regeneration may live for a few years or decades and then die. Longevity of regeneration less than a meter in height in the understory varies substantially, for example, from 3  or 4 years in cherry bark oak to over 100 years for Pacific silver fir (Oliver and Larson  1996). Together, these features significantly increase the structural complexity of the stand relative to the stem exclusion stage. Pre-­disturbance biological legacies have significantly decayed and declined, but snags and logs from the post-­disturbance forest begin to accumulate during this stage. The old-­growth, steady-­state, or old forest stage (i.e., the late-­successional forest) develops as overstory trees age and increase in height and biomass (Figure 17.4a) while invasion and regeneration continue. Senescence, chronic insects and pathogens, and minor and major disturbances cause tree crowns to further disintegrate or entire trees to die and disintegrate singly and in groups. As a result, some advanced regeneration and understory trees are able to reach the overstory. Second-­ and even third-­generation trees grow into the canopy, and thus canopy trees begin to vary in their size. Very old forests may lack any representative trees from the original ­post-­disturbance cohort, such that a 500-­year-­old forest may lack trees older than 300 years! Structural diversity is therefore maximized in the old-­growth stage, both in terms of the abundance of diverse tree structures as well as the patchiness of structure (Franklin et al. 2018). For example, in old-­field Piedmont forests, hardwoods dominate over declining loblolly pines, and relatively even-­ aged patches resulting from previous gaps now undergo a miniature version of the three previous stages. Importantly, the old-­growth stage is defined not only by tree age and size but also by abundant biological legacies (such as snags and coarse woody debris) and ecological processes that are very different from those in earlier stages (Franklin et al. 1981; Harmon et al. 1986; Oliver and Larson 1996). The importance of the old-­growth stage for forest ecosystems has been emphasized by many researchers and is reviewed extensively by Franklin et al. (2018). Old-­growth forests are ecologically unique in North America in large part because twentieth-­century forest management deemed them unsuitable or “overmature” for timber production and other resources, resulting in their harvesting and conversion to earlier successional stages. The structural diversity provided by the old-­growth stage is critical for its provision of many habitats and ecological niches, as well as for providing critical living space for rare or endangered species. The old trees that persist within many old-­growth forests have ecological significance of their own, for their provision of habitats in

How Does Succession Work?  427

their cavities, large branches, and complex crowns; resistance to future disturbance because of their thick bark; and their contributions to genetic diversity (Franklin and Johnson  2012). Old trees that die continue to have ecological value as snags, logs, and coarse woody debris, all of which are common in the old-­growth stage. In general, the ecological importance of individual trees could be summarized as increasing slowly as they age and increase in complexity, and then gradually declining once they die (Franklin et al. 2018). There are many possible variations of the four-­stage models presented in Table  17.4 ­depending on the site, vegetation, and disturbance characteristics specific to the ecosystem. One alternative is illustrated in Figure 17.4b, where pioneer species A dominates in the first two stages only to be disturbed in the midst of the understory reinitiation stage, causing accelerated succession of species B and C. Thereafter, a B–­C old-­growth stage would be characterized either by areas lacking disturbance, or perhaps by minor or major disturbance which, in turn, might lead to replacement of B by C, C by B, or B and C by A! Logging or windthrow of the overstory when advanced regeneration is present (Figure 16.12, Chapter 16) is one example of the Figure 17.4b alternative, and other specific examples are given by Abrams and Scott (1989). Again, for both primary and secondary succession, complete, stage-­by-­stage successions rarely, if ever, occur in nature. Disturbances constantly disrupt the gradual internal changes that occur in an ecosystem and may set back, accelerate, or permanently change successional ­trajectories. Moreover, we emphasize that succession is ecosystem-­dependent, and a predictable sequence of vegetation stages can be expected within the context of most ecosystems. Disturbances often only temporarily alter the sequence unless they act to restart succession or permanently change the site, as with erosion, deposition, fire, and windstorm. In such cases, the disturbance may initiate an entirely new successional sequence. For example, on well-­drained sand soils, ­nutrients are retained primarily in the accumulated soil organic matter. A severe fire that destroys the stand and the organic matter may actually change soil conditions so drastically that a new succession is initiated (Damman 1975). Even a vegetation stage itself can change the site to the degree that there is a resulting change in succession. In Newfoundland, for example, Damman (1975) reported a relatively predictable succession following fire on most sites unless a Kalmia heath became established. If such a heath became firmly established after fire, it initiated soil changes leading to thin, iron-­pan formation, water logging, and peat bog formation, preventing the recovery of forest vegetation and creating a whole new successional sequence. Although succession does not always follow predicted patterns, disturbance-­or vegetation-­mediated changes in site conditions may strongly determine the trajectory of the new pattern.

SUCCESSIONAL CAUSES, MECHANISMS, AND MODELS This section describes some of the key characteristics, causes, mechanisms, and models of succession. These features gained attention as the role of disturbance in forest ecosystems was increasingly appreciated, as the number of ecosystems in which succession was observed increased, and in reaction to the classical model.

Key Characteristics and  Regeneration Strategies Succession functions as the i­ nteraction of species with different genetically determined life histories and regeneration strategies. There are likely limitless possibilities for differences between species in their regeneration requirements, allowing replacement of the plants of one generation by those of the next (Grubb 1977). Ecologists often attribute these traits and strategies to the species themselves, but they are intimately related to the site conditions that elicited them (Chapter 3). The characteristics of forest trees paramount to succession include: understory tolerance; seed production, size, and dispersal; growth rate and longevity; tree crown architecture; resistance to insects and pathogens;

428  C h a p t e r 1 7   Forest Succession

biomass production; allocation of carbohydrates; and nutrient requirements. It has been ­hypothesized that the competitive ability of species is regulated during succession by the changing ratio between nutrient availability and light (Tilman 1985, 1988), and that community composition should therefore change whenever the relative availability of the two limiting resources changes. Known as the resource-­ratio hypothesis, this idea has been criticized for its lack of general applicability (Glenn-­Lewin and van der Maarel 1992, p. 31) and because its predictions have not been robustly tested (Miller et al. 2005, 2007). The vital attributes concept (Noble and Slatyer 1977, 1980) identifies the characteristics of a species that are vital in determining its role in successional sequences. Such attributes include means of recovery from disturbance, ability to establish in the face of competition, and time needed to reach critical life-­history stages. The basic trait of differential longevity is also critical for some successional sequences (Egler 1954, 1977; Drury and Nisbet 1973). For example, short-­lived pioneer species may be replaced by long-­lived, late-­successional species that simply outlive them, and they are not regenerated even by subsequent periodic disturbances. Each of these characteristics (e.g., seed size, nutrient requirements, vital attributes, longevity, etc.) evolved in response to site-­specific ecosystem conditions (Chapter 3), and broad generalizations without an ecosystem context can be misleading. For example, in certain ecosystems small seed size and a short-­lived pioneer role are key factors in explaining successional pathways where such species are replaced by heavier-­seeded, more understory-­tolerant, long-­lived species. However, not all small-­seeded species are short-­lived pioneers. For example, hemlocks (Tsuga) and cedars (Thuja) are small-­seeded, very tolerant, long-­lived groups that persist in certain kinds of ecosystems whose seedbed conditions (e.g., moist to wet sites; bare mineral soil) and disturbance regimes (e.g., fire, erosion, and deposition) favor them. By contrast, walnuts are heavy-­seeded but intolerant and fast-­growing; likewise, their vital attributes evolved and are expressed in specific kinds of ecosystems wherever they persist today in North America and Asia. In summary, the key life-­history attributes and regeneration strategies are inherent parts of many overlapping, sequential, ecosystem-­specific processes, such as regeneration (Chapter 4), competition and mutualisms (Chapter 13), and biomass accumulation and growth (Chapter 18). These processes, involving both site and plants, explain succession on an ecosystem-­by-­ecosystem basis. Walker and Chapin (1987) describe this process model for primary succession in Alaska, and details are available in a number of sources (Walker and Chapin 1986; Walker et al. 1986; Matthews 1992).

Availability and  Arrival Sequence of  Species Successional trajectories are strongly

affected by the availability of species and their arrival at a disturbed site. The timing and magnitude of species invasion vary depending on their occurrence in adjacent ecosystems, the spatial position and distance of the disturbed site in relation to seed sources, the site conditions and disturbance history of the disturbed area, and species life-­history traits. Differential arrival often explains why different successional pathways occur on similar sites. In secondary successions following windstorms or fires, surviving understory plants, clonal plants, and those germinating from the soil seed bank have immediate access to the site. For example, the differential arrival of species largely determined three pathways of primary succession in Alaska where a 220-­year glacial retreat progressively exposed parent material for colonization (Fastie  1995). The variability of species arrival was related to the distance from each site to the closest seed source of Sitka spruce at the time of deglaciation. In secondary succession following logging and fire in the Great Lakes region, some eastern white pine seed trees were able to survive, and the variable successional pathways on a given area today demonstrate the influence of these remnant seed sources (Palik and Pregitzer 1994).

Facilitation, Tolerance, and  Inhibition Connell and Slatyer (1977) proposed three alternative models of succession based on the mechanisms of facilitation, tolerance, and inhibition

How Does Succession Work?  429

A disturbance opens up a relatively large space

FACILITATION Individuals of any species in the succession could establish and exist as adults under the prevailing conditions

Only certain ‘pioneer’ species are capable of becoming established in the open space

TOLERANCE Modification of the environment by early occupants makes it more suitable for recruitment of ‘late-successional’ species

Modification of the environment by early occupants has little or no effect on subsequent recruitment of ‘latesuccessional’ species

In time, earlier species are eliminated through competition for resources with established ‘latesuccessional’ adults

The sequence continues until the current resident species no longer facilitate the invasion and growth of other species and/or no species exists that can invade and grow in the presence of the resident

INHIBITION Modification of the environment by early occupants makes it less suitable for recruitment of ‘late-successional’ species

As long as earlier colonists persist undamaged and/or continue to regenerate vegetatively, they exclude or suppress subsequent colonists of all species

If external stresses are present, early colonists may be damaged or killed and be replaced by species which are more resistant

F I G U R E  1 7 .5   Three autogenic mechanisms—­facilitation, tolerance, and inhibition—­that determine successional trends. Source: After Connell and Slatyer (1977) / University of Chicago Press, as modified by Begon et al. 1990. Reprinted with permission from American Naturalist, © 1977 by the University of Chicago. Modification reprinted with permission of Blackwell Science.

(Figure  17.5). Notably, many ecologists have found these models restrictive (Huston and Smith  1987) and oversimplified (Christensen and Peet  1981) because each was developed to ­represent a single, opposing, alternative pathway (Pickett et al. 1987). In reality, nature features an enormous variety of pathways and mechanisms that may simultaneously influence successional change in complex ways. Facilitation is analogous to the model of relay floristics (Figure  17.3a), whereby early ­successional species modify the environment to favor later successional species. In addition to the amelioration of environmental stress, facilitation may operate by increasing the availability of resources and enhancing colonization ability (Pickett et  al.  1987). Thus, a species with one

430  C h a p t e r 1 7   Forest Succession

combination of life-­history traits may establish and modify environmental conditions in a manner that favors individuals with other combinations (Huston and Smith 1987). Facilitation in this sense has been reported in many instances (e.g., Cooper 1923, 1931, 1939; Jenny 1941; Christensen and Peet 1981; Walker and Chapin 1986; Wood and del Moral 1987; Matthews 1992; Fastie 1995; Baumeister and Callaway 2006; Calder and St. Clair 2012). Chapin et al. (1994) investigated facilitation during succession following deglaciation at Glacier Bay, Alaska (Figure 17.6). Life-­history traits determined the pattern of succession, and initial site conditions and facilitation (where present) influenced the rate of succession as well as the composition and productivity of the late-­successional community. Facilitative effects were strongly related to the increase in nitrogen that is enhanced by all plants but especially by alder. Inhibition, where species established early inhibit subsequent colonization of other species by pre-­empting space and other resources, was also an important process at Glacier Bay (Figure 17.6). A common example of inhibition is when shrubs resist colonization by trees (Niering and Egler 1955; Webb et al. 1972; Niering and Goodwin 1974; Damman 1975; Chapin et al. 1994). The mechanism of tolerance relates to the ability of a species to tolerate low resource levels. In such situations, the initial species modify the environment, but this change has little or no effect on subsequent recruitment and growth of later successional species (Figure  17.5). Instead, ­late-­successional species replace early-­successional species either by active competition or by simply being longer-­lived. Active and passive concepts of tolerance are discussed by Pickett et al. (1987). Importantly, facilitation, inhibition, and tolerance describe relative, rather than absolute, processes (Huston and Smith  1987). The three processes almost always occur simultaneously with varying degrees of importance and are likely to occur in most primary and secondary ­successions. Various combinations of the three processes are reported for primary succession on glacier forelands (Matthews  1992; Chapin et  al.  1994; Fastie  1995), in old-­field succession in New York (Gill and Marks 1991), and on the North Carolina Piedmont. On Alaskan floodplains, STAGE: FACILITATIVE EFFECTS:

INHIBITORY EFFECTS:

Pioneer Survivorship

Germination (weak)

Dryas N (weak) Growth (weak)

Germination Survivorship Seed predation and mortality

Alder SOM N Mycorrhizae Growth

Germination Survivorship Seed predation and mortality Root competition Light competition

Spruce Germination

Germination Survivorship Seed predation and mortality Root competition Light competition N

F I G U R E  1 7 .6   Summary of facilitative and inhibitory effects (weak effects noted) for four successional stages on establishment and growth of spruce seedlings at Glacier Bay, Alaska, as determined from field observations, field experiments, and greenhouse studies. SOM = Soil organic matter. Source: After Chapin et al. (1994) / John Wiley & Sons. Reprinted with permission of the Ecological Society of America.

Change in Ecosystems  431

many different processes and factors (e.g., life history, chance, facilitation, competition, herbivory, etc.) affect how alder and spruce interact during succession, and no single successional process or model adequately describes successional change in these ecosystems (Walker and Chapin 1987).

CHANGE IN ECOSYSTEMS

Ecosystems are constantly changing over time and space: from night to day, seasonally, and year to year as climate and soil change, and as the cycle of organisms’ activity changes. As such, a regional or local ecosystem never can and never does reach a complete balance or permanence except for something arbitrarily chosen by humans for their own understanding.

END POINT OF SUCCESSION? Forest succession has been recognized from the earliest days of natural history study, but whether there is a last stage of succession has been debated for decades. Cowles observed in 1901: “The condition of equilibrium is never reached, and when we say that there is an approach to the mesophytic forest, we speak only roughly and approximately. As a matter of fact, we have a variable approaching a variable rather than a constant.” On the other hand, Clements defined the term “climax” relatively rigidly as the end point of succession and used it as the driver of his vegetation classification. Over decades many descriptors were added to the term to improve it conceptually, such as “polyclimax,” “polyclimatic climax,” “site climax,” “climax pattern,” or position of relative stability. Regardless of these modifications, the term implies a terminus of succession, and epitomizes succession within a vegetation context rather than an ecosystem context. In fact, the term “climax” is inappropriate within an ecosystem context. It is more useful to recognize pioneer or early-­successional, mid-­successional, and late-­successional species or groups for position in characterizing vegetation along the temporal scale for site-­specific ecosystems. These terms (early, mid, late) are not absolute but are relative to the specific ecosystem in the landscape. In any case, climax terminology permeates the literature such that it is useful to examine how the terms are used. Macroclimate is the dominant factor forming communities in Clements’ monoclimax theory, whereas other factors (physiography, soil, fire, biota) are of secondary importance. When other factors are of equal importance, a different climax community is possible or even likely for different physiographic settings, soil types, and disturbance factors. This is the ­polyclimax theory (Tansley 1939), which holds that biotic succession will move toward a climax for any combination of environment and organisms, but that the specific nature of the climax will vary with environmental factors and biotic conditions. Cowles (1899, 1901), Moss (1913), and especially Nichols (1923) contributed heavily to polyclimax theory. In fact, Nichols argued for a ­different physiographic climax on each site that differed more or less from the regional climatic climax. For example, in his classic studies of plant succession on the sand dunes of lower Lake Michigan, Cowles at first thought that succession on all sand dune sites would eventually reach the same forest stage of mixed mesophytic hardwoods. In a later study of the same area, Olson (1958) concluded that the drier, lower fertility, and more exposed sites would never support such a community but would instead be occupied more or less permanently with a black oak–­blueberry type. Furthermore, Olson suggested that although vegetation changes would continue to become progressively slower with time, but they would never reach complete stability. Over the years, the accepted concept of climax became broader, less rigid, and more closely approximated the polyclimax approach. Tansley (1939) was instrumental in broadening the concept by adopting a dynamic viewpoint in recognizing that “positions of relative equilibrium are reached in which the conditions and composition of the vegetation remain approximately constant for a longer or shorter time.”

432  C h a p t e r 1 7   Forest Succession

Polyclimax theory is also mired in its own terminology: climatic, edaphic, topographic, topoedaphic, fire, zootic, salt, etc. (see descriptions in Oosting 1956 and Daubenmire 1968). Many of these terms approximate or are merely more specific distinctions of Clements’ climaxes. For example, in the polyclimax approach the Clementsian climax was renamed “climatic climax;” Clements’ subclimax may be either an “edaphic” or “topographic climax;” Clements’ disclimax becomes a “fire” or a “zootic climax.” In addition, like Clements, only one climatic climax was ­recognized. A polyclimatic climax theory was later developed (see Meeker and Merkel 1984) that recognized more than one climatic climax for a macroclimatic region. Whittaker (1953, 1975) recognized a gradational climax pattern of vegetation corresponding to environmental gradients—­a continuum of climaxes. Whittaker (1953) wrote that climax ­composition was determined by all “factors” of the mature ecosystem—­properties of each of the species involved, climate, soil and other aspects of site, biotic interrelations, floristic and faunistic availability, chances of dispersal and interaction, etc. There is no absolute climax for any area, and climax composition has meaning only relative to position along environmental gradients and to other factors.

The “climax pattern” continuum has clear problems in application to management, and so Dyksterhuis (1949, 1958) proposed the use of site units to characterize succession and climax for range classification. This approach is termed site climax by Meeker and Merkel (1984). Other viewpoints and details of the climax concept have been presented by many authors ­(Phillips 1931, 1934–­35; Tansley 1935; Clements 1936; Cain 1939; Whittaker 1953, 1975; Daubenmire 1968; Langford and Buell 1969; White 1979; Matthews 1992).

SUCCESSION AS AN ECOSYSTEM PROCESS

A consistent thread throughout the abovementioned discussion is that a general model of succession based on species or communities alone is unrewarding. It is somewhat useful to develop a framework of stages based on the development of stands or populations, but specific site conditions are the primary determinant of the duration, distinctness, and predictability of those stages. Species’ life-­history traits are inseparably linked to the sites and disturbance regimes to which they are adapted. An ecosystem approach to forest succession provides a functional and geographical framework for understanding these multiple mechanisms and processes that drive succession. An ecosystem approach to forest succession requires that the first-­order focus is the regional and local ecosystem rather than component parts, such as vegetation or soil. Succession is a multidimensional process with many factors and processes occurring at multiple scales (Table 17.2) that can be assessed for specific regional and local ecosystems. A “top-­down” approach that begins with an understanding of the geographic hierarchy of regional and local ecosystems, their physical factors, biotic interrelationships, and past histories (Chapter 2) is often most appropriate in this regard. Each regional ecosystem will exhibit unique successional pathways and ­mechanisms, but their character and pattern will be further different among climatic–­ physiographic regions. For example, forest succession in the southern Appalachian Mountains is different than that of the Yellowstone Plateau, the slopes of the Sierra Nevada, or the southeastern Coastal Plain. Likewise, the four ecosystem regions in Michigan (Figure 2.10, Chapter 2) exhibit different patterns of succession due to differences in climate, physiography, soil, and vegetation. Within a given region in Michigan, ecosystem districts (Figure  2.10, Chapter  2) have their own characteristic successional patterns for local ecosystems occurring as sand plains, acid swamps, calcareous swamps, rocky ridges, or moist and fertile lower slopes. Variation in succession may even be expected for local ecosystems (Simpson 1990) due to past disturbance or management (logging, drainage), disturbance frequency and severity, seed source availability, and microsite variation, among other

Succession as an Ecosystem Process  433 Paludification Thermal erosion No fire

Balsam fir

Fire

Black spruce

Fire

Wet, mesic, and dry sites

es

Me

s

es

Paper birch

n

s it Dr y i la Flo o d p it ys Dr

es

Aspen

s it es

No fire

s it

White spruce

Floodplain sites-no fire

Drying Bog filling

?

ic

Long period no fire

Bog

Long period no fire (Eastern Canada)

Fire

Jack pine

F I G U R E  1 7 .7   General example of successional pathways involving black spruce on different ecosystems, defined by their site conditions, of the boreal forest in Alaska and Canada. Source: After Viereck (1983) / John Wiley & Sons. Reprinted with permission from SCOPE 18, The Effects of Fire in Northern Circumpolar Ecosystems, edited by Wein, R. W. and D. A. MacLean 1983 © SCOPE, John Wiley & Sons, Ltd., Chichester, UK.

factors. The factors that obscure successional trends are significantly reduced by specifically defining the geographic framework and working with more or less homogeneous site conditions. A general example of multiple, ecosystem-­specific successional pathways is presented in Figure 17.7 based on the site characteristics and disturbance regimes of the boreal forest of Alaska and Canada. On many wet, mesic, and dry sites, black spruce replaces itself with or without fire (Viereck 1983). In the southern regions of the boreal forest, colonization of burned black spruce forest by birch and aspen is more common than in the north. Long fire-­free periods in eastern Canada enable balsam fir to replace black spruce (Figure  17.7). Many other ecosystem-­specific pathways are described by Viereck (1983). A specific example of an ecosystem approach to succession comes from research on biomass accumulation and nutrient cycling in northwestern lower Michigan (Figure 17.8; Zak et al. 1986; Host et al. 1987; Host et al. 1988). Site-­specific ecosystems were first distinguished at two levels: landforms (outwash plain, ice-­contact drift, and moraine) and ecosystem types within landforms (named by overstory and ground-­cover species). Successional trends for landforms and ecosystem types were then estimated based on an understanding of site conditions and life-­history traits of the species in the understory and ground-­cover layers (Figure 17.8). Importantly, the oak cover types failed to distinguish the landforms and ecosystem types as well as the predicted successional pathways, suggesting that an understanding of entire ecosystems—­not just communities defined by dominant species—­is necessary to ascertain successional pathways. Moreover, different ­pathways were predicted for ecosystem types within each landform, and these pathways, based on multiple ecological factors, are the best estimate of successional trends available for making

434  C h a p t e r 1 7   Forest Succession Landform (Surface configuration)

Outwash Plain (Level to gently undulating)

Ice-contact Terrain (Hilly)

Parent Materials

Well-Sorted Medium Sand

White OakBlack OakWell-Sorted Northern Red Oak Medium Sand With Loamy Subsurface Textural Northern Bands Red Oak Medium Sand With Clay Loam Subsurface Textural Bands Fine or Loamy Sands

Moraine (Hilly)

Society of American Foresters Cover Type

Fine or Loamy Sands With Sandy Clay Loam Subsurface Textural Bands

Ecosystem

Black OakWhite Oak/ Vaccinium

Current Dominant Understory White Pine

Apparent Future Compositional Changes Loss of Oaks in Overstory Conversion to White Pine ?

White Oak (Suppressed)

Persistence of White Oak Formation of Oak-Sadge Savanna

Mixed Oak/ Trientalis

Red Maple

Loss of Oaks in Overstory Conversion to Red Maple

Mixed Oak/ Viburnum

Red Maple

Loss of Oaks in Overstory Conversion to Red Maple or Northern Hardwoods

Sugar MapleRed Oak/ Maianthamum

Sugar Maple

Loss of Red Oak in Overstory Replacement by Sugar Maple or Sugar Maple and Beech

Sugar Maple Sugar Maple-Basswood/ Sugar Maple Osmorhiza Basswood

Loss of Basswood in Overstory Replacement by Sugar Maple or Sugar Maple and Beech

Sugar Maple

F I G U R E  1 7 .8   Estimated future compositional change associated with two levels of landscape ecosystems (landforms and ecosystem types) in northwestern lower Michigan. Source: After Host et al. (1987) / Oxford University Press. Reprinted with the permission of the Society of American Foresters. Site conditions, overstory cover type (Eyre 1980), and current dominant understory species are also shown.

management decisions. Being specific to ecosystems that recur throughout the landscape allows the proposed trends to be monitored and tested over time. Such an approach is applicable for improved management, conservation, and restoration of ecosystems on many or most public and large private lands in North America.

EXAMPLES OF FOREST SUCCESSION PRIMARY SUCCESSION ON RECENTLY DEGLACIATED TERRAIN Primary succession is perhaps best studied on the land newly formed in front of a glacier, and such detailed studies of the glacier foreland have been underway since the late 1800s. Matthews (1992) provides a comprehensive review of primary succession on recently deglaciated terrain, examining physical and ecological processes and successional stages. From his geoecology viewpoint, succession in front of the retreating glacier is a process of the developing landscape when all parts of the ecosystem develop simultaneously. The most important visible features of the landscape in the glacier foreland are the landforms and their associated parent materials, and the developing vegetation is a function of the dominant factor terrain age. Other factors are subordinate: climate, physiography, parent material, organisms, as well as physical site change and disturbance. ­Matthews found that vegetation chronosequences are generally limited by differences among sites, both the initial environment and their subsequent environmental histories, but their careful use can provide approximations of succession at particular sites. For example, the chronosequence at Glacier Bay, Alaska is a series of eight stages that occur along a continuum: I. Early pioneer stage: mountain avens (Dryas drummondii) and prostrate willows, mosses, and herbs. II. Dryas mat: coalescing mats of Dryas. III. Late pioneer stage (after 15–­20 years): young Sitka alder and black cottonwood shorter than 2–­3 m.

Examples of Forest Succession  435

IV. Open thicket stage (after 20–­25 years): clumps of alder. V. Closed thicket stage (after 25–­30 years): dense alder with individual erect specimens of balsam poplar, shrubby willows, and Sitka spruce. VI. Balsam poplar line stage (after 30–­40 years) where poplars reach a height of 4–­5 m and form a canopy and visible line above the alder. VII. Spruce forest stage (at 75–­90 years) when spruce replaces alder. VIII. Spruce–­western hemlock stage (speculative because the forest may remain dominated by spruce even after 200 years). The horizontal and vertical structure of vegetation on the glacier foreland of the Grand ­ lacier d’Aletsch in the Swiss Alps has long been studied with permanent plots (Figure 17.9). The G sequence of herbs to shrubs to trees typically reflects the sequence of plant size, longevity, and competitive power, but trees may colonize much earlier in the sequence depending on seedbed conditions and seed availability. In the Canadian Rocky Mountains, establishment time ranges from 10 to 15 years to 80 years (Luckman 1986). Matthews’ review of primary succession in the glacier foreland reveals that succession may proceed in trajectories that are parallel, but also in those that are strongly divergent or convergent. Generally, divergent pathways are most clear during early stages and are associated with relatively severe physical environments; convergence is favored by relatively strong biotic controls. Biomass and species richness tend to increase during primary succession, although patterns of diversity vary considerably from site to site (Matthews 1992). Matthews notes that most of the general successional models (as previously described) are not well applied to recently deglaciated terrain. Such land instead is better described by a geoecological model that emphasizes physical environmental processes as well as biological processes. In addition, models of succession as a landscape process must incorporate spatial variation in site

150

% cover

F I G U R E  17 .9   Coverage of bare ground and five vegetative strata on recently deglaciated terrain, which is on terrain of increasing age, Grand Glacier d’Aletsch, Swiss Alps. Source: After Lüdi (1945). Reprinted with permission of the 17, Stiftung Rübel, Zurich, Switzerland.

100

50

0 4–6

25

45

70

85

Terrain age (years)

Bare ground

Dwarf shrubs (0.5 m)

Herbs

Trees (>2 m)

436  C h a p t e r 1 7   Forest Succession

characteristics through time. In sum, Matthews strongly emphasizes the properties (landforms and glacial sediments, climate, soil) and processes (erosion by water; frost weathering, heaving, and sorting; solifluction; wind deposition and erosion; and soil development) of the physical landscape. Matthews’ summary points concerning succession on recently deglaciated terrain provide real-­world insights into the mechanisms of succession discussed previously in this chapter: Allogenic and autogenic processes interact during succession and are often inseparable. Establishment is more important than migration (species arrival) in determining species composition in the harsh environment of glacial forelands. Here, it is likely that pioneers require some special physical, chemical, or microbial conditions as well as lack of competition. Multiple processes—­including physical as well as biological processes—­are characteristic in replacement of pioneers by later colonizers. Pioneer species may decline more due to physiochemical changes associated with leaching (nutrient depletion, balance of available nutrients, toxicity associated with pH changes) and other physical changes than to the rise of species of the next successional stage. The action of allogenic factors (physical processes and disturbances) as compared with autogenic action of the plants (the importance of life-­history traits) influences both the nature and rate of succession in the harsh physical environment of the glacier forelands. In favorable environments (high in resources; low in stresses), there is a rapid increase in the importance of autogenic processes through time (Figure 17.10, see “1”). This increase may reflect the rapid accumulation of biomass that occurs as soil water and nutrients become more favorable in a favorable climate or a relatively undisturbed site. By contrast, allogenesis, as a determinant of vegetational change, is reduced sharply (1′) as the intensity of physical processes and disturbances decrease plant biomass and vegetative control of the environment increases. As the environment becomes less severe, biological production is constrained, allogenic processes become relatively more important in any successional stage, and the control of succession by allogenic processes becomes longer. The relative importance of allogenesis in Figure 17.10 (curves 2′ and 3′) is highest in more severe environments. That is, on very severe sites, plants themselves modify the environment to a lesser extent than on favorable sites. This relationship can be applied at both the broad scale of regional

100

Relative importance [%]

3´ 2´ 1´ 50 1 2 3 0 Successional time

F I G U R E   17. 10  Relative importance of autogenic (1, 2, 3) and allogenic (1′, 2′, 3′) processes under increasing environmental severity (1 and 1′ with solid line = least severe; 2 and 2′ with dashed line = moderately severe; 3 and 3′ with dotted line = most severe) in the context of recently deglaciated terrain. The model assumes species/community context with autogenic processes relating only to vegetational processes; allogenic processes are a combination of physical processes and disturbance. Source: After Matthews (1992) / Cambridge University Press. Reprinted from The Ecology of Recently Deglaciated Terrain, © 1992 by Cambridge University Press, Reprinted with the permission of Cambridge University Press.

Examples of Forest Succession  437

ecosystems (increasing latitude, for example, boreal forest, tundra, polar desert) and the local scale within the different landforms of the glacier foreland. As described in the earlier text, however, autogenic and allogenic processes are not mutually exclusive, and their relationship to site severity applies only to the extent that these processes can be distinguished. It appears likely that they are most distinct in the harshest environments.

SUCCESSION FOLLOWING THE ERUPTION OF MOUNT ST. HELENS On May 8, 1980, Mount St. Helens erupted in the Cascade Mountains of Washington in the Pacific Northwest. The eruption produced a catastrophic disturbance to approximately 600 km2 of ­old-­growth and planted forests dominated by western hemlock, Douglas-­fir, and Pacific silver fir. The disturbance created by the eruption was extremely heterogeneous due to the diversity of geological processes that occurred. These processes included pyroclastic flows (a landslide and flows of hot gases and pumice fragments) that eliminated ecosystems and landforms over about 60 km2, a lateral blast zone that blew down or scorched forests over 500 km2, and the deposition of more than 5 cm of ash over 1000 km2 (Dale et  al.  2005a). Whether organisms survived the eruption depended heavily on which of these zones they resided in prior to the disturbance, with little survival in the area of pyroclastic flows, some survival of underground plant parts and burrowing animals or organisms protected by snow in the blast zone, and greater survival of taller organisms in the ashfall zone. The presence of snow was a major factor in organism survival and the subsequent rate and trajectories of succession (Figure 17.11). Much like we observe in areas of severe fires or windstorms, the heterogeneity of disturbance intensity resulting from the eruption had important consequences for post-­eruption succession. Unlike fires or windstorms, the pyroclastic zone generally buried or scoured away any organic material toward its terminus (Dale et al. 2005b), and succession there was therefore mostly primary (Dale and Denton 2018). However, secondary succession characterized the blast zone and the ashfall areas because there was survival of some to many organisms. Remarkably, some individuals of most plant species present before the eruption survived the disturbance in refuges of various forms (Del Moral  1983; Halpern and Harmon  1983; Adams et  al.  1987; Zobel and Antos  1992), and survivors represented most life-­forms and successional stages. For example, snowbanks that had persisted into early May of 1980 protected the tree seedlings and shrubs that they had buried, allowing them to immediately resume growth upon snowmelt (Franklin and MacMahon  2000). Together with populations dispersed into the disturbed area, these survivors facilitated and shaped secondary succession (Dale and Crisafulli 2018), and succession occurred most rapidly from areas with surviving organisms (Franklin and MacMahon 2000). Succession following the eruption of Mount St. Helens has been extensively studied, and a series of important lessons have emerged. The concept of biological legacies emerged from the aftermath of the Mount St. Helens eruption, and there they were found to be extremely important in directing the rates and trajectories of succession (Dale et  al.  2005a). Pre-­eruption legacies included dead organic material (mainly trees blown over or left standing but scorched in the blast zone), but also organisms that survived the blast but died before they were able to reproduce (Dale and Crisafulli 2018). Ecologists had largely ignored the importance of such legacies for ecosystem structure prior to 1980, but they were found to be important in shaping post-­eruption erosion and deposition processes as well as supporting many organisms with protective cover, food, habitat, and nutrients (Franklin and MacMahon 2000). An additional important lesson provided by succession at Mount St. Helens is that succession is extremely complex and does not always follow ecological theory. Heterogeneity in organism survival and in the distribution of biological legacies created many different points of initiation of succession across the landscape that resulted in many different pathways of succession that did not necessarily converge. In some cases, adjacent sites undergoing primary succession supported very

F I G U R E  1 7 .1 1  Top: Photo chronosequence taken from 1983 to 2004 showing changes in plant community development in the blowdown zone of Mount St. Helens at a site that was snow-­covered at the time of the eruption. Snow cover facilitated the survival of numerous plants, including conifer saplings. Bottom: Photo chronosequence showing changes in plant community development from 1983 to 2016 in the blowdown zone at a site that was not snow-­covered at the time of the eruption. Source: From Dale and Crisafulli (2018); photos by Peter Frenzen, USDA Forest Service.

Examples of Forest Succession  439

different plant communities, emphasizing the importance of random or chance events in the successional process (Dale and Crisafulli 2018). Specific to forests, several late-­successional coniferous species (e.g., western hemlock, Pacific silver fir) were found to colonize the primary succession sequence on the debris flow area within the first 3 years after the eruption, although seedling density greatly decreased by 5 years post-­disturbance (Dale et al. 2005b). These results highlight the differences between ecological theory and the reality of ongoing succession in real-­world ecosystems following complex disturbance events. Other succession research completed at Mount St. Helens through 2015 is summarized in detail by Crisafulli and Dale (2018).

SECONDARY SUCCESSION FOLLOWING FIRE IN PONDEROSA PINE FORESTS OF WESTERN MONTANA Fire leads to ecosystem change throughout forests of the western United States (Chapters  10 and 16). Fischer and Bradley (1987) described the major role of fire in determining successional pathways in grasslands and forests in western Montana. For example, ponderosa pine of the submontane zone in western Montana competes with grasses in dry ecosystems at low elevations and with Douglas-­fir in the cooler and moister ecosystems of the montane zone. In this submontane zone, fire acts to (i) maintain grasslands, (ii) maintain open ponderosa pine stands, and (iii) encourage ponderosa pine regeneration. Frequent fires maintain the grassland community by killing pine seedlings (stage 1; Figure 17.12a). Over occasional long fire-­free periods, ponderosa pine seedlings become established gradually as the result of a seedbed-­preparing fire (stage 2). The seedlings develop into saplings in the absence of further burning, and fires during this period may kill young trees (stage 3) or thin them (stage 4). In time, light surface fires produce an open stand of mature trees (stage 5) that may accumulate enough fuel to produce a stand-­replacing fire (stage 6) which is followed again by grassland (Figure 17.12a). Multiple successional pathways are evident in these forests using a starting point of open, park-­like, old-­growth ponderosa pine forest (Figure  17.12b). The interaction or absence of fire (timing and severity) and the life-­history traits of pine and grasses result in various states of pine stands (e.g., A, B2, C3, E2, etc.) and pathways. For example, starting with an open, park-­like, ­old-­growth pine stand (A), frequent fires (pathway 33) maintain grassland (K), but moderate-­ severity fire (pathway 5) leads to a closed-­canopy, multistoried pine stand (stage D1). A dense stand (El) occurs in the absence of fires and is either maintained by low-­severity fire (pathway 7) or succeeds to grassland (F) after severe fire (pathway 9). This example illustrates the complexity of succession and the lack of any end point of the successional process. Similar examples of fire ecology for the diverse range of ecosystems of eastern Idaho and western Wyoming are also available (Bradley et al. 1992). Quantitative succession models are beyond the scope of our treatment, but many simulation models have been developed for succession (Huston et  al.  1988; Botkin  1993). A mechanistic, ­ecological process model simulating fire succession of coniferous forests of the northern Rocky Mountains is also available (Keane et al. 1996), as is a commonly used forest succession simulator with versions specialized for individual forest regions of the United States (Crookston and Dixon  2005). Recent model development also includes a spatially explicit landscape-­scale succession model capable of incorporating climate change and additional disturbances (Scheller et al. 2007).

SECONDARY SUCCESSION AND GAP DYNAMICS Forest trees may establish and successfully reach the overstory canopy via three modes (Veblen 1992): (i) following severe disturbance, (ii) by establishing in overstory gaps, or (iii) rarely, by continuous regeneration in a disturbance-­free understory and recruitment into an overstory without discrete gaps. Understory intolerant species typically reach the overstory following severe

440  C h a p t e r 1 7   Forest Succession

disturbances, whereas mid-­tolerant and tolerant species often utilize relatively small treefall gaps. Tolerant species may recruit without disturbance if the canopy is not too dense, but they usually establish in a disturbance-­free understory (Lorimer et al. 1988; Busing 1994) and then require at least periodic episodes of overstory crown thinning. We examine regeneration in canopy gaps in this section. When an opening or gap occurs in the forest overstory, advanced regeneration and new stems colonizing the gap from the understory form a patch of vegetation that responds to the newly changed environmental conditions—­leading to a type of secondary succession termed gap dynamics ­(Pickett and White 1985). Watt (1925, 1947) described four phases (Table 17.4) in the development of pure European beech forests in England. One of these four phases was the “gap phase” that occurred in old-­growth beech stands. Gaps are initiated by major and minor

(a)

Firemaintained grassland

1 6 Stand-destroying fire (rarely occurs)

2 Prepare seedbed

Grass/forb

Seedlings PIPO Crowded, twostoried stand (rarely occurs) PIPO

Pole stand PIPO Firemaintained open, parklike stand PIPO

3 Return to grassland

Saplings PIPO

4 Thinning fire

5 Generalized forest succession Cool fire

Severe fire F I G U R E  1 7 .1 2  The role of fire in forest succession in warm and dry ponderosa pine (PIPO) ecosystems. (a) Generalized diagram of succession illustrating situations where grassland is maintained or where ponderosa pine is favored by a less-­frequent fire regime. (b) Specific diagram of multiple successional pathways starting from an open, park-­like old growth ponderosa pine stand. Numbers (1, 2, etc.) indicate successional pathways; letters (A, B) indicate states of pine forest or grassland. Source: After Fischer and Bradley (1987) / U.S. Department of Agriculture / Public Domain.

Examples of Forest Succession  441

(b) 8

D1 Closed canopy multistoried PIPO stand

E1 Crowded PIPO stand ; structure varies

2

4

7

Low

Low

Severe

9

Low

6

10 frequent fires

re ve Se

C1 Scattered PIPO overstory; dense PIPO understory

y An

t os M

33

B1 Open PIPO stand; unevenaged regen.

e at er od M

A Open, park-like, old growth PIPO

e at er od M

re ve Se

1

5

3

Most

Severe

re ve Se

12

11

Severe

17 K Firemaintained grassland

14

C2 Scattered PIPO overstory; dense PIPO understory

B2 Open PIPO stand; evenaged regen.

D2 Stagnant PIPO pole stand

13

Low

E2 Broken, decadent PIPO stand

15

Low

LowModerate C3 Scattered PIPO overstory; sparse PIPO understory

18

Low

D3 Open PIPO pole stand

19

Low

E3 Open park-like PIPO stand

21

Severe

16

F Grass

e ver Se

20

Go to A

Any

Succession in absence of fire Response to fire

Low Cool or light surface fire Mod. Fire of intermediate (moderate) severity Severe Hot, stand-destroying fire

F I G U R E  17 .1 2   (Continued)

disturbances, and most emphasis has been placed on severe disturbances such as windstorms, fires, flooding, and landslides (Chapter 16). However, fine-­scale or tree-­fall gaps also significantly affect the structure and composition of mesic forests around the world, especially those dominated by understory tolerant trees. Small canopy gaps are characteristic of many mesic forests of eastern and western North America. Larger, major disturbances occur at very long intervals in these forests, such that small gaps and slight but chronic disturbances to overstory crowns necessarily play a major role in tree regeneration. Lorimer (1977, 1980) found that mesic pre-­European colonization forests in eastern North America and the Appalachian Mountains were characterized by all-­aged stands—­stands with an inverse J-­shaped size-­class distribution (Figure 13.9, Chapter 13). In mesic old-­growth forests of eastern North America, Runkle (1982) estimated average annual canopy disturbance rates of less than 1% per year, suggesting that it requires 110–­125 years on average for the entire canopy to be disturbed. For both tropical and temperate forests, Veblen (1992) estimated an average canopy rotation time of approximately 100 years and a range of 50–­575 years. Late-­successional species can perpetuate themselves more or less indefinitely in and near small canopy gaps in the absence of larger or more severe disturbances. Regeneration of species such as beeches, sugar maple, western and eastern hemlocks, firs, and basswoods occurs when persistent seedlings in the understory fill the gaps, excluding less-­tolerant species that need to first establish in the newly formed gap. A canopy profile through a single-­tree gap in a sugar

442  C h a p t e r 1 7   Forest Succession

maple–­beech forest is illustrated in Figure 17.13a. The gap is 50 m2 in the tree canopy and affects 250 m2 at ground level and formed from the breakage of a single stem 13 m above the ground about 3 years before the profile was made. Such small gaps may be filled by (i) adjacent canopy trees expanding branches into it or (ii) already-­established saplings filling it from below. This process takes place in temperate forests in gaps smaller than approximately 0.04 ha; the abundance of intolerant species sharply increases in larger gaps (Runkle 1985; Stewart et al. 1991; Busing 1994). Many authors report that two or more gaps are required for most saplings of tolerant species to reach the overstory. Mean residence time in the understory for tolerant species is often long; for sugar maple, it is often over 100 years and can be as long as 230 years (Barden 1981; Canham 1985). Mesic Douglas-­fir–­western hemlock forests of the Pacific Northwest appear to exhibit slower gap dynamics than other forests (Spies et al. 1990). Gap formation is typically less than the range of 0.5–­2.0% reported for temperate deciduous forests (Runkle 1985). Nevertheless, vertical forest structure provides an important context for gap dynamics in these forests. In Figure 17.13b, a small gap occurs in an old-­growth Douglas-­fir–­western hemlock forest whose overstory trees are over twice as tall as those in the beech–­sugar maple forest in Figure 17.13a. Because of their great height and narrow crowns, the death of one or a few conifers in such forests may not transmit enough light for the regeneration of mid-­or intolerant species (Canham et al. 1990; Spies et al. 1990). Gaps may remain important for the regeneration of Douglas-­fir and other early-­successional species where trees are shorter and canopies less dense, such as mixed-­conifer forests (Spies and Franklin 1989). Larger gaps may allow mid-­tolerant and intolerant species (opportunists or gap-­phase species) to establish and reach the overstory. Such species have light, wind-­disseminated seeds with the ability to germinate on small patches of exposed mineral soil, such as those created by the uptorn roots of windthrown trees. Juvenile growth under conditions of partial shade and partial root competition is rapid so that at least some seedlings can outgrow and overtop ­already-­established tolerant species in the openings. Typical gap-­phase replacement species include yellow birch and white ash, and with increasingly larger gaps include tulip tree, northern red oak, black cherry, and sweet birch. For example, Webster and Lorimer (2005) found that yellow birch was able to take advantage of gaps as small as 100–­400 m2 in hemlock–­northern hardwood forests in northern Wisconsin, although those most successful in reaching the canopy were already present at the time of gap formation. Small gaps create a shifting mosaic during the transition and old-­growth stages of mesic forests (see Table 17.4), but the forest canopy is not a simple closed-­canopy matrix punctuated by gaps (Lieberman et al. 1989). Tree crowns thin with age as their branches die, are sheared off by wind, or are broken by colliding with the branches of adjacent trees. The intensity of light reaching the forest floor therefore changes over time, and gaps represent an extreme condition where light may travel unimpeded to the forest floor. Latitude and the height of trees determine how much light penetrates beyond the perimeter of the gap itself and under the branches of surrounding overstory trees (Canham et al. 1990). In addition, succession in canopy gaps depends on many factors in addition to light, because of spatial variability in the amount of vegetative cover, soil water and nutrients, occurrence of coarse woody debris, and bare soil. Other factors include those associated with different regional and local ecosystems: latitude, physiography, drainage, nutrient cycling, tree species composition, disturbance regime, mutualisms with animals, and mycorrhizae. Of primary consideration in assessing gap dynamics is the disturbance type that creates the gap. For example, the specific type of wind damage can affect succession; trees uprooted by wind may form larger gaps and provide bare soil for colonizing species, whereas stem breakage may favor existing advanced regeneration in a smaller, more shaded gap. Runkle (1981, 1982, 1985, 1990) discovered in a variety of mesic forests of the eastern United States that small gaps 50–­200 m2 favor tolerant species but allow opportunists to persist in low densities. Gaps caused by the death of one species that favor seedlings of another species (i.e., reciprocal

Examples of Forest Succession  443

(a) 30 FA

25

FG

Distance (m)

AS

AS

20 AS

AS

AS

15

FG

AS AS

10 AS

FG

5

FG

0 0

5

10

(b)

20 15 Distance (m)

25

30

70 PM

60

50

50

40

30

40 TH 30

20

20

10

10

G Snag

PM

N

TH

10 m

60

Meters

F I G U R E  1 7 .1 3   Examples of tree-­fall gaps in eastern and western North American ecosystems. (a) Overstory canopy profile through a single-­tree gap. The gap, 50 m2 in the tree canopy and affecting 250 m2 at ground level, was formed when one stem broke at a height of 13 m. Crowns with solid lines are within 5 m of the observer, and crowns with broken lines are between 5 and 10 m of observer. FA, Fraxinus americana (white ash); FG, Fagus grandifolia (American beech); AS, Acer saccharum (sugar maple). The vertical lines in the understory are FG; thick lines are AS. (b) Vertical and horizontal profile through a 450-­year-­old Douglas-­fir/western hemlock stand in western Oregon. Source: After Spies et al. (1990) / Canadian Science Publishing. Reprinted with permission of National Research Council Canada. The small gap in the transect center (8–­10 years old) was created when a large Douglas-­fir broke off and fell on a smaller hemlock, breaking its crown. Tree regeneration consists entirely of western hemlock seedlings less than 0.5 m tall. PM, Pseudotsuga menziesii (Douglas-­fir); TH, Tsuga heterophylla (western hemlock); G, canopy gap. Source: After Runkle (1990). Reprinted with permission of National Research Council Canada.

444  C h a p t e r 1 7   Forest Succession

replacement; Fox 1977; Woods 1979, 1984; Runkle 1981) are more common in low-­diversity forests such as those dominated by beech and sugar maple stands, although exceptions exist. In the southern Appalachian Mountains, for example, four mid-­tolerant to intolerant species—­northern red oak, white ash, black cherry, and tulip tree—­occupied 3% of the canopy area via infrequent success in gaps, especially those formed by multiple trees (Barden 1981).

Gap Specialists: American Beech and Sugar Maple  American beech and sugar maple,

two of the most common species in eastern North America, are long-­lived, very tolerant, late-­ successional species of mesic forests where severe disturbances are very infrequent. These two species are the most likely of all deciduous species to perpetuate themselves over many centuries because they are able to regenerate in shaded understories and utilize treefall gaps to gain the canopy. Where the two species co-­occur, beech occurs over a wider variety of sites because it is less nutrient-­demanding, tolerates wetter sites (Crankshaw et al. 1965; Leak 1978), and is more shade tolerant than sugar maple. For example, beech occurred on 256 soil types in pre-­European ­colonization forests of Indiana, second only in distribution to white oak (Crankshaw et al. 1965). However, many investigators have sought to explain the co-­existence of these late-­successional species in mesic forests by their performance in relation to the light relations of forest understories. A long-­term study of beech and sugar maple coexistence is a 16-­ha beech–­sugar ­maple-­dominated ecosystem that is part of Warren Woods in southwestern Michigan. Beech was the main overstory species in 1933 as it was prior to European colonization in the surrounding area (Kenoyer 1933). At that time, beech had 10 times the number of trees >15 cm diameter at breast height (DBH) and represented 82% of the basal area (Cain  1935). Three-­fourths of the treefalls from 1946 to 1976 were of beech (Brewer and Merritt 1978). Meanwhile, sugar maple was more common as seedlings and saplings. Long-­term studies of the understory show that both species maintained themselves over the entire area without marked replacement of one by the other, a relationship termed allogenic coexistence (Poulson and Platt  1996). In this case, “allogenic” is emphasized because wind creates the treefall gaps that appear to explain the observed pattern of species interaction. Notably, autogenic ecosystem factors, such as old (300 years), tall (30–­40 m) trees with broad crowns, hollow trunks, brittle wood, and shallow rooting, also are significant in gap formation, which has increased markedly since 1975 in this aging stand. Beech and sugar maple understory individuals are able to coexist in this ecosystem due to their differences in understory tolerance, tree architecture, and how they respond to light levels both in the understory and in treefall gaps (Runkle 1985; Canham 1988; Poulson and Platt 1996). Beech replaces sugar maple in the deeply shaded understory and in small gaps where its saplings exhibit horizontal (plagiotropic) growth to obtain light. Sugar maple is favored where multiple treefall gaps occur; sugar maple has more rapid vertical (orthotropic) extension growth enabling it to quickly colonize the gap. Where a mix of treefall gaps of different sizes occurs at appropriate frequencies, both species may coexist at near-­equal abundances over the forest as a whole (Figure 17.14). Despite the importance of species trade-­offs based on light relations, the Warren Woods example is incomplete without consideration of soil properties and nutrient and water relations specific to this local ecosystem. Soils at Warren Woods are sandy and strongly acid, and the high spring water table tends to restrict rooting to the acidic topsoil. The less nutrient-­demanding understory beech should be favored over sugar maple on the low nutrient status of this site and contribute significantly to beech dominance with minimum canopy disturbance. This relationship was reflected by the overwhelming dominance of beech overstory in 1933 (Cain 1935). As windstorm disturbances increase in frequency within the forest as they have over the last several decades, understory dominance may shift to sugar maple (Figure 17.14, panel 4). The proportion of the species in the understory and overstory may reflect soil and nutrient differences in other ecosystems where sugar maple is favored, provided sufficient light is available for sugar maple regeneration. For example, sugar maple dominates only in two favorable ecosystems in the White

Examples of Forest Succession  445 25 YR BEFORE TREEFALL

25 YR AFTER TREEFALL

no gaps nearby local gaps every 100–200 yr

10 × 10 metres

1 cm DBH and >l–­2 m tall are shown. Bc = beech canopy tree; Mc = sugar maple canopy tree. Source: After Poulson and Platt (1996) / John Wiley & Sons. Reprinted with permission of the Ecological Society of America.

Mountains of New Hampshire, whereas beech is more abundant in eight (Leak 1978). In addition, in the heavy-­textured, calcareous morainal soils that supported beech–­sugar maple forest prior to European colonization in central and southeastern Michigan (Quick  1923; Veatch  1953,  1959), sugar maple appears to be much more vigorous than beech and the tree replacement patterns differ

446  C h a p t e r 1 7   Forest Succession

from those at Warren Woods. For example, in 1997 in an old-­growth forest on a clay-­rich, calcareous, nutrient-­rich moraine in southeastern Michigan, sugar maple dominated the overstory (49% sugar maple to 23% beech for basal area), and its saplings overwhelmingly dominated the understory (62–­13% for number of stems). Notably, sugar maple has become even more dominant on this site over the last 20 years in the understory (now 66–­14% for number of stems). Experimental evidence demonstrates that growth of sugar maple seedlings under low light is enhanced 2 to 4 times on rich sites in part because of higher nitrogen and water availability (Walters and Reich 1997).

FIRE AND OAK DOMINANCE—­O AKS AT RISK Oaks dominate the overstory of dry and dry-­mesic sites in much of eastern North America today, but the current generation may be their last on all but the most extremely dry sites. Succession of oak forests to more mesophytic and understory tolerant species has been documented for many different regions and ecosystems (Christensen  1977; Lorimer  1984; McCune and Cottam  1985; Host et al. 1987; Pallardy et al. 1988; Hammitt and Barnes 1989; Abrams and Downs 1990; Nowacki et al. 1990; Nowacki and Abrams 1991; Abrams 1992; Shotola et al. 1992; Abrams 2003; Aldrich et al. 2005; Nowacki and Abrams 2008; Johnson et al. 2019). Although oaks dominate in the overstory, seedlings are relatively few and there is virtually no recruitment to the sapling understory. Such a trend, if continued, would likely create major differences in ecosystem structure and function, including shifts in both animal and plant communities. Upland oaks evolved in dry, fire-­prone ecosystems and are characterized by a suite of fire and dry-­site adaptations (Chapter 10; Abrams 1990). Depending on regional and local site conditions, fire seasonality, frequency, intensity, and severity are instrumental in maintaining oak regeneration and determining the distribution and occurrence of oak species. Widespread fire exclusion throughout much of the East is perceived as the major reason for lack of oak regeneration, as is herbivory by deer in many areas (Chapter 12). Red maple and sugar maple are the understory tolerant, mesophytic species replacing oaks in the understory on all but the driest sites. Opportunistic species such as black cherry, sweet birch, blackgum, and sassafras can also establish and develop in the absence of fire where oak advance regeneration fails and logging or wind disturbance to the oak overstory provides sufficient light. Successful oak regeneration depends on an abundance of oak saplings in the forest understory before disturbance, that is, the presence of advanced regeneration. There are many factors that drive oak regeneration, such as seed production and dispersal or animal damage to seeds and seedling (Lorimer  1993). However, fire is critically important for successful oak regeneration because it kills competitors and promotes resprouting of young oaks. In addition, fire provides a source of readily available nutrients for developing oak sprouts when it consumes organic matter. As mid-­tolerant species, young upland oaks are at a competitive disadvantage from many early-­ successional species in open understories as well as from the species that tolerate shade. Thus, mesophytic species are always invading the understories of oak forests by wind and animal dispersal, but oak dominance is maintained by fire. Light and moderate surface fires historically occurred at all stages of stand development in oak forests, indiscriminately killing tree seedlings, saplings, and shrubs. Few species have the resprouting ability exhibited by oaks (and their usual associates, the hickories) in the seedling or sapling stage, such that repeated fires act to purge the understory of mesophytic competitors in oak–­hickory forests. Young oak stems killed by fire resprout vigorously from well-­established root systems, although the thick bark of many young saplings allows them to survive surface fires. The density of the oak overstory may be high enough to deter or significantly slow the invasion of mesophytic species into the understory during stand initiation and stem exclusion. Fire is necessary to kill invaders and reduce shrub competition from the understory reinitiation stage onward, when

Examples of Forest Succession  447

oaks are themselves becoming established and forming advanced regeneration. When crown breakage or gaps form during the old-­growth stage, the advanced regeneration of oaks and hickories is well positioned to dominate the understory (as in Figure 13.13, Chapter 13). On the driest sites of ridges and fire-­prone barrens, oaks (blackjack, post, chestnut, black, northern pin) may be self-­replacing or only very slowly replaced by other species (Abrams 1992). In these ecosystems, fire, wind, soil movement, or other disturbances, together with other species adaptations to extreme sites, may maintain the intolerant oak species and associates of hickory or pine. Understanding a complex process such as forest succession is a tall order. Almost all ecological research (at least on secondary succession) has been conducted within the life span of existing forests, and therefore successional pathways must always be inferred or modeled rather than observed. It is notable that macroclimate has been considered to be the grand driver of succession since the concept was first articulated, and thus nearly everything we think we know about how forests are likely to change over time is likely to be altered in the current era of rapid climate change. From this perspective, despite the fact that forest succession has been studied in detail for well over a century, ongoing and future research on successional processes should remain at the forefront of forest ecology. We examine climate change and its effects on forests in the final part of this book.

SUGGESTED READINGS Abrams, M.D. (1992). Fire and the development of oak forests. BioScience 42: 346–­353. Buma, B., Bisbing, S., Krapek, J., and Wright, G. (2017). A foundation of ecology rediscovered: 100 years of succession on the William S. Cooper plots in Glacier Bay, Alaska. Ecology 98: 1513–­1523. Chapin, F.S. III, Walker, L.R., Fastie, C.L., and Sharman, L.C. (1994). Mechanisms of primary succession following deglaciation at Glacier Bay, Alaska. Ecol. Monogr. 64: 149–­175. Christensen, N.L. (1988). Succession and natural disturbance: paradigms, problems, and preservation of natural ecosystems. In: Ecosystem Management for Parks and Wilderness (ed. J.K. Agee and D.R. Johnson). Seattle: University of Washington Press.

Crisafulli, C. and Dale, V. (ed.) (2018). Ecological Responses at Mount St. Helens: Revisited 35 Years After the 1980 Eruption. New York: Springer 351 pp. Glenn-­Lewin, D.C. and van der Maarel, E. (1992). Patterns and processes of vegetation dynamics. In: Plant Succession: Theory and Prediction (ed. D.C. Glenn-­ Lewin, R.K. Peet and T.T. Veblen). London: Chapman and Hall. Matthews, J.A. (1992). Chapter  6. Plant succession: processes and models. In: The Ecology of Recently-­ Deglaciated Terrain, 250–­316. Cambridge: Cambridge University Press. Pickett, S.T.A., Collins, S.L., and Armesto, J.J. (1987). Models, mechanisms and pathways of succession. Bot. Rev. 53: 335–­371.

Carbon Balance of Trees and Ecosystems

CHAPTER 18

A

lthough forest ecosystems cover only 21% of the Earth’s land surface, they constitute a ­disproportionately large share of terrestrial plant mass (75%) and its annual growth (37%). Because plants are largely constructed of carbon (47%), the extent to which forests are altered by human activities (i.e., harvesting or conversion to other types of vegetation) has a substantial influence on the pattern in which carbon (C) is cycled and stored at local, regional, and global scales. Recent attention has focused on factors influencing the fixation and release of C by trees and forest ecosystems (i.e., their C balance), because the rising CO2 concentration of the Earth’s atmosphere and projected changes in global climate have the potential to alter the present-­day geographic distribution of forests and the rate at which they sequester CO2 from the atmosphere (Pastor and Post 1986; Melillo et al. 1993). Such a change is of both ecological and economic importance, because the production of plant matter in forest ecosystems not only influences the global cycling of C but also represents an indispensable source of food, fuel, fodder, and fiber for human populations around the Earth. Globally, forests also are home to countless species of animals, as well as vascular and non-­vascular plants. Importantly, aspects of climate change, such as rising atmospheric CO2, altered temperature and precipitation regimes, and the deposition of human-­ derived nitrogen (N), have the potential to alter the C balance of forest trees, forest ­ecosystems, and also that of the entire Earth. The C balance of plants (i.e., growth) is controlled by the quantity of CO2 fixed through photosynthesis and the rate at which photosynthetically fixed C is returned to the atmosphere by respiring plant tissues. Light, temperature, and the availability of soil water and nutrients ­constrain these physiological processes and consequently influence the productivity of plants and the ecosystems in which they occur. One needs to only examine desert, grassland, and forest ecosystems to understand that the processes controlling the productivity of individual plants also have a substantial influence on the productivity of terrestrial ecosystems. This situation is particularly true in forests, in which the growth and metabolism of plants far outweigh that of any other organism. From an ecological perspective, the flows of C and energy are somewhat interchangeable; that is, plants transduce solar energy into C-­based biochemical energy via photosynthesis. Further, it is plant-­derived biochemical energy that drives the subsequent flow of energy through entire terrestrial ecosystems. In this chapter, we focus on the physiological processes of plants that control the C balance (i.e., the flow of energy) of terrestrial ecosystems, especially forests. Importantly, the principles we present here apply equally to the C dynamics of all terrestrial ecosystems, whether they are tundra, desert, or grassland. We discuss how climate and soil nutrient availability influence the C balance of plants, and hence their growth. We then build upon this information to understand how environmental factors influence the C balance of ecosystems at local, regional, and global scales.

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

449

450  C h a p t e r 1 8   Carbon Balance of Trees and Ecosystems

CARBON BALANCE OF TREES

Only a small fraction of the radiant energy reaching the Earth’s surface (2%) is used by green plants to assimilate atmospheric CO2 into organic compounds via photosynthesis. These compounds are used to construct new plant tissue, maintain existing tissue, create storage reserves, or provide defense against insects and pathogens. Simply stated: trees grow (or gain C) when the amount of CO2 fixed through photosynthesis exceeds the amount of CO2 lost from respiring leaves, branches, stems, and roots. Understanding the factors influencing the balance between photosynthesis and respiration is of ecological importance, because plants with the greatest net C gain under a specific set of environmental conditions (i.e., light, water, and nutrient availability) are often the best competitors. It follows that fast-­growing plants require rapid rates of photosynthesis, but the converse is not always true. As you will learn, the net C gain of plants is not influenced by photosynthetic rate alone (Ceulemans and Saugier 1991). In the following section, we trace the flow of photosynthetically fixed CO2 to the construction and maintenance of plant tissue, focusing on how environmental factors influence the net C gain of plants (i.e., growth).

PHOTOSYNTHESIS, DARK RESPIRATION, AND LEAF C GAIN The processes of photosynthesis and respiration occur in the leaves of all plants, and the balance of these more or less opposing physiological processes controls the net C gain of leaves. Gross photosynthesis is the total amount of C plants assimilate from the atmosphere. However, a portion of that fixed C is returned to the atmosphere from leaves as CO2 during dark respiration. This process is so termed because leaf respiration can only be determined in the absence of C fixation and is hence measured on unilluminated leaves. In leaves, the process of dark respiration provides the biochemical energy (i.e., ATP) to repair proteins and membranes mediating photosynthesis. Net photosynthesis is the balance between gross photosynthesis and leaf dark respiration, and it represents the amount of photosynthate available for the growth and maintenance of non-­ photosynthetic tissue, as well as that allocated to storage and defense. Light intensity, temperature, and the availability of water and soil nutrients (especially N) all influence photosynthesis, respiration, and the C gain of leaves. Rates of net photosynthesis measured under saturating light, optimum air temperature, low vapor pressure deficit, and ambient atmospheric CO2 represent the maximum photosynthetic capacity of plants (see Table  18.1). Care should be taken in interpreting Table  18.1, because maximum photosynthetic capacity varies with leaf canopy position (shade versus sun leaves), as trees develop from a juvenile to an adult phase (see ponderosa pine), soil nutrient availability, and measurement technique. Also be aware that maximum photosynthetic capacities are often poorly correlated with the C gain of forest trees (Ledig and Perry 1969), because environmental factors (light, temperature, water, and soil nutrients) constrain photosynthesis under field conditions, often well below its maximum. Nevertheless, Table 18.1 illustrates several important physiological differences among forest trees that have important ecological implications. The maximum photosynthetic capacity of forest trees ranges from 2 to 25 μmol m−2  s−1 (Table 18.1), generally lower than that of most agronomic crops (Ceulemans and Saugier 1991). Shade-­intolerant deciduous species have some of the highest photosynthetic capacities of all trees (up to 25 μmol m−2 s−1; Table 18.1), whereas those of shade-­tolerant deciduous species are much lower (3–6 μmol m−2 s−1; Table 18.1). Recognize that shade tolerance is only one of several characteristics enabling a species to persist in the understory (i.e., understory tolerance; Chapter  13). Although some conifers have high photosynthetic capacities (see Scots and Monterey pine), most are modest when compared to those of deciduous trees. Canopy architectures efficient at intercepting solar radiation can outweigh moderate photosynthetic rates, allowing some trees to attain large C gains despite the relatively low net

Carbon Balance of Trees  451

Table 18.1  The photosynthetic capacity of trees under conditions of saturating light, optimum temperature and water regimes, and ambient atmospheric CO2 (350 μmol mol−1). Rates are expressed in micromoles of CO2 assimilated by 1 square meter of leaf during 1 second. Shade-­understory tolerance

Maximum photosynthetic capacity μmol CO2 m−2 s−1

Tolerant Deciduous trees sugar maple

3–4

American beech

4–5

hop-­hornbeam American basswood flowering dogwood

6 3 4–6

Coniferous trees European silver fir

2–4

white fir

8

grand fir

4–5

balsam fir seedling

2

Sitka spruce

3–9

Norway spruce

2–7

subalpine fir

9

Mid-­tolerant Deciduous trees white oak

2

northern red oak

4–5

red maple

3–5

white ash

4–5

yellow birch

9

Coniferous trees Engelmann spruce

3–5

Douglas-­fir

2–6

Intolerant Deciduous trees black oak

7

sassafras

6

black cherry

4

sweet gum

6–11

tulip tree

7–17

paper birch

10

black locust

13–17

red ash

20–25

bigtooth aspen cottonwood

14 15–19 (Continued )

452  C h a p t e r 1 8   Carbon Balance of Trees and Ecosystems

Table 18.1  (Continued) Shade-­understory tolerance

Maximum photosynthetic capacity μmol CO2 m−2 s−1

trembling aspen

20–22

black poplar hybrid

20–25

Coniferous trees lodgepole pine

3–9

Ponderosa pine seedlings mature trees old-­growth Scots pine Monterey pine loblolly pine

25 5–14 4 10–16 17 3–6

Source: All values are expressed on a one-­sided leaf area basis and are summarized from Bazzaz (1979), Ceulemans and Saugier (1991), Jurik et al. (1988), Larcher (1969), Wallace and Dunn (1980), and Walters et al. (1993).

photosynthetic rates of individual leaves. The rapid growth and modest net photosynthetic capacities (2–10 μmol m−2 s−1) of some conifers (e.g., Douglas-­fir) illustrate this point (Table 18.1). Coniferous canopies disperse incoming solar radiation over a larger number of leaves than broad-­ leaved canopies, enabling them to capture more solar energy at high light intensities. The ­orientation of coniferous leaves and branches also produces less self-­shading of leaves lower in the canopy, enabling them to photosynthesize at a relatively high rate. Also note that the photosynthetic capacities in Table 18.1 are expressed on an area basis (μmol CO2 per square meter of leaf surface per second), a factor that can be compensated for by canopy architecture as described in the earlier text. Additionally, coniferous species often maintain higher leaf areas than deciduous species, creating a larger surface area to capture solar radiation. For example, Douglas-­fir and many true firs maintain leaf areas 2–6 times greater than that of some broad-­leaved species. As such, low net photosynthetic rates in some conifers are offset by an efficient light-­gathering canopy architecture, enabling them to achieve relatively large C gains, despite relatively low rates of maximum net photosynthesis.

LIGHT AND LEAF C GAIN Net photosynthesis varies substantially with light intensity, and the response of trembling aspen (shade intolerant) and northern red oak (mid-­tolerant) leaves to increasing light intensity illustrates this important point (Figure 18.1). Sun leaves of trembling aspen attain a greater light-saturated rate of net photosynthesis (i.e., maximum photosynthetic capacity; see right-­hand portion of Figure  18.1) than do the sun leaves of red oak. The shade leaves of trembling aspen also have greater light-­saturated rates of net photosynthesis than do the shade leaves of northern red oak, but rates in the shade leaves of both species are lower than those of sun leaves. Although shade leaves photosynthesize at a lower rate than sun leaves, they can contribute as much as 40% of the C assimilated by tree canopies (Schulze et al. 1977). Plants attaining high rates of net photosynthesis, such as trembling aspen and other shade intolerant species, typically contain large amounts of chlorophyll and the enzymes used for CO2 fixation (Field and Mooney 1986), which represent the biochemical machinery plants use to convert light energy into biochemical energy. High respiration rates are associated with plant tissues

Carbon Balance of Trees  453 30

25 Aspen sun leaf 20

Light-saturated photosynthetic rates

Net photosynthesis (mg CO2/dm2•h)

F I G U R E  18 .1   The photosynthetic response to changing light intensity for trees of contrasting tolerance. Trembling aspen is an intolerant species, whereas northern red oak is a mid-­tolerant species that can persist in the forest understory. Source: After Loach (1967) / John Wiley & Sons.

15

Oak sun leaf

10

Aspen shade leaf 5 Oak shade leaf 0

F I G U R E  1 8 .2   The relationship between photosynthetic capacity and leaf dark respiration for deciduous broad-­leaved (closed circles), evergreen broad-­leaved (half-­open circles), and coniferous (open circles) tree species. Source: After Ceulemans and Saugier (1991) / John Wiley & Sons. Reprinted from Physiology of Trees, © 1991 by John Wiley & Sons, Inc. Reprinted with permission of John Wiley & Sons, Inc.

Dark respiration (μmol•m–2•s–1)

0.2 Compensation points

0.4

0.6

0.8

1.0

Relative light intensity

y = 0.0913x + 0.455 r = 0.750

4

3

2

1

0

0

5 10 15 20 25 Photosynthetic capacity (μmol•m–2•s–1)

30

containing large amounts of metabolically active enzymes (Amthor 1984); such high rates of respiration are needed to replace and repair enzymes and membranes involved in the biochemical processes mediating photosynthesis. Consequently, leaves with high net photosynthetic rates also respire rapidly. Leaf dark respiration represents a relatively constant proportion (7–10%) of the maximum photosynthetic capacity of many forest trees (Figure 18.2; Ceulemans and Saugier 1991). However, rates of net photosynthesis rarely attain their maximum under field conditions, because light, soil nutrients, or water are often in short supply. Under these conditions, dark respiration can account for 10–20% of annual net photosynthesis in a wide range of coniferous and deciduous tree canopies (Ryan et al. 1994).

454  C h a p t e r 1 8   Carbon Balance of Trees and Ecosystems

Because rates of net photosynthesis vary with light intensity, whereas dark respiration remains constant across a range of light intensities, the ability of a species to persist in low light environments is related to the balance of these physiological processes. In Figure 18.1, the light intensities at which net photosynthesis is zero (i.e., light compensation point) for the sun and shade leaves of trembling aspen are substantially higher than those of northern red oak. This relationship results from high dark respiration rates that maintain rapid rates of photosynthesis in aspen leaves. Dark respiration is lower in shade tolerant species, thereby enabling them to maintain positive rates of net photosynthesis (i.e., positive leaf C gain) at much lower light intensities. This ability illustrates the point that high photosynthetic rates alone do not insure a high C gain in all light environments.

TEMPERATURE AND LEAF C GAIN Air temperature directly controls leaf C gain by influencing both rates of gross photosynthesis and dark respiration (Figure 18.3). The difference between these two physiological processes (i.e., net photosynthesis) reaches a maximum between 15 and 25 °C for most temperate trees (Kozlowski and Keller 1966; Mooney 1972). Rates above and below this range rapidly decline because low and high temperatures limit the process of CO2 fixation. In Figure 18.3 we see that gross photosynthesis increases with rising temperature to an optimum and then rapidly declines with a further increase in temperature. Dark respiration also increases with rising temperature, attains a maximum, and then declines markedly as temperature continues to rise. However, the response of dark respiration to temperature differs from that of gross photosynthesis in several important ways (F­igure 18.3). First, the maximum rate of dark respiration occurs at a much higher temperature than does the maximum rate of gross photosynthesis. Also, the slope of the gross photosynthesis and respiration curves differs at high temperatures (i.e., 35–50 °C). In this range of temperatures, gross photosynthesis declines rapidly, whereas dark respiration continues to increase. Note also that leaf C gain approaches zero as temperature rises between 35 and 43 °C. At temperatures above 43 °C, rapid rates of dark respiration combined with slow rates of gross photosynthesis result in a negative C

Optimum temperature range for net photosynthesis Net CO2 balance = 0

Gross photosynthesis

Respiration Net loss in carbohydrate production

Net gain in carbohydrate production

0

10

20

30

40

50

60

Temperature, °C F I G U R E  1 8 .3   The temperature response of gross photosynthesis, dark respiration, and leaf C gain. Gross photosynthesis and dark respiration in leaves are temperature-­dependent processes, and leaf C gain is maximized at the temperature where the difference between these processes is greatest. Source: After Daniel et al. (1979) / McGraw-­Hill. Reprinted from Daniel, T., J. A. Helms, and F. S. Baker, Principles of Silviculture, 2nd ed., © 1979 by McGraw-­Hill, Inc. Reprinted with permission of The McGraw-­Hill Companies.

Carbon Balance of Trees  455

balance and the net loss of C from the leaf. Thus, you can see that changes in rates of gross p­hotosynthesis and dark respiration in response to temperature have a great impact on leaf C gain (Berry and Bjorkman  1980; Berry and Downton  1982), controlling whether it is positive (net C gain) or negative (net C loss).

WATER AND LEAF C GAIN Water is important for plant growth and its availability directly influences photosynthesis and leaf C gain. Plants are faced with a fundamental trade-­off because CO2 fixation and water loss occur simultaneously in leaves. Trees open their stomata allowing CO2 to diffuse from the atmosphere to enzymatic sites of CO2 fixation in the leaf mesophyll. At the same time, water vapor diffuses out of stomata in response to differences in water potential between the leaf and atmosphere (see Chapter 9). Declines in leaf water content from 5 to 10% often have little influence on photosynthesis (Hanson and Hitz 1982), but a drop in leaf water potential beyond this level can sharply reduce CO2 fixation. Low leaf water potentials cause stomata to close, restricting the loss of water to the atmosphere and the flow of CO2 into the leaf. At leaf water potentials of −1.0 to −2.5 MPa (F­igure 18.4), stomatal closure begins to limit the diffusion of CO2 into the leaf, thereby slowing rates of photosynthesis. As leaf water potentials fall below a critical level, stomata completely close and photosynthesis ceases. Although the species in Figure  18.4 inhabit a wide range of sites (dry-­mesic uplands for white pine and northern red oak; wet lowlands for speckled alder, red ash), leaf water potential limits photosynthesis over a surprisingly narrow range of values. Plants occurring in dry sites have several physiological and morphological adaptations enabling them to maintain high leaf water potentials at relatively low leaf water contents (percent water). Physiological changes in cell wall elasticity, membrane permeability, and the concentration of solutes within plant cells can maintain high leaf water potentials (i.e., less negative), allowing photosynthesis to occur at relatively low leaf water contents (Bradford and Hsiao 1982). Trees inhabiting dry environments also have a higher rate of gas exchange, lose turgor at a lower leaf water potential, and have greater root:shoot ratios, leaf thicknesses, guard cell lengths, and stomatal densities than trees from mesic sites (Abrams et al. 1994). Such adaptations provide dry-­site species with increased water-­use efficiency (μmol CO2 fixed/mmol of H2O transpired), allowing them to fix more CO2 per unit of water lost.

12

Net carbon uptake by leaves (mg CO2 dm–2 h–1)

F I G U R E  1 8 .4   The relationship between leaf water potential and net photosynthesis for deciduous and coniferous trees occurring in different habitats. Source: After Hinckley et al. (1981) / with permission of Elsevier. Reprinted from Water Deficits and Plant Growth, © 1981 by Academic Press, Inc. Reprinted with permission of Academic Press, Inc.

Alnus rugosa

10 8

Fraxinus pennsylvanica

6

Quercus rubra 4

Pinus rigida

2 0

Pinus strobus

0

–1.0

–2.0

Leaf water potential, MPa

–3.0

456  C h a p t e r 1 8   Carbon Balance of Trees and Ecosystems

SOIL NITROGEN AVAILABILITY AND LEAF C GAIN Photosynthesis as well as other biochemical processes in the leaf are influenced by the supply of soil nutrients. Nitrogen is a constituent of chlorophyll, other light-­harvesting pigments, photosynthetic membranes, and the enzymes mediating photosynthesis. Their presence makes leaves rich in N compared to most other plant tissues. Thus, the supply of N can have a large influence on the ability of a leaf to fix CO2 from the atmosphere. For example, leaf N concentration is directly related to the photosynthetic rate of a wide array of herbaceous plants (Field and Mooney 1986). Net photosynthesis in forest trees also increases with leaf N concentration, but coniferous and deciduous species respond much differently (Brix 1971; Linder et al. 1981). In deciduous species, net photosynthesis is particularly responsive to leaf N concentration (Linder et  al.  1981). For example, the highest rates of net photosynthesis for European white birch occur at the greatest leaf N concentration (5%), and rates sharply drop as leaf N concentrations decline (Figure 18.5). Coniferous species, on the other hand, lack the ability to increase photosynthetic enzyme contents to the same extent as deciduous trees, making photosynthesis much less responsive to changes in leaf N concentration, which often reflects soil N availability. Photosynthetic rates in conifers, such as Douglas-­fir, initially increase up to 25% following N fertilization, but rates rapidly decline with time (Brix 1971). However, conifers often increase whole-­tree C gain following N fertilization by producing more foliage, thus creating a larger photosynthetic surface area to capture CO2 (Brix and Ebell 1969; Brix 1971). In deciduous trees, N fertilization often increases both the photosynthetic rate of individual leaves as well as leaf production, resulting in a larger and more photosynthetically active canopy.

CONSTRUCTION AND MAINTENANCE RESPIRATION Forest trees are unique compared to other types of terrestrial vegetation due to their size and longevity. As a consequence of their large mass, forest trees require substantial amounts of photosynthate (i.e., carbohydrates produced by photosynthesis) to maintain the function of living cells located in the non-­photosynthetic tissues of branches, stems, roots, and reproductive structures. Photosynthate also is required for the construction of new leaves to capture light, roots to forage for water and nutrients, and reproductive structures to insure regeneration. The construction and maintenance of these tissues require the products of photosynthesis to generate cellular energy (i.e., ATP). Respiration during the biosynthesis of new tissue is construction respiration, whereas respiration used to biochemically sustain already existing living tissue is maintenance

Net photosynthesis (mg CO2 /g leaf /h)

30

20

10

0 0

110

220

230

Photon flux density, μE/m2/s

440

0

1

2

3

4

5

% Nitrogen in leaf

F I G U R E   18. 5  Leaf N concentration is an important factor controlling the rate of net photosynthesis in deciduous species. The highest rates of net photosynthesis for European white birch occur at saturating light intensities and high leaf N concentrations. Source: Linder et al. (1981) /U.S. Department of Energy / Public Domain.

Carbon Balance of Trees  457

respiration. These physiological processes, which return fixed CO2 to the atmosphere, have an important influence on the C balance of all plants. Construction Respiration: Respiration provides the cellular energy (ATP) used to c­onvert the products of photosynthesis into the biochemical constituents of new plant tissue. Thus, photosynthetically fixed CO2 is lost to the atmosphere (as CO2) during the synthesis of new plant tissue. This loss (i.e., construction respiration) is calculated knowing the biochemical constituents of a particular plant tissue and the biosynthetic pathways by which they are formed. This information can be used to determine the amount of CO2 produced during the biosynthesis of a particular plant compound or organ. Glucose, a product of photosynthesis, is often considered the substrate for biosynthesis when calculating the cost of construction respiration during the biosynthesis of a plant compound or a tissue (Penning de Vries 1975; Chung and Barnes 1977). Look at Table 18.2 and note that the needles and shoots of loblolly pine contain large proportions (%) of carbohydrates, lignin, and phenolic compounds, whereas nitrogenous compounds, lipids, and organic acids are present in smaller amounts. The biosynthetic pathways giving rise to these compounds require different amounts of substrate (i.e., glucose) and produce different amounts of CO2 as they are biochemically synthesized. For example, the synthesis of 1 g of carbohydrate during cell wall construction requires 1.2 g of glucose as substrate, of which 11% is lost as CO2 (Table  18.2). Although lipids are present in relatively small amounts, their construction p­roduces relatively large amounts of CO2 compared to other biosynthetic pathways. By knowing the constituents of a particular plant tissue and the amount of CO2 lost during their biosynthesis, one can calculate construction respiration for plant tissues. For example, during the construction of 1 g of loblolly pine needles, approximately 19% of substrate glucose is respired during biosynthesis. Similarly, construction respiration consumes approximately 16, 18, and 17% of the glucose used to synthesize loblolly pine shoots, cambium, and roots, respectively (Chung and Barnes  1977). Clearly, accounting for these losses of C is essential for understanding the C balance of whole plants, as well as the C balance of forest ecosystems. It is important to realize that construction respiration is a fixed cost to the plant. Regardless of temperature or other environmental conditions that a plant faces, the amount of photosynthate used to synthesize a protein, carbohydrate, or for that matter a leaf, is invariant. The process of construction respiration may occur at a faster or slower pace depending on temperature, but the amount of photosynthate used to construct a compound or tissue remains the same. This differs dramatically from maintenance respiration, which is highly response to environmental conditions. Table 18.2  The biochemical constituents of the needles and shoots of loblolly pine, and the proportion of carbon respired during their biosynthesis. Needles Compound

Shoots

Construction respirationa

-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­Percent -­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­-­

Carbohydrates

35.7

38.0

11

Lignin

23.1

23.3

14

Phenolic compounds

21.0

20.0

29

9.3

8.4

25

Nitrogenous compounds Lipids

5.6

5.3

49

Organic acids

3.8

3.5

32

 Calculated as the percent of glucose lost as CO2 during the synthesis of a compound in the left-­hand column. Source: After Chung and Barnes (1977) / Canadian Science Publishing. Reprinted with permission of National Research Council Canada.

a

458  C h a p t e r 1 8   Carbon Balance of Trees and Ecosystems

Maintenance Respiration: Carbohydrates produced by photosynthesis are used to sustain the function of living plant tissues during the process of maintenance respiration. Protein synthesis and replacement, membrane repair, and the maintenance of ion gradients across cell membranes all require the expenditure of energy initially derived through photosynthesis (Ryan 1991). The rate of maintenance respiration, like all other enzymatic processes, increases predictably with temperature. For most plants, maintenance respiration roughly doubles for every 10 °C increase in ambient temperature. That is, maintenance respiration increases exponentially with a linear increase in temperature. This fundamental fact has important implications for the C balance of plants, as well as entire ecosystems, as the Earth’s climate continues to warm from fossil fuel burning. Maintenance respiration also is strongly influenced by tissue N concentration (Waring et al. 1985), wherein tissues with high N concentrations also have high rates of maintenance respiration. In fact, leaf maintenance respiration for a wide range of plant species can be predicted from N concentration using the following equation (Ryan 1991):



RM

0.0106 N

where RM is leaf maintenance respiration (mmol of CO2 mol tissue C−1 h−1) and N is the total nitrogen concentration of that tissue (mmol N mol tissue C−1). This relationship arises because the majority of N present in plant tissue occurs in proteins, and a large proportion (60%) of maintenance respiration is used for their repair and replacement (Penning de Vries 1975). Such a relationship also contributes to the high rates of leaf dark respiration that accompany rapid photosynthetic rates (see Figure 18.2). Herein lies an important link between soil N availability, leaf N concentrations, and the physiological processes that control plant C gain. Greater soil N availability can increase the N concentration of plant tissues. This can lead to greater rates of photosynthesis in leaves and more rapid rates of maintenance respiration in non-­photosynthetic tissue. However, increases in maintenance respiration of non-­photosynthetic tissues (i.e., plant roots) are surpassed by increases in net photosynthesis in leaves with higher N concentrations. In combination, these responses often lead to higher plant C gains when soil N availability increases. As trees develop from juvenile to adult phases, the ratio of photosynthetic to non­­­ photosynthetic tissues decreases, a factor that substantially influences maintenance respiration and the C balance of whole trees. Canopy leaf area reaches a maximum relatively early in tree development, thereby imposing a limit on canopy photosynthesis and the supply of photosynthate for tissue construction and maintenance. Although the supply of carbohydrate is constrained by canopy photosynthesis, the demand for carbohydrate rises as the proportion of non-­photosynthetic tissue continues to increase during tree ontogeny. For example, stemwood accounts for a large proportion of total tree mass (approximately 50–70% in mature trees) and contains both living (sapwood) and dead tissue (heartwood). Although the proportion of living, non-­photosynthetic tissue in stems decreases as trees mature, the absolute mass of living non-­photosynthetic tissue continues to increase. Consequently, the amount of photosynthate allocated to the maintenance of living cells in stemwood dramatically increases as trees develop from a juvenile to an adult (Agren et al. 1980; Waring and Schlesinger 1985; Ryan 1988). Remember that the product of photosynthate used to maintain the mass of live, non-­photosynthetic tissue in mature trees cannot be used for the construction of new tissue. Consequently, the greater cost of tissue maintenance in mature trees results in a decline in annual growth. Annual growth also can be slowed by reductions in the net photosynthetic rate of adult trees. As you will read later in this chapter, increases in maintenance respiration and declines in growth as trees mature have important implications for the C balance of forest ecosystems.

Carbon Balance of Trees  459

ALLOCATION TO STRUCTURE, STORAGE, AND DEFENSE In the previous paragraphs, we have discussed physiological processes influencing the fixation and loss of C by forest trees. Accounting for these processes, one can construct a C budget illustrating how plants partition C into the growth and maintenance of leaves, stem, roots, reproductive structures, as well as the production of defense and storage compounds. Such an approach can shed light onto how light, water, and nutrients influence the allocation of photosynthate into organs that aid in gathering light or the acquisition of water and soil nutrients. Summarized in Table  18.3 is the annual C budget for a young Scots pine tree (Agren et al. 1980). This budget quantifies the partitioning of photosynthetically fixed CO2 into the growth of various plant tissues and accounts for their respiration during construction and maintenance. Approximately 88% of fixed C in this young tree was used to produce new tissue, whereas only 10% was consumed during construction and maintenance respiration. The production of new roots to forage for water and nutrients in soil represented a large investment of photosynthate, accounting for 56% of annual net photosynthesis. In combination, root growth and respiration comprised almost 62% of annual net photosynthesis, indicating a substantial proportion of fixed C was allocated to belowground growth and metabolism to acquire soil resources. The amount of photosynthate plants allocate to growth, storage, and the production of defense compounds is under strong genetic control. Patterns of C allocation vary among species, within a species seasonally as well as with increasing age. Environmental factors also play an important role in influencing the allocation of C into structure, storage, and defense. Figure 18.6 illustrates the C allocation priorities in lodgepole pine. Allocation of photosynthate to new foliage and buds represents the highest priority for carbohydrate, followed by the production of new roots. Tree growth often is limited by light, water, and nutrients, making it necessary for the plant to first meet the carbohydrate demand of tree crowns and roots. The storage of carbohydrate in leaves, stem, and coarse roots has a lower priority than allocation to new leaves and roots. Allocation of Table 18.3  The carbon budget of a 14-­year-­old Scots pine tree. Assimilation

Allocation

Percent of total

g C year−1 Net photosynthesis

1723

Growth Current needles

286

16.6

Branch axes

132

7.7

Stem

145

8.4

Roots

960

55.6

1523

88.4

49

2.8

Total growth Construction and maintenance respiration Stem Branch axes Roots Total respiration Growth + respiration Unaccounted net photosynthesis Total Source: After Agren et al. (1980).

15

0.9

109

6.3

173

10.0

1696

98.5

27

1.5

1723

100.0

460  C h a p t e r 1 8   Carbon Balance of Trees and Ecosystems

Pr ote ve cti e Ch ca mi 5 ls

New Leaves 1

Buds 1

Canopy Storage 3

Diameter Growth 4 Stem Storage 3

Storage 3

New Roots 1-2

F I G U R E  18.6  The allocation of photosynthate into structure, storage, and defense in lodgepole pine. Tree components with low numbers (e.g., one) represent strong priorities for allocation, whereas components with high numbers indicate a low priority for allocation. Source: Waring and Pitman (1985) / John Wiley & Sons. Reprinted with permission of the Ecological Society of America.

carbohydrate to stem growth follows that to storage and the production of chemicals to defend against herbivory, which is the lowest priority. Patterns of C allocation vary substantially over the growing season, wherein certain tissues function as sinks or sources for carbohydrate. Nowhere is this more evident than during the phenological development of deciduous trees. Storage reserves in buds, branches, stems, and roots function as a source of carbohydrate to build new leaves early in the growing season. Once the expanding canopy attains a positive C balance, the priorities of allocation shift. In sugar maple seedlings, for example, large amounts of photosynthate are allocated to the production of new leaves early in the growing season (Table 18.4). The proportion of C allocated to stems and coarse roots increases following canopy development (i.e., mid growing season), and represents the greatest sink for photosynthate late in the growing season. The amount of carbohydrate stored in stems, coarse roots, and fine roots also increases over the growing season, attaining the highest priority late in the growing season. These carbohydrate stores are then used to construct new leaves, as well as new fine roots, at the start of the next growing season.

Carbon Balance of Trees  461

Table 18.4  The seasonal C allocation to leaves, stem, and roots in sugar maple seedlings. Percent of growth allocated to: Leaves

Stem

Coarse rootsa

Fine rootsb

Early growing season

85

5

0

10

Mid growing season

30

30

25

15

Late growing season

0

5

60

35

 Coarse roots are 2–5 mm.  Fine roots are Mg > K > Fe

This sequence partially reflects mineral solubility and partially the degree to which these elements participate in the formation of secondary minerals, such as phyllosilicate clays. Table 19.1 summarizes nutrient inputs from mineral weathering for forest ecosystems occurring in different climates and on different parent materials. Notice that the weathering of dolomitic parent materials (CaCO3 and MgCO3) beneath limber and bristlecone pines yields relatively large amounts of Ca compared to weathering of adamellite (SiO2:NaAlSi3O8:CaAl2Si2O8). The weathering of Ca-­and Mg-­bearing minerals represents an important input of these nutrients in many terrestrial ecosystems, satisfying 35% of Ca and Mg required for the annual growth of some temperate

Table  19.1  Rates of  mineral weathering for  forest ecosystems occurring in  different climates and  on ­different soil parent materials. Ecosystem

Parent material

Location

P

K

Ca

Mg

kg ha  y

−1 −1

Wet tropical

Alluvium

Venezuela

−

−

6

1

Douglas-­fir

Volcanic tuff

Oregon

−

2

47

12

Limber and

Dolomite

California

−

4

86

52

Bristlecone pines

Adamellite

California

−

8

17

2

Trembling aspen

Glacial till

Wisconsin

1

4

7

–

Northern hardwood

Glacial till

New Hampshire

13

7

21

4

Source: Data have been summarized from Boyle and Ek (1973), Fredriksen (1972), Hase and Foelster (1983), Likens et al. (1977), Marchand (1971), and Wood et al. (1984).

490  C h a p t e r 1 9   Nutrient Cycling

forests (Table 19.1; Likens et al. 1977). Similarly, the weathering Ca-­, Fe-­, and Al-­phosphate minerals provides an important source of P for plant growth. It had been thought that mineral weathering was not a source of N to plants, because few rocks contained N in their chemical structure. Recently, a global analysis of rock chemical composition has revealed that more N resides within them than previously thought. This realization suggests that N released from rock weathering could compose 8–26% of annual N inputs to terrestrial ecosystems, prior to the anthropogenic input of N from atmospheric deposition and fertilizer use (Houlton et al. 2008). Figure 19.2 summarizes the amounts of N contained in surface rock (Figure  19.2a), the amounts annually released through weathering (Figure 19.2b), and the percent contribution of annual N inputs supplied via mineral weathering. This summary has dramatically revised our thinking of how N enters terrestrial ecosystems.

ATMOSPHERIC DEPOSITION Over the past several decades, ecologists have become increasingly interested in understanding the geographic pattern and amount of nutrients entering terrestrial ecosystems from the atmosphere. In many portions of the Earth, human activity has altered rainfall chemistry and hence the atmospheric deposition of nutrients in dramatic ways. In the northeastern United States, central Europe and broadly across Asia, terrestrial and aquatic ecosystems receive enhanced inputs of NO3−, SO42−, and H+ (i.e., acidic deposition or acid rain) from fossil fuel burning. Both NO3− and SO42− contain essential plant nutrients, and elevated inputs of these ions have altered the rate and pattern by which N and S are cycled and stored within terrestrial ecosystems. Atmospheric deposition occurs through three processes: (i) wet deposition—­the addition of nutrients contained in rain or snow, (ii) dry deposition—­the direct deposition of atmospheric particles and gases to vegetation, soil, or water surfaces, and (iii) cloud deposition—­the input of small, nonprecipitating water droplets (in clouds and fog) to terrestrial surfaces (Fowler  1980; Lovett and Kinsmann 1990; Lovett 1994). Wet and dry deposition occur in all terrestrial ecosystems, but their importance as a source of nutrients varies greatly from place to place. By contrast, cloud deposition is generally restricted to coastal and mountainous regions that are frequently immersed in clouds or fog (Lovett 1994), like coastal giant redwood forests of northern California, United States. Wet deposition occurs when solid particles (0.2–2 mm diameter) and gases in the atmosphere dissolve in water droplets that fall to Earth as rain or snow (Lovett 1994). This process is controlled by several factors that vary markedly from one region to another and result in a broad range of nutrient inputs from wet deposition. In North America, continental-­scale patterns of wet deposition are relatively well understood (Figure 19.3). For example, anthropogenic emissions of N and S in the Midwest, Southeast, and Northeast result in widespread patterns of increased NO3− and SO42− deposition over the eastern United States. By contrast, ammonium (NH4+) and Ca2+ inputs from wet deposition are relatively low in the eastern United States but are elevated in the eastern Great Plains region and the Midwest. The plowing of agricultural fields and traffic on unpaved roads cause Ca2+ associated with dust particles to be carried eastward from the Great Plains region by the prevailing winds. Ammonium released into the air from heavily fertilized agricultural fields and volatilized from animal manure in concentrated feeding operations is transported eastward in a similar manner. Wet deposition entering forest ecosystems can pass through the canopy as throughfall or portions can flow down tree stems in stemflow. The chemical composition of wet deposition can be greatly altered by the “wash off” of dry material accumulated on plant surfaces, the leaching of nutrients from plant tissues, and the direct assimilation of nutrients into leaves. Potassium, Ca2+,

Nutrient Additions to Forest Ecosystems  491

(a)

Rock N reservoir (kg N m–3) 1.90 1.43

0.95 0.48

0.00

(b)

Chemical N flux (kg N ha–1 yr–1) 10.0+ 7.5

5.0 2.5

0.0

(c)

Increased modern N input from rock (%) 100+ 75

50 25

0 F I G U R E  1 9 .2   (a) The amount of N contained in surface minerals, (b) the annual release of N from minerals, and (c) the percent of annual ecosystem N inputs supplied by the weathering of N from surface minerals. Cool colors (blue) indicate low inputs and amounts, whereas warm colors (red) indicate the opposite. Source: Houlton et al. (2008) / American Association for the Advancement of Science.

492  C h a p t e r 1 9   Nutrient Cycling 5 6

5

5

10

10

20 20

20

18 14 18

14 10 10 14 14

20 14 20 20 2–

SO 4

5

–

10

NO3

10

6

1.5 2.5

2.5

1.0

1.5

1.5

1.5

1.5

3.5 2.5

2.5 2.5 2.5

+

NH 4

1.5

1.0

2.5 2.5 3.5 3.5

1.5

1.0 1.0

1.5 1.0 1.5

1.0

1.5

Ca

2+

1.5

2.0

1.5 1.0 2.0 1.5

F I G U R E  1 9 .3   Geographic patterns of sulfate (SO42−), nitrate (NO3−), ammonium (NH4+), and calcium (Ca2+) in wet deposition across the continental United States. Values are in kilograms per hectare per year (kg ha−1 y−1). Source: Lovett (1994). Reprinted with permission by the Ecological Society of America.

and Mg2+ are easily leached from leaves (Tukey 1970), and their concentrations in throughfall can be greater than that of the original wet deposition. This may be particularly important in tropical forests in which species-­specific differences in canopy leaching are related to the abundance of epiphytes (Schlesinger and Marks 1977). In some cases, however, forest canopies directly assimilate water-­soluble nutrients (i.e., N) from wet precipitation, thereby lowering their concentration in throughfall (Olson et al. 1981). Stemflow is significant in that it returns a relatively concentrated supply of nutrients directly to the base of the tree (Gersper and Holowaychuk 1971), where they may accumulate and be taken up by the tree. Determining the extent to which the “wash off” of dry material and leaching influence the nutrient concentration of throughfall and stemflow can be difficult. However, it appears that 85% of the SO42− in throughfall originates from the “wash off” of dry material deposited on leaves ­(Lindberg and Garten 1988), making the dry deposition of nutrients an important process in forest ecosystems. The high amount of leaf surface in forests makes them effective collectors of dry materials suspended in the atmosphere. Dry deposition is a complex process involving atmospheric chemistry, wind velocity, and canopy characteristics like leaf shape, orientation, spatial arrangement, and area (Lovett  1994). Solid particles (< 5 μm diameter) enter terrestrial ecosystems through gravitational sedimentation, the process by which the force of gravity pulls particles suspended in the atmosphere toward the Earth’s surface. Some particles are small enough to be kept aloft by wind turbulence and enter ­terrestrial ecosystems only when they impact the surface of vegetation. Even smaller dry particles and gases directly enter the plant through stomata and lenticels.

Nutrient Additions to Forest Ecosystems  493

The movement of atmospheric particles toward leaf surfaces is determined by the change in particle concentration from the atmosphere to the leaf surface and the resistance to flow along that path (Garland 1977). Atmospheric scientists estimate the input of particles to terrestrial ecosystems by determining their deposition velocity:



Deposition velocity

Rate of particle dryfall mg cm 2 of leat s Concentration in air mg cm

3

1

.



With knowledge of deposition velocity (cm sec−1), atmospheric concentration (mg cm−3), and leaf area (m2  m−2), atmospheric scientists estimate the amount of nutrient-­containing dry particles entering a particular forest ecosystem. Total dry deposition (mg m−2) is simply the mathematical product of atmospheric concentration, deposition velocity, and canopy leaf area. Dry deposition of nutrients is somewhat more variable than wet deposition for several reasons (Figure  19.3). The ratio of wet to dry deposition increases with distance from the source, because readily dry-­deposited materials are depleted from the air as it travels downwind (Ollinger et al. 1993). Second, differences in canopy structure and leaf morphology among trees cause nutrients in dry material to be captured with different efficiencies, depending on overstory composition. Nonetheless, dry deposition can bring significant amounts of nutrients into forest ecosystems relative to those contained in wet deposition. Figure 19.4 summarizes wet, dry, and cloud depositions for a series of forest, prairie, and agricultural ecosystems distributed across North America. Notice that the ratio of dry to wet deposition is highly variable among these ecosystems, reflecting differences in the distance to nutrient sources and canopy characteristics. The prairie ecosystem (AR) located downwind from Chicago, Illinois, United States receives relatively greater inputs of S and N in dry deposition than would be expected for an ecosystem in the Midwest (Figure 19.4) because it is relatively close to a major source of anthropogenic NO3− and SO42− production. Cloud deposition results when small, nonprecipitating water droplets containing dissolved nutrients come in contact with terrestrial surfaces, such as tree canopies. In contrast to wet deposition, tree canopies exert a substantial influence on the amount of nutrients entering forest ecosystems from cloud deposition. This process is a particularly important nutrient input in ecosystems that lie in coastal and mountainous regions (WF and CD in Figure 19.4), whereas in continental areas (e.g., Midwest), cloud deposition contributes relatively minor amounts of nutrients to terrestrial ecosystems. For example, relatively large quantities of SO42+, NO3−, and NH4+ enter spruce/fir ecosystems in the northeastern United States and red spruce ecosystems at high elevation in the Appalachian Mountains through cloud deposition (Figure  19.4). Over 50% of the atmospheric deposition of NH4+ in these ecosystems occurs through cloud deposition.

BIOLOGICAL FIXATION OF NITROGEN Biological inputs of N to forest ecosystems occur through symbiotic associations of soil bacteria and tree roots, free-­living bacteria in litter and mineral soil, as well as epiphytic lichens living on the external surfaces of trees and decaying wood (Melillo  1981; Waughman et  al.  1981; Boring et al. 1988). Several genera of soil bacteria, actinobacteria (filamentous bacteria), and cyanobacteria have the ability to reduce or “fix” atmospheric N2 into NH4+. Depending on the fixation process, NH4+ can be directly used by the microbial cell or it can be transported from the microorganism to the host plant as N-­containing organic compounds. The process of N2 fixation is catalyzed by the

494  C h a p t e r 1 9   Nutrient Cycling

nitrogenase enzyme system, which requires substantial quantities of energy (i.e., ATP). The ­following equation summarizes this process:



N 2 16 ATP 8 e

10 H

Nitrogenase

2 NH 4

H 2 16 ADP Pi .

The high energy cost (i.e., ATP) associated with N2 fixation has important ecological implications and gives rise to widely different rates of fixation depending on the fixation process and the microorganisms involved. Soil organic matter provides small amounts of energy for fixation by free-­ living soil bacteria, whereas carbohydrates from photosynthesis can fuel rapid rates of N2 fixation by symbiotic bacteria and actinobacteria inhabiting plant roots. As a result, quantities of N entering terrestrial ecosystems from N2 fixation are highly variable and range from as little as 1 kg N ha−1 y−1 for free-­living soil bacteria to 200 kg N ha−1  y−1 from symbiotic fixation in the roots of legumes, alder, and other higher plants (Cole and Rapp 1981). Rates of free-­living N2 fixation by free-­living organisms are low and equivalent to rates of atmospheric deposition (Figures 19.2 and 19.3) in most terrestrial ecosystems. Free-­living heterotrophic bacteria of the genera Azotobacter and Clostridium fix atmospheric N2 into forms that plants can assimilate after their cells die and cellular constituents are decomposed by other soil microorganisms. Because soil organic matter supplies the energy for free-­living fixation, this process is greatest in soils with relatively high organic matter contents (Granhall 1981). In the forests of the Pacific Northwest and northeastern United States, free-­living N2 fixation has been observed in decaying logs (Roskoski 1980; Silvester et al. 1982), a relatively rich-­energy substrate for heterotrophic bacteria. In most cases, the input of N to forest ecosystems from free-­living N2 fixation is relatively small, approximately 1–5 kg N ha−1 y−1 (Boring et al. 1988). Free-­living cyanobacteria (blue-­green algae) are common soil microorganisms that can, in some instances, be a source of N in some terrestrial ecosystems. Fixation by these photosynthetic microorganisms is associated with high-­light environments, and in most forests, fixation by cyanobacteria is limited by low light levels at the soil surface. However, algal crusts that form on the surface of some desert soils can have exceptionally high rates of N2 fixation (Rychert et al. 1978). Nitrogen-­fixing cyanobacteria also enter into mutualistic associations with some fungi to form N2­­­ fixing lichens that bring N into forest ecosystems. Nostoc, an N2-­fixing cyanobacteria, is one of two species of algae present in the lichen Lobaria oregana that inhabits the crowns of old-­growth Douglas-­fir. Lobaria is estimated to fix approximately 8–10 kg N ha−1 y−1, an amount comparable to the input of N via atmospheric deposition (see Figure 19.4). Nitrogen-­fixing actinobacteria and unicellular bacteria can infect the fine roots of certain higher plants and induce the formation of root nodules, which are the location of symbiotic N2 fixation. These small (< 5 mm), round, hollow structures are formed on individual roots and enclose large numbers of bacterial or actinobacterial cells, the agents of N2 fixation. Nitrogen fixed by the microbial symbiont is transferred to the plant, and carbohydrate supplied by the host in turn provides energy for the microbial fixation of N2 in the root nodule. Because the cost of N2 fixation is subsidized by a supply of carbohydrate (i.e., energy) from the plant, symbiotic fixation can bring substantial quantities of N into terrestrial ecosystems, far surpassing fixation by free-­living bacteria and cyanobacteria. Legumes are perhaps the most important plants with N2-­fixing bacteria inhabiting their roots. They occur as overstory species in both temperate and tropical forests. However, not all legumes have the ability to fix N2 in association with Rhizobium, their bacterial symbiont. Black locust, honey locust, acacias, and mesquite are trees of the legume family that occur in temperate forests; there are hundreds of similar species in the tropics. Also, many herbs and shrubs of this family inhabit the forest floor throughout much of the Earth, many becoming dominant following forest harvesting or fire. Soil bacteria belonging to the Rhizobium genus exclusively infect the roots of legumes and enter into a symbiotic N-­fixing relationship.

(a)

(b) NO3-N Deposition Rate (kg•ha–1•yr–1)

S Deposition Rate (kg•ha–1•yr–1)

20 40 30 20 10

16 12 8 4 0

0 WF

HF

SC

OR

CW

CD

PN

AR

PW

WF

TH

HF

SC

OR

CW

CD

PN

AR

PW

TH

(c) NH4-N Deposition Rate (kg•ha–1•yr–1)

10 Cloud 8

Dry Wet

6 4 2 0 WF

HF

SC

OR

CW

CD

PN

Code

Site Location

Vegetation

WF HF SC OR CW CD PN AR PW TH

Whiteface Mountain, New York Huntington Forest, New York State College, Pennsylvania Oak Ridge, Tennessee Coweeta, North Carolina Clingmans Dome, North Carolina Panola, Georgia Argonne, Illinois Pawnee site, Colorado Thompson site, Washington

Spruce/fir forest Beech/maple/birch forest Corn/wheat/soybeans Loblolly pine forest White pine forest Red spruce forest Meadow Meadow/forest Prairie Douglas-fir forest

AR

PW

TH

Elevation (m)

Annual Precipitation (cm)

1000 500 393 300 725 1740 212 229 1641 220

115 97 121 114 144 203 130 122 28 114

F I G U R E  19.4  Atmospheric deposition of S, NO3−-­N, and NH4+-­N in agricultural, prairie, and forest ecosystems in North America. The relative contribution of wet, dry, and cloud deposition is depicted. Note that cloud deposition is an important process in coastal areas and at high elevations. Source: Lovett (1994). Reprinted with permission of the Ecological Society of America.

496  C h a p t e r 1 9   Nutrient Cycling

In the southeastern United States, black locust readily establishes following a major disturbance, such as fire or clear-­cut harvest. Fixation in the root nodules of this leguminous tree is a substantial input of N to the forests of the southern Appalachians, particularly during the early to intermediate stages of ecosystem development (Waide et al. 1988). Four years following clear-­cut harvest, rates of N2 fixation equaled 48 kg N ha−1  y−1, a substantial input of N relative to atmospheric inputs (see Figures  19.2 and  19.3). Rates subsequently increased to 75 kg N ha−1  y−1 at 17 years following harvest, and after 38 years of ecosystem development rates declined to 33 kg N ha−1 y−1 (Waide et al. 1988). Compare these rates with mineral weathering and atmospheric deposition, which are orders of magnitude slower. Actinobacteria (filamentous bacteria) of the genus Frankia form a symbiotic relationship with the roots of Alnus, Ceanothus, Casuarina, Elaeagnus, Comptonia, and Myrica. This symbiotic relationship is particularly important in the Pacific Northwest, where N2 fixation by red alder and Frankia can substantially increase the amount of N entering forest ecosystems. One example comes from a comparison of 38-­year-­old red alder and Douglas-­fir stands that initially established adjacent to one another on the same soil type. Differences in N pools between them resulted directly from N2 fixation by the alder–Frankia symbiosis (Table 19.2). An additional 85 kg N ha−1 y−1 have accumulated under red alder, an input of N far exceeding atmospheric deposition (see Figure 19.3). Much of the N2 fixed by the red alder–Frankia resides in forest floor and mineral soil, where it is eventually released during organic matter decomposition. The shrub Ceanothus velutinus also occurs in the Pacific and Inland Northwest and enters into an N2-­fixing symbiosis with Frankia that can contribute as much as 100 kg N ha−1 y−1 following forest fire or harvest (Youngberg and Wollum 1976). Greater N availability resulting from N2 fixation on N-­poor soils can dramatically increase rates of biomass accumulation and nutrient cycling during ecosystem development. In the Hawaiian Islands, Myrica faya is an exotic N2-­fixing tree that has greatly altered the N cycle of native forests (Vitousek et al. 1987). This invasive, early-­successional species establishes on volcanic ash flows that were formerly colonized by Metrosideros polymorpha, a native, non-­N2 fixing tree. In the absence of M. faya, N accumulates in ash-­flow soils at a very slow rate, constraining rates of ecosystem development and productivity (approximately 5 kg N ha−1 y−1 from free-­living fixation and atmospheric deposition). Nitrogen fixation by M. faya brings an additional 18 kg N ha−1 y−1 into the ecosystem, increasing the annual rate of N accumulation by a factor of three. The N2 fixed by Myrica eventually enters the soil via leaf and root litter, and during decomposition, relatively greater amounts of N are released into soil solution where they are assimilated by plant roots. Table  19.2  The accumulation of  N in  forest ecosystems dominated by nitrogen-­ fixing (red alder) and  non‑nitrogen-­fixing (Douglas-­fir) tree species. The  overstory of  each ecosystem is 38 years old and differences in the amount of N contained in each ecosystem pool reflect the fixation of nitrogen by red alder. Ecosystem pool

Douglas-­ fir

Nitrogen kg N ha−1

Red alder

Annual accumulation kg N ha−1 y−1

Overstory

320

590

7.1

Understory

10

100

2.4

Forest floor

180

880

18.4

Mineral soil

3270

5450

57.4

Total ecosystem

3780

7020

85.3

Source: Cole and Rapp (1981) / Cambridge University Press. Reprinted from Dynamic Properties of Forest Ecosystems, © 1981 by Cambridge University Press. Reprinted with the permission of Cambridge University Press.

Nutrient Cycling within Forest Ecosystems  497

NUTRIENT CYCLING WITHIN FOREST ECOSYSTEMS

Nutrients entering forest ecosystems from mineral weathering, atmospheric deposiAtmospheric Deposition tion, and biological fixation can enter soil solution where they are absorbed by plant roots. Within the plant, nutrients taken up from soil solution participate in a wide array of physiological processes, and in Translocation some cases, growth-­limiting nutrients are Aboveground Litter Fixation removed (i.e., translocated) prior to the Production shedding of some plant tissues. The proGaseous Loss duction of plant litter above and below the soil surface eventually returns nutrients to forest floor and mineral soil, where they are released from litter by soil microorganisms. During the process of litter decomposition, soil microorganisms incorporate organiOrganic Matter Forest Floor and Belowground cally bound nutrients (and C) into their Mineral Soil Litter Production biomass and release excess nutrients into Nutrient Uptake soil solution where they can again be taken Decomposition and Assimilation and Mineralization up by plant roots. Thus, the rate at which Nutrients in nutrients flow within forest ecosystems is Soil Solution Mineral controlled by the physiological activities of Weathering Leaching plants and soil microorganisms, and their requirement for growth-­limiting nutrients. In the sections that follow, we trace the flow of nutrients within forest ecosystems. The cycling of nutrients within ecosystems occurs through: (i) root nutrient absorption via the processes of uptake and assimilation, (ii) nutrient allocation to biomass construction and maintenance, (iii) nutrient translocation from senescent tissue, (iv) the return of nutrients in aboveground and belowground litter, and (v) the microbially mediated release of inorganic nutrients into soils solution (i.e., mineralization) during organic matter decomposition. These processes are highlighted in the reduction of Figure  19.1, which is located above. Central to this discussion is the ability of plants to forage for nutrients in soil solution, a process mediated by physiological and morphological adaptations of plant roots. Also of importance is the growth dynamics of soil microorganisms during litter decomposition, because they control the release of organically bound nutrients from plant litter and soil organic matter. As you will learn later in this chapter, soil microorganisms are the “gate keepers” of N availability within forest ecosystems.

NUTRIENT TRANSPORT TO ROOTS Nutrients must diffuse to the root surface from soil solution before they can be taken up and ­incorporated into biologically active compounds. The uptake of nutrients by plant roots is initially constrained by rates of organic matter decomposition, mineral solubility, cation/anion exchange reactions, as well as diffusion through soil solution. Although these processes control nutrient availability in soil, the fact that nutrients are present in soil solution or on exchange sites does not ensure uptake by plant roots. In the paragraphs that follow, we discuss the processes controlling the flow of nutrients to the root surface, uptake from soil solution, and their incorporation into biologically active compounds within the plant.

498  C h a p t e r 1 9   Nutrient Cycling

Nutrients in soil solution move toward root surfaces in response to two processes: mass flow and diffusion. Mass flow is a passive process in which ions move with the flow of water for transpiration; they are physically carried along from soil solution to the root surface. Diffusion occurs when ions move from a region of high concentration to one of low concentration, for example, from soil solution to the root surface. In most soils, the concentration of nutrients in solution is low enough that mass flow does not meet the demand of actively growing plants. Rapid uptake by plant roots also causes some nutrients to be virtually absent near their surface, thereby forming zones of depletion around individual roots. Consequently, the supply of nutrients to roots is constrained by the rate at which ions diffuse toward root surfaces from areas of higher concentration in soil solution (Nye 1977). Nutrients such as PO43−, K+, and NH4+ diffuse slowly in water and rapid uptake depletes their concentration near root surfaces so that concentrations increase as the distance from the root surface increases. Although NO3− is very mobile in soil solution and diffuses rapidly toward roots, a high demand by most plants maintains low NO3− concentrations throughout soil solution. Because of its high rate of diffusion, mass flow may meet the NO3− demand of some forest trees. The size of the zone of depletion surrounding an individual root is directly proportional to its radius and the rate at which a particular ion diffuses in water. The radius of the zone of depletion (Rdepletion in cm) for any ion is



Rdepletion

a 2 Dt ,

where a is the root radius (cm), D is the diffusion coefficient for an ion (cm2 sec−1), and t is time in seconds (Nye and Tinker  1977). One also can compare root size (cm3) to the zone of depletion (cm3) it creates using the following expression:



a 2 Dt

2

/a 2

Given this relationship, smaller roots exploit a greater volume of soil than larger roots (Fitter 1987). Recall that large roots have higher construction costs than small roots, indicating that plants “invest” relatively greater amounts of photosynthate into the production of large-­diameter roots (Chapter  18). By forming small-­diameter roots, plants increase the volume of soil exploited for water and nutrients, while reducing the amount of photosynthate allocated to root production. It is likely that this is why roots generally less than 1 mm in diameter are the nutrient-­ and ­water-­absorbing organs of forest trees. Once nutrients arrive at the root surface, they must be taken up from soil solution and incorporated into biologically active compounds. In some soils, the supply of some nutrients is excessive, and they are actively excluded from uptake. It is not uncommon to observe accumulations of CaCO3 adjacent to the roots of desert shrubs growing on highly calcareous soils (Klappa 1980). In most instances, however, nutrient supply is low (especially for N and P), and plants actively forage for nutrients in soil. Nutrient foraging occurs through the physiological mechanisms of nutrient uptake and through morphological changes in root architecture that increase the absorptive area of plant root systems.

NUTRIENT UPTAKE AND ASSIMILATION BY ROOTS Nutrients are incorporated into biologically active compounds (i.e., amino acids, nucleic acids) in a two-­step process consisting of uptake and assimilation. Uptake is the physiological process by which nutrients in soil solution are actively transported across cell membranes into roots. ­Assimilation occurs when inorganic nutrients transported into plant cells are biochemically

Nutrient Cycling within Forest Ecosystems  499

incorporated into organic compounds such as amino acids, nucleic acids, lipids, or other biologically active compounds. High rates of enzymatic uptake are one physiological means by which plants can maximize acquisition of growth-­limiting nutrients. For example, the uptake of nutrients from soil solution is mediated by enzymes associated with the membranes of very-­fine roots (< 0.5 mm diameter). The rate of enzymatic activity mediating nutrient uptake increases with increasing nutrient concentrations in soil solution until the capacity of the enzyme system is saturated. Uptake capacity for nutrients varies widely among tree species, and this process can be highly responsive to soil temperature (Figure 19.5). In a comparison of trees of the taiga, rapidly growing species such as

6

3

Trembling Aspen

4 3

2

2 1 0

1 0

3

10

20 0

Paper Birch

2

2

1

1

0

0

10

20

0

0

10 Green Alder

0

10

Phosphate concentration 4 Nitrate absorption (µmol·h–1·g–1)

10° 20°

20

Ammonium absorption (µmol·h–1·g–1)

5 Phosphate absorption (µmol·h–1·g–1)

60

Balsam Popular

2 1 4000

Paper Birch

0 4

3

3

2

2

1

1

0

0

600

30

30

20

20

10

10

1200

0

2000

4000

Paper Birch

0

30

20

20

10

10 0

2000

4000

0

40

30

4

1

4

40

Trembling Aspen

0

2000

4000

Green Alder

0

Ammonium concentration

2

2000

40

40

(µmol·L–1)

3

0

50

0

3

0

60

50

0

20

Balsam Poplar

Balsam Poplar

2000

4000

(µmol·L–1)

Trembling Aspen

0

2000

4000

Green Alder

0

0 2000 4000 Nitrate concentration (µmol·L–1)

F I G U R E  1 9 .5   Nutrient uptake by four taiga trees in response to nutrient concentration and soil temperature. Uptake rates of NH4+, NO3−, and PO43− were measured on excised roots of seedling growing under laboratory conditions. Source: Chapin et al. (1986). Reprinted from Oecologia, © Springer-­Verlag Berlin, Heidelberg 1986. Reprinted with permission of Springer-­Verlag, New York, Inc.

500  C h a p t e r 1 9   Nutrient Cycling

balsam poplar and trembling aspen have relatively high uptake capacities for NH4+, NO3−, and PO43−, compared to the slower-­growing paper birch and green alder (Chapin et al. 1986). Balsam poplar and trembling aspen also occur on relatively warm soils in the taiga and exhibit rapid increases in nutrient uptake with rising soil temperature (Figure 19.5). Although the productivity of many temperate forests is constrained by soil N availability, we understand relatively little regarding the ecological importance of NH4+ versus NO3− uptake by overstory trees. Ammonium (NH4+) often is the dominant form of N in forest soils, and some trees exhibit a physiological preference for NH4+ over NO3−. This relationship can be observed in Figure  19.5 for taiga trees, and it also has been documented for several overstory trees in temperate forests (Figure 19.6). In the fine roots of Douglas-­fir and sugar maple, rates of NH4+ uptake far exceed rates of NO3− uptake, suggesting that these widely distributed trees have a

35

30

30

25

25

20

20 15NH+ 4 15NO– 3

15 10

15 10

5 0

5 0

400

200

600

800

µmol 15NO–3·g–1·h–1

35

+

µmol 15NH 4·g–1·h–1

(a)

0

1000

µM 15NH+4 or 15NO–3 10

10

8

8

6

6

4

4

2

2

µmol 15NO–3·g–1·h–1

µmol 15NH+4·g–1·h–1

(b)

0 0

200

400

600

µM 15NH+4 or

15NO– 3

800

1000

F I G U R E  19 .6   Ammonium and NO3− uptake in the fine roots of (a) sugar maple and (b) Douglas-­fir. At any

given concentration, rates of NH4+ uptake exceed those of NO3−, suggesting that these overstory species have a physiological preference for NH4+ over NO3−. Source: Kamminga-­Van Wijk and Prins (1993) / Springer Nature. After Kamminga-­Van Wijk and Prins 1993 and Rothstein et al. 1996. Panel (a) reprinted from Oecologia, © Springer-­Verlag Berlin, Heidelberg 1996. Reprinted with permission of Springer-­Verlag, New York, Inc. Panel (b) reprinted from Plant and Soil, 1993, Vol. 151, “The kinetics of NH4+ and NO3− uptake by Douglas-­fir from simple N-­solutions and from solutions containing both NH4+ and NO3−” by C. Kamminga-­Van Wijk and H. Prins, pp. 91–96, © 1993 by Kluwer Academic Publishers. Reprinted with kind permission from Kluwer Academic Publishers.

Nutrient Cycling within Forest Ecosystems  501

physiological preference for NH4+, especially for sugar maple which has very low maximum rates of NO3− uptake. Western hemlock and jack pine commonly occur on acidic soils with extremely low NO3− availability. These overstory species also have a limited capacity for NO3− uptake and appear to satisfy their demand for N through rapid NH4+ uptake (Lavoie et al. 1992; Knoepp et al. 1993). In jack pine, NO3− uptake enzymes are present at very low levels in fine roots (Lavoie et al. 1992), and when seedlings are supplied with NH4+ or NO3− as a sole source of N, those supplied with NH4+ attain twice the biomass of seedling grown solely on NO3−. (Lavoie et al. 1992). This observation suggests that jack pine has a low physiological capacity for NO3− uptake, even when concentrations in soil solution are relatively high. The inability of this species to use NO3− suggests that it would be a poor competitor in NO3−-­rich soils, relative to plants that have rapid rates of NO3− uptake and assimilation. There are several reasons why NH4+ uptake is more rapid than NO3− uptake in many forest trees. First, NH4+ is often the dominant form of N in soil solution and small amounts can strongly inhibit enzymes involved with NO3− uptake and assimilation. Second, plants must reduce NO3− to NH4+ before it can be assimilated into biologically active compounds, a process that requires substantial amounts of energy (ATP). When this reaction occurs in leaves, it is subsidized by energy from the light-­harvesting reactions of photosynthesis. However, many woody plants have the ability to reduce NO3− in roots, which requires the transport of reducing compounds from the leaf. Most often, plants with rapid rates of NO3− uptake and assimilation occur in high-­light habitats in which excess energy from photosynthesis can be allocated to NO3− reduction. A second physiological mechanism for maximizing the uptake of limiting nutrients is via the production of root enzymes that release nutrients from soil organic matter. This is particularly true of PO43−, which diffuses slowly in soil solution. Phosphatases are enzymes produced by plants and microorganisms that act upon soil organic matter to yield PO43− from organically bound P. The release of these enzymes into soil appears to be an important mechanism for increasing PO43− supply to roots in PO43−-­poor soils. Phosphatase activity associated with the roots of arctic tundra plants can supply up to 65% of their annual PO43− demand (Kroehler and Linkins 1988).

ROOT ARCHITECTURE, MYCORRHIZAE, AND NUTRIENT ACQUISITION Root Architecture  In addition to physiological mechanisms, plants also forage for nutrients by altering root architecture and growth. Ecologists have become increasingly interested in the structure and function of plant root systems, because plants appear to alter root branching patterns, or architecture, in response to soil nutrient availability (Fitter 1987; Fitter and Stickland 1991). ­Several analyses suggest that fine-­root systems consisting of only a main axis and primarily laterals ­(herringbone architecture; Figure 19.7) are the most efficient at nutrient acquisition (Fitter and Stickland  1991; Berntson and Woodward  1992). The herringbone arrangement maximizes the ­volume of soil exploited by roots, while minimizing the overlap between the zone of depletion created by adjacent fine roots; however, the construction cost of this type of root system is high. It requires the production of larger diameter roots (e.g., higher construction), each of which transports a greater proportion of the total flow of water and nutrients into the aboveground portion of the plant (Figure  19.7). By contrast, the mean root diameter of a dichotomously branched root system is much smaller than that of a herringbone architecture, resulting in relatively lower construction costs. This spatial arrangement of roots allows the depletion zones of adjacent roots to overlap, and thus it is less efficient at foraging for nutrients than the herringbone architecture. Differences in the cost of constructing and maintaining these root systems, and the efficiency with which they explore soil for nutrients, suggest that root-­system architecture should vary in response to soil nutrient availability. Plants growing in nutrient-­poor soils should allocate relatively more photosynthate to a highly efficient root system for nutrient uptake (i.e., herringbone

502  C h a p t e r 1 9   Nutrient Cycling F I G U R E   19. 7  Examples of plant root systems with different degrees of branching. Source: Fitter (1987) / John Wiley & Sons.

Herringbone

Intermediate

Dichotomous

branching pattern), whereas those growing in nutrient-­rich soil should invest relatively less C into a dichotomously branched root architecture. These predictions hold for different herbaceous plants (dicots) from nutrient-­rich and nutrient-­poor soils, and also hold for the same species of herbaceous plants grown in soils of different fertility (Fitter and Stickland  1991; Berntson and Woodward 1992; Taub and Goldberg 1996). The fine-­root system of forest trees appears to respond in a similar manner, but unfortunately, few studies have focused on these organisms. The fine roots of pin cherry rapidly proliferate within experimental patches (125 cm3) of N-­rich soil by increasing rates of production and the degree of branching, that is, their fine-­root system becomes more dichotomous (Pregitzer et al. 1992). Although there are clear differences in the construction cost of fine-­root systems, it is also important to consider the maintenance cost of these structures, because maintenance costs could exceed the cost of their construction. This is especially true for fine roots with high N concentrations (enzymes for ion uptake), but, unfortunately, we presently do not have sufficient observations to understand the relationships among construction costs, maintenance costs, and fine-­root longevity for a large number contrasting plant species. Importantly, we still are learning how form and function of plant root systems are related, as well as how this might be coordinated with aboveground plant form and function (i.e., canopy geometry, maximum rates of net photosynthesis). Recent global analyses of plant root systems indicate that plants effective in nutrient-­rich soil have a lower specific root length (i.e., smaller diameter per unit length) and higher N concentration than plants which are good competitors for nutrients in nutrient-­poor soils (Weigelt et al. 2021). These root traits correspond to aboveground form and function, wherein leaves of plants in nutrient-­rich soil are thinner, have high N concentrations, and more rapid rates of maximum net photosynthesis than those from N-­poor soils. Like leaves, fine roots with high N concentrations likely have high maintenance costs to fuel the rapid uptake of nutrients from soil solution. Although this is an intriguing parallel, we still lack a ­comprehensive understanding of how fine root form (architecture) and function (ion uptake) are coordinated among forest trees or any other plant species. This remains an important gap in our understanding of the belowground physiology and natural history of plants as well as how they are coordinated, or not, with the aboveground plant attributes.

Mycorrhizae  Forest trees increase their ability to forage for nutrients in soil by entering into a symbiotic relationship with soil fungi. This form of mutualism, termed a mycorrhiza or literally a root-­fungus, increases the volume of soil exploited by an individual plant. Mycorrhizal fungi are important for the successful establishment of some tree seedlings. Mycorrhizae also facilitate an important nutrient cycling process—­plant nutrient uptake and, to some extent, the decay of soil organic matter. In this section, we further explore the extent to which mycorrhizal fungi increase the ability of plant roots to forage for nutrients within soil. Root hairs are not common in nature, because tree roots are almost universally colonized by mycorrhizal fungi. In pines, spruces, firs (and all other genera of the Pinaceae), birches, beeches,

Nutrient Cycling within Forest Ecosystems  503 F I G U R E   1 9 . 8   Shown are the manner in which the hyphae of ectomycorrhizal fungi (ECM) and arbuscular mycorrhizae (AM) fungi penetrate plant roots. A notable difference between is the fact that ECM hyphae grow throughout root cell wall, whereas AM hyphae oppress plant cell membranes and form arbuscules, which are the site of photosynthate and nutrient exchange.

Ectomycorrhiza

Arbuscular mycorrhiza

Hyphopodium

Mycelium

Mantle Arbuscules

Hartig net

Spore

oak, basswoods, and willows, ectomycorrhizal fungi (ECM; ecto meaning outside) form a sheath or mantle surrounding fine roots giving them a characteristic swollen appearance (Figure  19.8). Fungal hyphae penetrate the space between the outer cortical cells, but do not enter individual root cells. Over 2400 species of fungi are known to form ECM on North American trees (Marx and ­Beattie  1977). A less conspicuous group, arbuscular mycorrhizae (AM), forms no sheath, but individual hyphae grow within and between epidermal and cortical cells of roots. Many plant species form AM, including cultivated crops, grasses, and most tropical tree species (Janos 1987). Temperate species forming AM include redwood and many hardwoods: maples, ashes, tuliptree, sweet gum, sycamore, black walnut, and black cherry (Marx and Beattie  1977; Harley and Smith 1983). Ectomycorrhizae and AM are effective accumulators of nutrients (and water) because they increase the absorbing surface area of tree roots, thereby allowing the plant to forage for nutrients in a greater volume of soil. Recall that small fine roots (approximately 0.1 mm in diameter) more effectively exploit soil for nutrients than larger roots. Fungal hyphae (10 μm in diameter) are much smaller than fine roots, which further increases the ability of plants to exploit soil for nutrients. Additionally, the hyphae of mycorrhizae can extend great distances into the forest floor and mineral soil. For example, 1 mm3 of soil can contain 4 m of ECM hyphae that transport water and nutrients back to the host plant. In a 450-­year-­old Douglas-­fir ecosystem in Oregon, 5000 kg ha−1 of ECM occur in the surface soil (0–10 cm), composing over 11% of the total root biomass (Trappe and Fogel  1977). Moreover, a single Douglas-­fir–Cenococcum ectomycorrhizae can form 200– 2000 individual hyphae, some of which extend over 2 m into soil and form more than 120 lateral branches or fusions with other hyphae. As compared to non-­mycorrhizal roots, those infected with mycorrhizal fungi tend to be more metabolically active, exhibiting rapid rates of respiration and ion uptake. Mycorrhizae are widely known to increase the PO43− uptake of forest trees growing in P-­poor soils; however,

504  C h a p t e r 1 9   Nutrient Cycling

mycorrhizae also facilitate the uptake of other plant nutrients such as N (Bowen and Smith 1981). The mycorrhizae formed by the fungus Paxillus involutus are able to take up substantial amounts of NH4+ from soil solution, assimilate the NH4+ into amino acids, and transfer the newly formed amino acids to European beech (Finlay et  al.  1989). Paxillus involutus also is able to assimilate NO3−, but it does so at a substantially lower rate. In addition to the direct uptake of ions from soil solution, mycorrhizae have the ability to produce extracellular enzymes which facilitate organic matter decomposition (Antibus et al. 1981; Dodd et al. 1987) and also produce organic acids which release plant nutrients from soil minerals (Bolan et al. 1984). Ectomycorrhizae and AM fungi differ from one another in many important ways that have implications for plant nutrient supply. Foremost, AM fungi are evolutionary ancient and are thought to have mediated the transition of aquatic plants to those on land. They represent an evolutionary narrow group of fungi (i.e., monophyletic), which assist plants with the acquisition of both water and phosphorus. These organisms produce phosphatase enzymes which release PO43− from organic matter, as well as organic acids that solubilize PO43− from soil minerals. By contrast, ECM have evolved more recently and have arisen from multiple saprotrophic fungal ancestors (i.e., polypheletic; approximately 80 independent times). While some ECM have retained genes that encode enzymes mediating organic matter decay, others have lost them during their evolution into root symbionts (Pellitier and Zak  2018). Isotopic evidence suggests that ECM who have retained genes for saprotrophic function can modify soil organic matter and provide the N contained within it to a host plant (Pellitier et al. 2021). This appears particularly important where soil N supply is slow and large quantities of N are locked with soil organic matter. It is interesting to note that a large number of boreal trees associate with ECM whose genomes retain genes with decay function, likely a response to slow rates of organic matter decay in these cold ecosystems. Because mycorrhizal fungi are heterotrophic microorganisms, they require a supply of ­carbohydrate for growth, for uptake of nutrients from soil, and for translocation of nutrients to the host plant. The cost of forming mycorrhizae is not inconsequential. Carbohydrate formed by the plant that could otherwise be used for growth, maintenance, or other physiological functions must be transferred to the mycorrhizal fungi in return for a greater nutrient supply. In a Pacific silver fir ecosystem, ECM compose only 1% of total ecosystem biomass, but their growth and maintenance consume over 15% of net primary production (Vogt et al. 1983). Because of the high cost of forming mycorrhizae, plants growing in nutrient-­poor soil tend to form more mycorrhizae than those growing in nutrient rich-­soils, because they are more dependent on them for nutrient acquisition.

PLANT LITTER AND THE RETURN OF NUTRIENTS TO FOREST FLOOR AND SOIL Plant litter production above and below the soil surface directly controls the amount of nutrients returned to the forest floor and mineral soil, and therefore constitutes important processes controlling the cycling of nutrients within forest ecosystems. Nutrients absorbed by roots and mycorrhizae are used for a wide variety of physiological functions, including the growth of new tissue, the maintenance of existing tissue, storage, and the production of defense compounds—­fates similar to that of photosynthetically fixed CO2 (Chapter 18). Prior to abscission, both coniferous and deciduous trees withdraw nutrients from leaf tissue and transport them into the nearby branches for storage. These materials are removed from storage when growth commences in spring and are used to form new leaves and construct a canopy. However, it is not clear whether nutrients are withdrawn from fine roots prior to their death. Consequently, both the production of litter and the concentration of nutrients contained within it control the rate at which nutrients absorbed by plants are returned to the soil from which they came. In the paragraphs that follow, we explore geographic patterns of aboveground and belowground plant litter production and the extent to which plants withdraw nutrients prior to leaf abscission, a process controlling the nutrient concentration of leaf litter.

Nutrient Cycling within Forest Ecosystems  505

Leaf and Root Litter Production  On a global basis, leaves account for approximately 60–

75% of total aboveground litterfall in forest ecosystems; the remaining proportion is composed of woody material (approximately 30%) and reproductive structures (1–20%). In Chapter 18, we found that global patterns of aboveground net primary production (ANPP) were strongly related to climate, particularly temperature and actual evapotranspiration. Leaf litter production in forest ecosystems also is strongly related to global patterns of climate, because of the aforementioned relationship. In Figure 19.9, leaf litter production is low at high latitudes where short growing seasons limit plant growth; it increases toward the equator where plant growth can occur throughout the entire year. Also notice that substantial variation occurs at any particular latitude (Figure 19.9). This regional-­scale variability in leaf litter production undoubtedly results from the modification of climate by physiography (i.e., slope and aspect), differences in soil water and nutrient availability, or disturbance. Recall that leaf biomass (or area) places the upper limit on the amount of photosynthetically fixed CO2 available for NPP (Chapter 18). As a result, leaf biomass and ANPP are positively related across a wide array of terrestrial ecosystems, including deserts, grasslands, and forests (Webb et al. 1983). Because leaves are temporary plant tissues, even in evergreen coniferous forests, one would expect that patterns of leaf litterfall should be related to ANPP. The solid line in Figure 19.10 indicates a 1:1 relationship between ANPP and leaf litter production, wherein all ANPP is ­allocated to leaf production. At low values of ANPP (i.e., < 7.5 Mg ha−1 y−1), note that leaf litter production represents a large proportion of ANPP. At higher rates of ANPP, leaf litter constitutes a smaller proportion, reflecting a greater allocation of photosynthate to other plant tissues and functions. Although we have a clear understanding of global patterns of leaf litterfall, our knowledge of the production of plant litter below the soil surface is incomplete. It is clear, however, that the death of fine roots represents a significant proportion of NPP in many forest ecosystems, often exceeding leaf litter production. Notice that fine-­root litter represents 40–330% of leaf litter in the temperate coniferous and deciduous forests summarized in Table 19.3. In the Pacific Northwest, the production and death of roots can account for 59–67% of NPP in Pacific silver fir ecosystems, whereas leaf litter production ranges from 6 to 8% of NPP (Vogt et al. 1983). Fine roots often have nutrient contents (N and P) greater than or equivalent to leaf litter, suggesting that fine-­root litter represents a substantial transfer of nutrients from plant tissue to forest floor and mineral soil.

18 Total litter production (Mg·ha–1·y–1)

F I G U R E  1 9 .9   Global patterns of leaf litter production in forest ecosystems. Source: Modified from Bray and Gorham (1964) and O’Neill and DeAngelis 1981. Modified from Advances in Ecological Research, ©, 1964 by Academic Press, Ltd. London. Reprinted with permission of Academic Press, Ltd. London and from Dynamic Properties of Forest Ecosystems, © Cambridge University Press 1981. Reprinted with permission of Cambridge University Press.

16 14 12 10 8 6 4 2 0

0

10

20

30

40

Latitude (°N)

50

60

70

506  C h a p t e r 1 9   Nutrient Cycling F I G U R E   19. 10  The relationship between aboveground net primary production (ANPP) and aboveground litter production in temperate and boreal forest ecosystems. The solid line indicates a 1:1 relationship between ANPP and the production of aboveground litter. At low levels of ANPP, virtually all production is allocated to leaves, and the departure from the 1:1 line at high levels of ANPP indicates a greater allocation of production to non-­photosynthetic tissues. Source: After O’Neill and DeAngelis (1981) / Cambridge University Press. Reprinted from Dynamic Properties of Forest Ecosystems, © Cambridge University Press 1981. Reprinted with permission of Cambridge University Press.

Total litterfall (Mg·ha–1·y–1)

20

15

10

5

0

0

5

10

15

20

Aboveground net primary production (Mg·ha–1·y–1)

Table 19.3  Leaf litter and root litter production in forest ecosystems from different geographic locations in temperate North America. The ratios of fine-­root litter to leaf litter indicate that the production of fine-­ root litter in forest ecosystems ranges from 40 to 330% of leaf litter. Litter Ecosystem

Leaf

Mg ha−1 y−1

Fine-­root

Root:leaf

Temperate coniferous Pacific silver fir

1.5–2.2

4.4–7.4

2.9–3.3

Douglas-­fir

2.8–3.0

2.4–2.7

0.9

Scots pine

2.1

2.0

1.0

White pine

2.9

2.6

0.9

Mixed pine

3.1

2.6

0.8

White spruce

2.7

1.6

0.6

Red pine

2.5–5.3

1.3–2.0

0.4–0.5

Black oak

4.1

5.9

1.4

Northern red oak

4.2

5.2

1.2

White oak

3.6

4.1

1.1

Sugar maple

2.9

4.0

1.4

Paper birch

2.8

3.2

1.1

Mixed hardwoods

4.4

2.7

0.6

American beech

3.2

1.3

0.4

Temperate deciduous

Source: Vogt et  al. (1986) / with permission of Elsevier and Nadelhoffer et  al.  1985. Modified from Advances in Ecological Research, © 1986 by Academic Press, Inc. Reprinted with permission of Academic Press, Inc. and from Ecology by the permission of the Ecological Society of America.

Nutrient Cycling within Forest Ecosystems  507

Recall that the return of nutrients to the forest floor and mineral soil is controlled by the production of plant litter and its nutrient concentration. From the previous paragraphs, it should be clear that forest ecosystems dramatically differ in the production of litter, both above and below the soil surface. In the following paragraphs, we discuss the withdrawal, or retranslocation, of nutrients prior to the shedding of leaf litter. This process directly influences the concentration (i.e., percent) of nutrients in litter, which together with litter production controls the total return of nutrients in plant litterfall.

Nutrient Retranslocation  The biochemical CO2-­fixing machinery of leaves requires substan-

tial quantities of N and P to form photosynthetic proteins, as do the nutrient uptake and assimilation mechanisms of fine roots. As a result, green leaves and live fine roots contain relatively high concentrations of nutrients compared to other plant tissues, such as stems and branches. Because low quantities of N and P in soil often limit plant growth, plants have evolved mechanisms to conserve growth-­limiting nutrients within their tissues. As noted earlier in the text, in both coniferous and deciduous trees substantial quantities of N and P (and other nutrients) are retranslocated from leaves prior to abscission and are mobilized into storage. Much of the N and P removed from leaves resides within adjacent branches, stem, and large structural roots, where they can be remobilized to meet the demand of newly forming leaves when growth resumes. However, the extent to which forest trees retranslocate nutrients from senescent leaves is highly variable (Killingbeck 1996); it changes from year to year within an individual species and changes among species from different habitats. The retranslocation of many nutrients reflects their availability in soil and their demand within the plant for subsequent growth. In Brazilian rainforest trees, 17–73% of N and 41–83% of P are translocated from leaves prior to abscission. Although retranslocation efficiency (i.e., percent removed) varies among rainforest species, they all reduce P to similar concentrations in abscised leaves (i.e., 0.04–0.06% P; Scott et  al.  1992). The retranslocation of N is not as complete and relatively large amounts of N (>1.0%) remain in abscised leaves (Scott et al. 1992). Differences in the amount of P and N remaining in abscised leaves of rainforest trees reflect the fact that P availability in soil, not N, limits the productivity of wet tropical forests (Vitousek 1984). The extent to which forest trees retranslocate nutrients from senescing leaves directly influences the nutrient-­use efficiency of litter, which is defined by the following expression:



Nutrient use-efficiency

Dry weight of leaf litterfall kg ha 1y

1

Nutrient content of leaf litterfall kg ha 1y

1

.



Plants that retranslocate large quantities of nutrients prior to leaf abscission have high ­nutrient-use efficiencies and return relatively few nutrients (per unit biomass) to the forest floor in litterfall. On a global basis, there is a remarkable relationship between the N-­ and P-­use efficiency and nutrient availability in tropical, temperate, and boreal forests. In Figure 19.11, the amount of nutrients returned in litterfall (x-­axis) is used as a surrogate for nutrient availability—­large amounts of nutrients residing in litterfall occurs where their availability in soil is high. Notice that N-­use efficiency is inversely related to the amount of N contained in litterfall, with the exception of tropical forests (Figure 19.9a). The inverse relationship between N-­use efficiency and litterfall N in temperate coniferous and deciduous forests indicates that N-­use efficiency is high in ecosystems in which relatively small quantities of N are annually cycled (i.e., those with low litterfall N contents). In tropical ecosystems, litterfall N varies widely (10–180 kg N ha−1 y−1), but N-­use efficiency is relatively constant (approximately 80; Figure  19.11a). The fact that N-­use efficiency changes ­little, whereas the amount of N annually cycled to the forest floor varies by a factor of three,

508  C h a p t e r 1 9   Nutrient Cycling

(a)

Tropical

368 N-Use Efficiency Dry weight:N (ratio of litterfall)

Coniferous 280

Deciduous

240

Mediterranean Shrublands

200

N-Fixing Species

160 120 80

223 224

40 0 0

20

40

60

80

100

N in litterfall (kg·

(b)

5076

120

·

ha–1

140

180

200

y–1)

4548

4000 P-Use Efficiency Dry weight:P (ratio of litterfall)

3600 3200 2800 2400 2000 1600 1200 800

12 14

400 0

0

1

2

3

4

5

P in litterfall (kg·

6

·

ha–1

7

8

9

10

y–1)

F I G U R E  1 9 .1 1   The relationship between plant nutrient-­use efficiency and the amount of nutrients returned in leaf litterfall for tropical, coniferous, temperate deciduous, and Mediterranean shrublands. Forests dominated by N-­fixing plants are denoted with a diamond. Plant nutrient-­use efficiency is estimated as the ratio of litterfall mass to its (a) N or (b) P content. The amount of N and P returned in litterfall (kg N or P ha−1 y−1) is a relative index of soil N or P availability. Source: After Vitousek (1982) / John Wiley & Sons. Reprinted with permission from American Naturalist, © 1982 by the University of Chicago. Reprinted with permission of the Chicago University Press.

indicates that N availability does not influence the N-­use efficiency of tropical forest trees. In contrast to this pattern, P-­use efficiency and the amount of P returned in litterfall are inversely related in tropical forests (Figure 19.9b). This pattern suggests that P translocation from senescent leaves is high (i.e., high P-­use efficiency) where P availability is relatively low within the ecosystem. Also notice that there is a weak relationship between P-­use efficiency and litterfall P in other forest ecosystems. These patterns are consistent with the observation that N generally limits the productivity of boreal and temperate forests, whereas a limited supply of P constrains productivity in wet tropical forests.

Nutrient Cycling within Forest Ecosystems  509

Taken together, these observations suggest that limited supplies of N or P result in greater translocation prior to leaf abscission, thereby increasing N-­ or P-­nutrient-­use efficiency of leaf litter. This is likely an evolved mechanism that permits plants growing in low N or P soils to conserve these growth-­limiting nutrients. Forest trees that conserve nutrients through a high ­nutrient-­use efficiency should have a competitive advantage on nutrient-­poor sites, and it is not uncommon to observe species replacement along gradients of nutrient availability (Pastor et al. 1984; Zak and Pregitzer 1990). Individual species also are able to alter nutrient-­use efficiency in response to nutrient supply, which may allow them to persist in soils of markedly different fertility. Northern red oak, for example, occurs on soils with a wide range of N availability in the northern Lake States; N-­use efficiency in this species drops as soil N availability increases; similar responses have been observed in sugar maple, basswood, and American beech (Zak et al. 1986). Although fine-­root mortality represents a significant input of litter to forest floor and mineral soil, we do not yet understand the extent to which soil nutrient availability influences the nutrient-­use efficiency of fine roots. While it is relatively easy to collect leaves prior to and following abscission, fine roots present a unique challenge to gain this same insight—­they remain buried in soil and are difficult to observe and analyze for nutrient content. The combined influence of plant litter production and nutrient-­use efficiency on the amount of N returned to forest floor and mineral soil is summarized in Table 19.4. Notice that N inputs from fine-­root litter are equivalent to or greater than the amount of N contained in leaf litter, regardless of whether forests occur in tropical, temperate, and boreal regions (i.e., root:leaf N ratio >1). The combined amounts of N entering forest floor and mineral soil are generally greatest in tropical forests, in which leaf decay is rapid. Recall rates of N input from atmospheric deposition (see Figure 19.3) and free-­living fixation (approximately 1–5 kg N ha−1 y−1) and realize that leaves and fine roots annually contribute 10 times more N to forest floor and mineral soil. Clearly, aboveground and belowground litter production and its nutrient-­use efficiency ­represent an important process influencing the internal cycling of nutrients within terrestrial ecosystems.

NUTRIENTS IN THE FOREST FLOOR Large amounts of organic matter and nutrients can accumulate on the soil surface in boreal and cool temperate forests, emphasizing the importance of the forest floor in controlling the flow of nutrients within these ecosystems (Table  19.5). Nevertheless, forest floor organic matter and nutrient content do not vary in a predictable manner among tropical, subtropical, and warm-­ temperate forests. In fact, the quantity of organic matter and nutrients in the forest floor of dry and montane tropical forests differ greatly from wet tropical forests and are more similar to that in temperate forests (Anderson and Swift 1983). Nevertheless, several important generalizations can be drawn from the information in Table 19.5. First, the accumulation of organic matter and nutrients in the forest floor of cold-­temperate and boreal evergreen forests is much greater than that in warm-­temperate, subtropical, or tropical forests. In temperate and tropical regions, deciduous forest floors contain half the organic matter of evergreen forest floors, regardless of whether they are dominated by needleleaf or broadleaf evergreen species. Although the greatest forest floor N and P contents occur in boreal evergreen forests, there is substantial variation among the N and P contents of tropical, subtropical, and temperate forest floors. The quantity of organic matter and nutrients contained in the forest floor of tropical, temperate, and boreal forests reflect two opposing processes: the production and the decomposition of plant litter. Consequently, it is not surprising that patterns of leaf and root litter ­production alone do not well reflect differences in forest floor mass among different forest ­ecosystems (compare Tables 19.4 and 19.5). Temperature, precipitation, and the biochemical constituents of plant

510  C h a p t e r 1 9   Nutrient Cycling

Table 19.4  The nitrogen content of leaf and fine-­root litter in forest ecosystems from different geographic locations. Litter Ecosystem

Leaf

Fine root

Root:leaf N

119

255

1.9

36

44

1.2

White pine

21

40

2.0

Mixed pine

16

36

2.2

white spruce

28

22

0.8

red pine

12

19

1.6

black oak

31

79

2.5

Northern red oak

30

62

2.0

white oak

26

47

1.8

sugar maple

23

47

2.0

paper birch

25

43

1.7

24

26

1.1

kg N ha−1y−1

Tropical Broadleaf evergreen Warm temperate Broadleaf deciduous Cold temperate Needleleaf evergreen

Broadleaf deciduous

Boreal Needleleaf evergreen

Compare the magnitude of nitrogen entering the soil from leaf and fine-­root litter and notice that the N content of fine-­root litter is equivalent to or greater than that of leaf litter. Source: Vogt et  al. (1986) / with permission of Elsevier and Nadelhoffer et  al.  1985. Modified from Advances in Ecological Research, © 1986 by Academic Press. Inc. Reprinted with permission of Academic Press, Inc. and from Ecology by the permission of the Ecological Society of America.

litter all influence the rate at which leaf and root litter is decomposed by soil microorganisms; differences in these factors contribute to the variation in forest floor accumulation among ecosystems in Table 19.5. Earthworm activity is also of importance, because these organisms incorporate fresh leaf litter into surface mineral soil horizons which accelerates the decomposition process (Edwards and Bohlen 1996). In the absence of earthworms, relatively thick, distinct organic horizons develop over thin A horizons. By contrast, discontinuous and thin organic horizons form over thick, organic-­matter-­rich A horizons in the presence of earthworms. A simple model considering plant litter production and decomposition has been used to gain insight into forest floor dynamics (Olson  1963). This mass-­balance approach assumes that the annual rate of litter decomposition equals the annual rate of plant litter production, such that the amount of organic matter in the forest floor is unchanged. This condition only occurs very late in forest ecosystem development, and it is where this approach has been used to estimate the rate of decomposition. Under this assumption, a constant proportion (k) of forest floor organic matter decomposes on an annual basis:



Litter Production

k Forest Floor Mass .

Nutrient Cycling within Forest Ecosystems  511

Table 19.5  Forest floor mass and nutrient contents in tropical, temperate, and boreal forest ecosystems. These values have been summarized for  a wide range of  forest ecosystems occurring in  tropical, temperate, and boreal regions of the Earth. Forest floor Organic matter (Mg ha−1)

N

Broadleaf evergreen

22.6

325

8

Broadleaf semideciduous

2.2

35

–

Broadleaf deciduous

8.8

–

14

Broadleaf evergreen

22.1

121

5

Broadleaf deciduous

8.1

–

–

Ecosystem

P kg ha

−1

Tropical

Subtropical

Warm temperate Broadleaf evergreen

19.1

60

4

Needleleaf evergreen

20.0

362

25

Broadleaf deciduous

11.5

163

12

Needleleaf evergreen

44.6

200

10

Broadleaf deciduous

32.2

624

50

44.6

875

81

Cold temperate

Boreal Needleleaf evergreen

Source: Vogt et al. (1986) / with permission of Elsevier. Modified from Advances in Ecological Research, © 1986 by Academic Press, Inc. Reprinted with permission of Academic Press. Inc.

In this equation, k represents the decomposition rate constant for the entire forest floor. With knowledge of litter production and forest floor mass, values of k are derived using the following equation:



k y

1

Litter production Mg ha 1y Forest floor mass Mg ha

1

1

.



When decomposition is rapid, relatively small amounts of organic matter accumulate in the forest floor, and values for k are typically greater than 1 y−1. This situation occurs in late-­successional tropical forests in which microbial activity has the potential to respire more organic matter than that contained in aboveground and belowground litter production (Cuevas and Medina 1988). By contrast, decomposition is slow in cold, boreal forests where substantial amounts of organic matter and nutrients accumulate in the forest floor; values of k are approximately 0.01 y−1. If the forest floor is neither gaining nor losing organic matter (i.e., it is in equilibrium), then k can be used to estimate the mean residence of time of organic matter or nutrients (l/k), which is the average time organic matter or nutrients reside in the forest floor. Values range widely among the forest ecosystems listed in Table 19.6, primarily reflecting differences in temperature, precipitation, and plant litter biochemistry. The mean residence time for organic matter, N, and P in the forest floor of tropical forests is less than 1 year (Table 19.6), whereas these constituents can reside within the forest floor of boreal forests for tens to hundreds of years.

512  C h a p t e r 1 9   Nutrient Cycling

Table 19.6  The mean residence time of organic matter, nitrogen and phosphorus, in the forest floor of tropical, temperate, and boreal forest ecosystems. Mean residence time of forest floor organic matter is calculated by dividing the forest floor mass (kg ha−1) by the annual leaf litterfall mass (kg ha−1 y−1); residence times for nutrients are calculated in the same manner.

Ecosystem

Organic matter

Mean residence time in yearsa N

P

Tropical Broadleaf evergreen

2

2

1

Broadleaf semideciduous

0.4

0.2

–

Broadleaf deciduous

0.9

–

2

Broadleaf evergreen

6

3

2

Broadleaf deciduous

2

–

–

Broadleaf evergreen

3

1

2

Needleleaf evergreen

5

14

11

Broadleaf deciduous

3

5

4

Needleleaf evergreen

18

33

22

Broadleaf deciduous

10

19

11

60

138

225

Subtropical

Warm temperate

Cold temperate

Boreal Needleleaf evergreen

Mean residence times are 1/k. Source: Vogt et al. (1986) / with permission of Elsevier. Modified from Advances in Ecological Research. © 1986 by Academic Press, Inc. Reprinted with permission of Academic Press, Inc.

a 

Relatively large amounts of N and P (Table 19.6) in the forest floor of cool temperate and boreal forests, combined with long residence times (1/k), emphasize the importance of the forest floor as a site for nutrient storage in these forest ecosystems. Also, compare the mean residence times of coniferous (i.e., needleleaf) and deciduous forests, and notice that organic matter and nutrients reside in the forest floor of coniferous forests for relatively longer periods of time. This pattern reflects the fact that conifers contain greater proportions of organic compounds that are not easily metabolized by soil microorganisms, thus they have slower rates of litter decomposition. Mean residence times in Table 19.6 are calculated using rates of aboveground litter production alone, which comprise only 28–56% of total litter input to forest floor (see Table 19.4). Using leaf litter production as the primary input of plant litter, the mean residence time of organic matter in the forest floor of cool temperate forests ranges from 8 to 67 years. The range of values is much lower when fine-­root litter is included in the calculation of mean residence time (e.g., 5–15 years; Vogt et al. 1986), further emphasizing the importance of belowground litter production in studying the storage and cycling of nutrients in forest ecosystems.

ORGANIC MATTER DECOMPOSITION AND NUTRIENT MINERALIZATION To fully understand the process of decomposition and the release of nutrients from plant litter and organic matter, one needs to consider the factors that control microbial growth and maintenance in soil. Once abscised leaves, dead roots, and other plant tissues enter forest floor and mineral soil, they

Nutrient Cycling within Forest Ecosystems  513

become a growth substrate for saprotrophic microorganisms. During the decomposition process, soil bacteria, actinobacteria, and fungi assimilate the organic compounds contained in plant litter into their cells for biosynthesis (i.e., growth and maintenance), albeit at different rates depending on the types of compounds contained in plant litter. Whether soil microorganisms release nutrients into, or assimilate nutrients from, soil solution depends on the biochemical constituents of plant litter and their suitability for microbial growth and maintenance. In the paragraphs that follow, we discuss how plant litter functions as a substrate for microbial growth and maintenance in soil, and, in turn, how these microbial processes influence the supply of growth-­limiting nutrients to plants. As you will read later in this section, the amount of energy that soil microorganisms derive from plant litter controls the processes of decomposition and soil N availability.

Biochemical Constituents of Plant Litter  The leaves, stems, and roots of plants are con-

structed of a remarkably small set of organic compounds, regardless of differences in phylogeny and growth form. Plant litter is largely composed of cellulose (15–60%), hemicellulose (10–30%), lignin (5–30%), protein (2–15%), fats (1%), and soluble compounds such as sugars, amino acids, nucleic acids, and organic acids (10%). These organic compounds carry out essential metabolic functions, provide physical support, and defend plants against herbivores and pathogens. Upon entering soil, however, they become substrates that yield different amounts of energy for microbial growth and maintenance. The process of decomposition is mediated by the microbial production of enzymes that harvest the biochemical energy contained in compounds formed during the biosynthesis of plant tissue. In most terrestrial ecosystems, microbial growth in soil is limited by usable forms of energy, and amounts contained in annual litter production are only sufficient to maintain (i.e., no growth) microbial populations in soil. The microbial production of enzymes that make the energy contained in plant compounds available is energetically expensive. Consequently, energy produced during the microbial metabolism of a particular plant compound must surpass the cost of the enzymes used to harvest its energy in order for microorganisms to use it as a growth substrate. The use of a plant-­derived organic compound as a substrate for microbial growth is determined by: (i) the amount of energy released by breaking different types of chemical bonds, (ii) the size and three-­dimensional complexity of molecules, and (iii) its nutrient content (N and P). Look at Figure 19.12 and notice that the complexity of chemical bonds, three-­dimensional structure, and nutrient content varies dramatically among the primary constituents of plant litter. Simple sugars (i.e., carbohydrates), such as glucose, are among some of the first products of photosynthesis, and although their concentration in fresh litter is low ( 128

F I G U R E 2 2 . 2  Housing density in 1940 and 2000 directing landscape change in the midwestern United States. Housing growth occurred both as suburban sprawl along the fringes of metropolitan areas and rural sprawl in previously unaltered forested areas. Source: Modified from Radeloff et al. (2005).

Forest Fragmentation and Connectivity  615

FOREST FRAGMENTATION AND CONNECTIVITY

Among the first concerns of landscape ecology was the idea of landscapes changing as a result of human activities. The most concerning of these changes was the idea that human land use such as forest clearing for agriculture, colonization, or forest harvesting was acting to break up large, contiguous tracts of forests with deleterious ecological effects (Harris 1984). Given its occurrence at broad spatial scales and its emphasis on spatial heterogeneity, forest fragmentation was a natural focus for landscape ecologists. The classic example of forest fragmentation was shown in a township in southern Wisconsin where land clearing and farming created an increasingly fragmented forest from 1882 (when only 30% of the original forest remained) to 1950 (Figure 22.3). By 1954,

1831

1882

1902

1950

F I G U R E 2 2 . 3  Fragmentation of a forested area of Cadiz Township, Green County, Wisconsin, during the period of European colonization. The township is 6 miles on a side. The shaded areas represent land remaining in, or reverting to, forest in 1882, 1902, and 1950. Source: Curtis (1956) / with permission University of Chicago Press.

616  C h a p t e r 2 2   Forest Landscape Ecology

Curtis (1956) found only 3.6% of the original forest remained, and 77% of what remained was so heavily grazed that tree regeneration was absent. The forest fragments are therefore functionally isolated from other and larger areas of forest. Much of our understanding of forest fragmentation and its ecological effects is based in patch-­based ecological theory, as described in the following text.

PATCHES IN FOREST ECOLOGY Many aspects of landscape ecology today are based on ecologists’ conceptualization of landscapes as a mosaic of patches of habitats, ecosystems, or cover types. The patch–corridor–matrix model, as described by Forman (1995), suggests that the extent and configuration of patches on a landscape is what defines landscape pattern. As described in the earlier text, patches are to some extent human constructs and are defined based on the phenomenon under consideration, such as forest ages, forest types, or vegetation type (e.g., forest vs. field). Patches are considered to be distinct areas at a given point in time, with boundaries different enough from what surrounds them that they can be identified as a relatively homogenous area at a given spatial scale. Corridors are narrow, linear patches of a patch type that differs from those on either side (Forman 1995) and typically connect similar patch types on a landscape. Corridors may function as breeding habitat for organisms that enables gene flow across a landscape, may allow dispersal or migration across a landscape, or may act as a barrier or a filter to the movement of organisms. The most common or connected landscape element is known as the matrix, which is often considered to be an unusable space to the organisms or phenomenon of interest. Thus, the matrix in Figure 22.3 is non-­forested area beginning in 1882. Patch-­based ecological theory is heavily influenced by the theories of island biogeography (MacArthur and Wilson  1967) and metapopulations (Levins  1969,  1970). Island biogeography theory was originally developed to explain community dynamics on oceanic islands and suggests that the number of species on island communities is affected by the size of the island and its distance from the mainland from which organisms disperse. Larger islands support more species because they have more complex habitat and large populations less threatened by extinction. In addition, islands further from the mainland support fewer species because they are colonized less frequently. Therefore, large islands closest to the mainland are likely to have the most species, while small islands far from the mainland are likely to have the fewest species. Island biogeography theory is rather easy to transpose onto fragmented landscapes, allowing assumptions about biodiversity in small, isolated patches of forest compared to larger, well-­connected fragments. Metapopulation theory adds an additional layer of complexity to patch-­based theory. Metapopulations include groups of populations that are spatially separated, as they would be when occupying fragments of habitat on an altered landscape. Though separated in space, the populations interact through dispersal among the patches, forming a “population of populations” (Levins 1969). Metapopulation theory predicts that a specific population in the metapopulation may go extinct, leaving some patches on the landscape suitable but unoccupied, but the patch will eventually be re-­colonized by immigration from an occupied patch. Thus, the metapopulation as a whole remains stable even if some of the populations within it do not. Patch-­based theory is one of the foundational aspects of forest landscape ecology and still permeates the field today. In general, most forest ecologists have realized that the natural world is not simply a series of patches surrounded by a featureless, unhabitable matrix. In such a scenario, every species—­including trees and other plants—­would “perceive” a very different landscape, making management or conservation nearly impossible. Instead, the matrix itself is made up of a mosaic of different patch types that may be suitable for dispersal or other activities of life history but not for breeding (Figure 22.4). As such, the matrix surrounding patches may not be completely analogous to an inhospitable ocean surrounding islands. Forest fragmentation such as that shown in Figure 22.3 may indeed create functionally isolated forests and woodlots for trees because the

Forest Fragmentation and Connectivity  617

F I G U R E 2 2 . 4  Distribution of eastern white pine on a portion of Harvard Forest, Massachusetts. White pine forest is highly fragmented when only the white pine patch type and the matrix are considered (left), but in reality the matrix forms a complex mosaic (right) of patches of other conifers (pink), mixed conifer stands (lighter greens), open areas (yellow), hardwood stands (orange), mixed hardwood and conifer (brown), and wetlands (blue).

nature of the matrix may prevent dispersal, regeneration, and/or establishment. We cannot assume, however, that some organisms, particularly animals, may not be able to traverse the matrix even as the patches within it become smaller and more isolated. This landscape-­mosaic view is probably a more accurate representation of the natural world than that of the patch–corridor– matrix view, and its use for understanding and characterizing landscapes is at the forefront of forest landscape ecology.

FOREST FRAGMENTATION As described in the earlier text, human activities have reduced total forest area and have fragmented remaining forests into progressively smaller patches isolated by adjacent plantations, roads, or agricultural and urban development (Harris 1984; Saunders et al. 1991). Forested land in the 48 contiguous United States is estimated to have occupied 400  million ha in the 1500s (­Harrington 1991), and it was reduced by about 53% to approximately 188 million ha by the 1920s. By the early 1990s only 3–5% of pristine old-­growth forest remained, principally in the Pacific Northwest (Miller 1992). Forest fragmentation is widespread in the eastern United States, where remaining forests date only from the 1920s and 1930s (Rudis 1995; Vogelmann 1995) and small, isolated woodlots interspersed in farmlands and suburbia comprise about 40% of the deciduous forest (Terborgh 1989). We discuss in the following text the ecological effects of fragmentation (and associated forest loss).

Ecological Effects  Forests have always been naturally fragmented by large disturbances even

prior to European colonization (Chapter 16). Disturbances such as fires, windstorms, floods, and avalanches characteristic of regional and local ecosystems created openings of various sizes on the

618  C h a p t e r 2 2   Forest Landscape Ecology

landscape, providing the heterogeneity necessary for a diverse array of animals and plants. Pattern on modern landscapes differs from that prior to European colonization, however, at least in part because changes in land use, conversion of cover types, or other drivers of current fragmentation often prevent or reduce regeneration in disturbed patches. An important consideration regarding forest fragmentation is that many of its effects are actually due to the loss of forest cover rather than simply its subdivision into relatively small, isolated patches (Fahrig  2003). Most landscape ecologists agree that fragmentation affects forests on a landscape by (i) reducing the amount of forest, (ii) increasing the number of forest patches, (iii) reducing the size of the forest patches, and (iv) increasing the isolation of the forest patches. However, these aspects are not independent of each other. For example, removal of some patches via forest loss would reduce the number of patches but increase the isolation of those that remain ­(Figure 22.5a), but subdivision of patches would increase the number of patches and reduce their size while decreasing their isolation (Figure 22.5b). Thus, forest fragmentation and loss are closely related to each other but are not equivalent. An increase in the number of patches on a landscape is clearly caused by fragmentation rather than loss (With 2019), but both patch size and isolation may be at least as well explained by loss as by fragmentation (Bender et al. 2003; Fahrig 2003). The important assertion, therefore, is that forest fragmentation changes the function of the patches of forest that remain as well as decreasing their size and increasing their isolation (van den Berg et al. 2001). Fragmentation effects may be classified into those resulting from patch size and isolation, and those resulting from edge effects. Predictions about the effects of decreasing patch size on the diversity and abundance of species (most often animals) lean heavily on island biogeography ­theory: larger patches are expected to favor larger populations compared to smaller patches, and thus (a)

Loss of patches: # patches reduced, patch isolation increased, patch size unchanged

(b)

Subdivision of patches: # patches increased, patch isolation decreased, patch size decreased

F I G U R E 2 2 . 5  Potential effects of forest loss on landscape pattern. Fragmentation is expected to increase the number of small, isolated patches as it proceeds on a landscape, but these effects are not always consistent and some may be achieved by forest loss alone. (a) Forest loss may increase patch isolation yet reduce the number of patches, and (b) subdivision may increase the number of patches and patch size yet decrease patch isolation.

Forest Fragmentation and Connectivity  619

extinction rates are expected to be lower in larger patches. Most species have a minimum patch size requirement (Diaz et al. 2000), and thus fragmentation of forests may result in patches too small to provide adequate heterogeneity for territory size, food supply (Whitcomb et  al.  1981), or other required features such as streams or wetlands. As described in the earlier text, patch isolation reflects the amount of forest in the landscape more so than its fragmentation, but isolated patches are less likely to be recolonized if a population within them goes extinct, depending on how well dispersers are able to move through the matrix among the patches (Ricketts 2001). For example, small mammals are more likely to move among forest patches through open habitat if the patches are close together and exposure to predation in the open areas will be limited. The most unambiguous indicator of fragmentation is the amount of edge that is increased as the process takes place. The amount of edge increases proportionately as patch size decreases, often represented as a ratio of the length of the patch perimeter to the patch area. Edge effects are the ecological consequences of sharp boundaries between two patch types, and they tend to be stronger than effects of patch area (With 2019). In forests, edge effects are often represented by adjacent forests and disturbed areas or openings such as those created by harvesting; edges thus ring the perimeter of each forest patch. The forest edge is characterized by relatively sharp boundaries with microclimatic, vegetational, and biotic conditions markedly different than in the interior of the forest. In a series of studies of clear-­cut harvesting in old-­growth Douglas-­fir forests of the Pacific Northwest, Chen et al. (1992, 1995) found warmer air and soil temperatures during the day and cooler temperatures at night at the edge compared to the interior forest, as well as lower relative humidity, higher short-­wave radiation, and exponentially higher wind speed. These effects extended 30 to over 240 m into the forest from the edge. Although not tested explicitly for an edge effect, canopy cover, basal area, and trees per unit area decreased in the forest nearest the edge, dominant tree growth increased, tree mortality increased, and tree regeneration was impacted in a species-­specific way. The abrupt, human-­created sharp boundaries along forest edges differ from those to which many forest species have adapted. Edges favor species adapted to them and select against those requiring interior conditions (Whitcomb et  al.  1981; Yahner  1988,  1995; Saunders et  al.  1991), potentially changing species richness. Such edge effects are evident with many forest animals but are particularly obvious with birds. Edge species of birds that forage and breed near forest edges are relatively independent of forest patch size and may be unaffected or even positively affected by fragmentation. By contrast, forest interior species—­those that depend on large forest tracts and avoid edges and their effects—­may decline in numbers with fragmentation. For example, Robbins et al. (1989) found that encountering a breeding pair of an area-­dependent bird species such as the scarlet tanager is much more likely (>70% probability) in a 100-­ha forest compared to a tract smaller than 10 ha (1000 ha) were generally absent until 1988; only three fires >4 ha burned from 1795 to 1978, all of which were 250 years). The even-­aged pine canopy that developed after fire deteriorates as the stands age and is replaced by trees able to survive in the understory and seed into gaps. Most of the subalpine plateaus in Yellowstone are dry and dominated by lodgepole pine, but subalpine fir and Engelmann spruce are common and may ­co-­dominate with pines on occasional moist sites, particularly in older forests. This relatively homogeneous site-­vegetation pattern contrasts with the diverse mosaic of physiography, human-­use history, and disturbance regime of central New England. As a result, the forest landscape mosaic in Yellowstone is largely based on forest age and structure rather than species composition. Large fires in 1739, 1755, and 1795 reduced the high proportion of old-­growth forest present in the early 1700s and replaced it with early-­successional forests (Figure 22.11). As the forests burned by these fires aged, middle-­successional stages became most abundant around 1800 and gradually transitioned to late-­successional forest. After creating a map of the mosaic resulting from multiple fires over 200 years, Romme (1982) applied species diversity indices (Chapter 14) to the mosaic to characterize landscape diversity from 1778 to 1978. Changes in landscape diversity were quantified using (i) a weighted average of richness (number of stand ages), evenness (proportion of the landscape occupied by each stand age), and patchiness; and (ii) the Shannon–Wiener index. Both measures show that the large fires of 1739, 1755, and 1795 created high diversity in the late 1700s and early 1800s, and a lack of large fires in the late 1800s during a 70-­year period with no major fires resulted in low diversity ­(Figure 22.12). Two small fires and a series of mountain pine beetle outbreaks in the twentieth century led to differential rates of succession that increased landscape diversity (Romme 1982). These differences in landscape diversity over time in turn had important effects on bird and ungulate populations, which are described in detail by Romme and Knight (1982). To this point, we have discussed heterogeneity in Yellowstone as having resulted from a series of stand-­replacing fires over several centuries. As described in the earlier text, single disturbance events also may create complex landscape patterns, particularly if they affect broad spatial extents (Foster et al. 1998), as illustrated by the extensive Yellowstone fires of 1988. As described in the earlier text, very large, stand-­replacing fires were a dominant factor on the Yellowstone landscape prior to European colonization, with such fires occurring at intervals of 100–300 years (Romme and Despain  1989). Due to prolonged, extreme drought and high winds that occurred

Middle

% Watershed Area

60 50 40 30

Late

20 10

40 20

Early % Watershed Area Burned 1739 1795 1755 1738 ‘58

1834 ‘78

‘98 1818 ‘38

1905 ‘58 ‘78

1949

‘98 1918 ‘38

‘58

‘78

Year F I G U R E 2 2 .1 1   Percentage of the 73 km2 Little Firehole River watershed occupied by early, middle, and late stages of forest succession from 1738 to 1878 in Yellowstone National Park, northwestern Wyoming. Source: Romme and Knight (1982) / with permission of Oxford University Press.

100

(a) Average of Richness, Evenness, and patchiness Indices

90

80

(b) .90

Shannon Index

.80

.70

1778 ‘98 1818 ‘38 ‘58 ‘78 ‘98 1918 ‘38 ‘58 ‘78 Year

Time of effective fire control

F I G U R E 2 2 .1 2   Changes in landscape diversity in the Little Firehole River watershed from 1778 to 1987 in Yellowstone National Park. (a) Average of richness, evenness, and patchiness indices; Y axis = percentage of maximum possible average of richness, evenness, and patchiness indices. (b) Shannon–Wiener diversity index; Y axis = percentage of maximum possible Shannon–Wiener diversity index. Maximum possible Shannon–Wiener index would occur where all stand ages have equal coverage in the landscape. Source: Romme and Knight (1982) / with permission of Oxford University Press.

Disturbances on Landscapes  629

during the summer of 1988, fires burned approximately 250 000 ha of lodgepole pine forest. In contrast to expectations at the time, what resulted was not a homogenous landscape, but a markedly heterogeneous one, including a mosaic of burned and unburned patches with varying burn severities (Turner et al. 1994). Sooner than 5 years after the 1988 Yellowstone fires, lodgepole pine regeneration varied substantially across the landscape, from areas with no seedlings to areas with >500 000 seedlings per hectare (Turner et al. 2004). Seedling densities truly varied in a fine-­grained mosaic, with mean patch size 1.5 ha, 68 patches occurring per 100 ha, and similar patches separated on average by only 150 m (Kashian et al. 2004). Overall, small, dense patches of seedlings occurred within a matrix of large, sparser patches across the landscape (Figure 22.13). Variation in burn severity was important in determining post-­fire seedling density; areas burned by severe surface fires had higher seedling densities than areas burned by crown fire or light surface fires (Turner et al. 1994). Lodgepole pine has serotinous cones (Chapter 10) that require heat to open, but may be incinerated by highintensity fires. While crown fires may have incinerated many cones and light surface fires may have been insufficient to open them, severe surface fires likely provided the appropriate burn temperatures to maximize post-­fire regeneration (Turner et al. 1994). The expression of the serotinous trait by lodgepole pine also varies across the landscape and is not well understood, but higher proportions of trees producing serotinous cones in a stand are correlated with shorter fire intervals and moderately aged stands (Tinker et al. 1994; Schoennagel et al. 2003). The heterogeneity created by the 1988 fires in Yellowstone has many implications for the future landscape. Although succession and stand dynamics will proceed as the post-­fire forests age,

N

Legend Sapling density (stems/ha) 0–1000 1001–5000 5001–10 000 10 001–20 000 20 001–50 000 > 50 000 Lakes Roads 0

5

10

Kilometres 20

F I G U R E 22 .1 3   Heterogeneity of lodgepole pine regeneration density across the area burned in Yellowstone National Park in 1988. Regeneration density varied over six orders of magnitude in a fine-­grained, complex spatial mosaic. Source: Kashian et al. (2004) / with permission of Canadian Science Publishing.

630  C h a p t e r 2 2   Forest Landscape Ecology

the landscape pattern created in 1988 is likely to leave a long-­lasting legacy of the fires for decades to centuries. Studies of stand structural dynamics on the unburned portion of the Yellowstone landscape found high variability in stand density that decreased over time (Kashian et al. 2005b). Most stands on the landscape exhibited classic dynamics of a fire-­regenerated, shade-­intolerant species, showing evidence that they had initially been relatively dense and their density had decreased over time due to self-­thinning (Chapter 17). However, several stands 50–150 years old had clearly regenerated sparsely after the fire that last burned them, allowing gradual and continuous recruitment to occur and density to remain stable or increase over time. Younger stands were not only more dense, but their density varied more across the landscape; stand density was generally similar among stands 200 years and older (Kashian et al. 2005b). The initial patterns of regeneration density produced by the 1988 fires largely remained in 2012 such that convergence had not yet occurred, but the mechanisms that would likely cause convergence across the landscape were observed (Turner et al. 2016). Notably, smaller fires ( 55 EVAPOTRANSPIRATION (cm/yr)

9–11 12 13 14 15 16 17 18 > 19 PHOTOSYNTHESIS (Mg · ha–1· yr–1) F I G U R E 2 2 .1 6  Heterogeneity of evapotranspiration and net photosynthesis as determined by leaf area index and an ecosystem model for a 1540 km2 landscape in northwestern Montana. Heterogeneity of the ecosystem process closely shows physiographic differences of the forested landscape. Source: Running et al. (1989) / John Wiley & Sons.

Interactions of Landscape Patterns and Ecological Processes  635

FOREST CARBON DYNAMICS Incorporating heterogeneity is particularly important for creating carbon budgets and for understanding the carbon balance on landscapes because of the variability of forest species composition, age, and other factors. A carbon balance may be determined from leaf to global scales, but net ecosystem productivity (NEP; also termed net ecosystem carbon balance, or NECB) in a forest essentially represents the balance between carbon fixed into live biomass via photosynthesis and that lost from dead biomass as it decomposes (Chapter 18). The importance of NEP for large landscapes relates to whether a region acts as a carbon source (losing more carbon to the atmosphere than it gains) or a carbon sink (gaining more carbon than it is loses to the atmosphere). Because forests are a major pool of carbon on the earth’s surface, and they act to remove carbon from the atmosphere, disturbances that impact forests are especially important in determining whether a landscape is a source or sink of carbon and therefore how it contributes to climate change (Chapter 20). As discussed in Chapter 18, forests are typically carbon sources immediately after a disturbance when small, live trees with relatively low ANPP do not compensate for carbon lost from high levels of dead, decomposing biomass killed by the disturbance. Depending on the forest type, NEP reaches its maximum several years or decades after the disturbance when tree growth is most vigorous (primary productivity is highest) and most dead biomass has decayed away, and forests then act as strong carbon sinks. By contrast, older forests have accumulated large stocks of carbon even as primary productivity declines and dead biomass begins to re-­accumulate, and NEP approaches zero. Many studies employ a chronosequence approach to determine how carbon storage changes after a disturbance, whereby a series of stands are sampled that vary in age, and all differences among the stands are assumed to occur due to the effects of succession (Chapter 17). For example, Rothstein et  al. (2004) sampled 11  jack pine stands ranging in age between 1 and 72 years in northern Michigan. They found that jack pine-­dominated forests are initially a carbon source for 6 years after wildfires because decomposition exceeds primary productivity, then shift to a strong carbon sink until stands matured and primary productivity declined. The patterns described in the earlier text are very common for systems where disturbances are stand-­replacing and may also vary with stand structure (Figure  22.17; Kashian et  al.  2006). Stand density is important because it affects the rate of carbon uptake and storage by vegetation. Because they have higher leaf area and photosynthate accumulates more quickly, dense stands should accumulate more carbon quicker to form a strong sink after the disturbance. Sparse stands should take longer to shift to positive NEP and achieve a weaker sink, but the duration of the sink is likely to be prolonged compared to dense stands (Kashian et al. 2006). Notably, Kashian et al. (2013) found that stand density was important for carbon storage in forests in Yellowstone National Park that had not yet reached canopy closure, but that tree size and age were most important after this point. Canopy closure represents the point at which stands reach their maximum leaf area, and thus changes in leaf area may be a critical driver of NEP after disturbances. Given the heterogeneity of both stand age and structure across landscapes, estimates of NEP at broad spatial scales are extremely useful because they account for the variability that directly affects carbon storage. Landscapes dominated by young or sparse forests may be carbon sources while those dominated by older or dense forests may be carbon sinks, or any combination therein. In addition to the influence of physiography as described by Running et al. (1989), species composition also imposes spatial heterogeneity in carbon dynamics across landscapes. For example, boreal forests with mixed conifer and hardwood composition were found to store 47% more carbon in the aboveground vegetation compared to black spruce forests and 44% more than jack pine forests (Martin et  al.  2005). In part, mixed stands were able to accumulate carbon more quickly because they contain species with contrasting shade tolerances and thus are able to form multilayered canopies, supporting a greater leaf area. Landscape carbon studies able to incorporate all factors that are spatially heterogeneous are very difficult and remain elusive, but clearly accounting for such heterogeneity is critical in understanding carbon storage at broad spatial scales.

636  C h a p t e r 2 2   Forest Landscape Ecology 150

ΔC Vegetation (g C m-2 yr-1)

Dense 100 Sparse

50

Total NEP (g C m-2 yr-1)

0 100 Dense

50

Sparse

0 –50 –100 0

50

100 150 Age since fire

200

250

F I G U R E 2 2 .1 7   Carbon balance following disturbances for lodgepole pine forests of contrasting structure in the central Rocky Mountains. Carbon accumulates quickly and to a higher maximum for a short duration in dense stands compared to sparse stands, which are slower to accumulate carbon. These differences in vegetation drive similar changes in net ecosystem productivity (NEP). Source: Kashian et al. (2006).

Incorporating the heterogeneity present on landscapes into our understanding of carbon dynamics has implications for predicting how forested regions may contribute to atmospheric carbon and climate change. Any disturbance to a forest will release carbon to the atmosphere if it kills trees and will therefore temporarily shift the landscape to a carbon source. The amount of carbon released, and the strength of the source, will vary with the type and severity of the disturbance and the heterogeneity of these factors that may occur across a landscape. For example, stand-­replacing wildfires will create a strong and immediate carbon source due to combustion and widespread tree mortality and subsequent decomposition, but less-­severe surface fires may create only a weak source if relatively few trees are killed. Likewise, moderate insect outbreaks or windstorms will create a weaker, more prolonged carbon source because a higher proportion of trees survive the disturbance, and variation in their severity may affect the strength or duration of the source.

NUTRIENT DYNAMICS Although it is clear that landscape heterogeneity in nutrient pools and fluxes is the rule rather than the exception for most systems, fewer researchers have been successful in both quantifying and explaining such patterns (Turner and Gardner 2015). Much of this work in temperate forests has focused on the nitrogen cycle, given that nitrogen is often the limiting nutrient for forest productivity (Chapter  19). A series of studies by Zak et  al. (1986,  1989) and Zak and Pregitzer (1990) related the spatial variation in nitrogen-­related processes to the patterning of landscape ecosystems in northern Lower Michigan. In comparing nine ecosystems located on various landforms and soil textures, net mineralization was lowest in a dry oak ecosystem and highest in a mesic northern hardwood ecosystem (Table 22.1). Nitrification was also highest in the northern hardwood ecosystem and far higher than in any other ecosystem sampled. Overstory biomass was asymptotically related to differences in mineralization, but was unrelated to nitrification rates. Thus, the spatial heterogeneity in at least some nutrient dynamics closely tracks the spatial distribution of ecosystems across the landscape (Zak and Pregitzer 1990).

Interactions of Landscape Patterns and Ecological Processes  637

Table  22.1  Association of  mineralization and  nitrification rates with  landscape ecosystems in northern Michigan. Landform

Percent coarse and medium sand

Mineralization (μg N/g)

Nitrification (μg N/g)

Black and white oak

Outwash plain

72

 52.0

 0.9

Black and white oak

Outwash plain

70

 49.6

 0.3

Black, white, and red oak

Kame and moraine

67

 47.2

 0.2

Red, white, and black oak

Kame and moraine

63

 60.2

 0.2

Red and white oak

Kame and moraine

55

 63.0

 0.6

Red oak and red maple

Kame and moraine

41

 81.8

 0.2

Sugar maple and red oak

Moraine

61

 75.5

 4.1

Sugar maple and red oak

Moraine

56

 93.4

 1.2

Sugar maple, basswood, and white ash

Moraine

55

127.8

45.5

Dominant overstory species

Source: Zak et al. (1989) / with permission of Canadian Science Publishing.

At broad spatial scales, spatial variability in some nitrogen transformations may be explained by the heterogeneity of soil texture. For example, Reich et al. (1997) found a strong influence of soils on mineralization for forests in Wisconsin and Minnesota, with higher rates of mineralization on finer-­textured soils. Likewise, Groffman and Tiedje (1989) and Groffman et  al. (1992) examined rates of denitrification, which tends to occur on wet sites in the absence of oxygen, in forest soils of southern Michigan. Rates of denitrification across the landscape varied with soil texture and associated drainage, and were differentially responsible for regional estimates of denitrification. Loamy soils represented 47% of the forest but 73% of the denitrification, and clay soils represented 9% of the forest but 22% of the denitrification. Meanwhile, sandy soils represented 44% of the forest but only 5% of the denitrification. Together, these studies are examples of ecosystem processes whose heterogeneity may be predicted by the heterogeneity of physical site factors across the landscape. Heterogeneity of nutrient dynamics at broad spatial scales appears to be most easily related to physical site factors, but a general framework that associates nutrient dynamics with more general characteristics of landscape pattern has been more fleeting. However, an experimental study of forest fragmentation in Kansas may have begun to uncover a relationship between patch size and soil nitrogen dynamics (Billings and Gaydess 2008). The study was designed with the assumption that edge effects in fragmented forests result in changes in dominant tree species that in turn may affect soil nutrient cycling. The experimental site contained small (4 × 8 m2) and large (50 × 100 m2) patches of successional forest that were established in 1984. Large patches had greater rates of woody plant colonization, higher plant species richness, and likely higher aboveground net primary productivity (ANPP) per unit area than smaller patches, all of which became evident after about 15 years of succession in the experimental fragments. These differences in vegetation apparently resulted in greater rates of mineralization and nitrification in soils of the small patches compared to the large patches, associated with greater fine root biomass and root nitrogen concentration. Soils from small patches also had higher potential denitrification rates compared to large patches, which is often associated with faster nitrogen cycling rates (Venterea et al. 2003). Although these patterns have yet to be rigorously tested in other landscapes, the potential that forest patch size—­or landscape configuration in general—­may govern the spatial heterogeneity in nutrient cycling is intriguing.

638  C h a p t e r 2 2   Forest Landscape Ecology

Much like the landscape ecosystem approach we reference throughout this book, landscape ecology focuses on the importance of scale, is concerned with the effects on a specific place on a landscape by the ecosystems that surround it, and emphasizes physical site factors to some degree. The landscape ecosystem approach involves mostly a focus on natural systems, with perhaps some acknowledgement that humans are a significant disturbance to those systems. Additively, landscape ecology explicitly includes humans as an important variable in understanding spatial heterogeneity, including how their perceptions, values, and decisions shape landscapes. We have not addressed this human dimension in our brief overview of forest landscape ecology, but we do so in detail in our treatment of sustainability in the chapter that follows.

SUGGESTED READINGS Fahrig, L. (2003). Effects of habitat fragmentation on biodiversity. Annu. Rev. Ecol. Evol. Syst. 34: 487–515. Forman, R.T.T. (1995). Land Mosaics, the Ecology of Landscapes and Regions. Cambridge University Press 632 pp. Foster, D.R. and Boose, E.R. (1992). Patterns of forest damage resulting from catastrophic wind in central New England, USA. J. Ecol. 80: 79–98. Landres, P.B., Morgan, P., and Swanson, F.J. (1999). Overview of the use of natural variability concepts in managing ecological systems. Ecol. Appl. 9: 1179–1188.

Romme, W.H. (1982). Fire and landscape diversity in subalpine forests of Yellowstone National Park. Ecol. Monogr. 52: 199–221. Turner, M.G. (1989). Landscape ecology: the effect of pattern on process. Annu. Rev. Ecol. Evol. Syst. 20: 171–197. Turner, M.G. and Gardner, R.H. (2015). Landscape Ecology in Theory and Practice, 2e. New  York: Springer 502 pp. With, K.A. (2019). Essentials of Landscape Ecology. Oxford, UK: Oxford University Press 656 pp.

Sustainability of Forest Ecosystems

CHAPTER 23

A

t many instances in this book, we have suggested that humans have had far-­­reaching ­influences on forests, both directly and indirectly. Their influences have been discussed throughout the first 22 chapters of this book, including impacts on the overriding effects of ­climate (Chapter 20); extensive alteration of disturbance regimes (Chapter 16); intentional and accidental introduction of non-­­native plants, animals, and pathogens that eventually became invasive (Chapter 21); widespread reductions in biological and ecological diversity (Chapter 14); and fragmentation of forest landscapes (Chapter 22), among many others. Forest ecology does not exist in a vacuum, in that every forest ecosystem on Earth has in some way been influenced by humans. Although we struggle to understand how the natural world works, we cannot completely study forest ecosystems independent of human influence, especially when we attempt to consider the future. This final chapter grapples with the concept of sustainability: its meaning, its role in ecology, and its applications to forest ecology.

CONCEPTS OF SUSTAINABILITY

Without question, humans are part of ecosystems and are not external to them. Although sometimes difficult to imagine because of their wide-­­ranging influence, humans are supported by the physical site factors of landscape ecosystems as much as any other organism. At the same time, however, humans have a disproportionate influence on the ecosphere and have profoundly changed regional and local ecosystems in both dramatic and subtle ways. Perlin (1989) provides a remarkably powerful historical record of human destruction of forests that accompanied the development of civilizations from the Bronze Age in Mesopotamia to nineteenth-­­century America. In several papers, Kay (1994) speculated that the role of indigenous people is probably underestimated in its importance of changing the biota in western North America. He concluded that the modern concept of wilderness as areas without human influence is a myth because indigenous people changed the composition of entire plant and animal communities by limiting ungulate numbers and purposefully modifying the vegetation with fire. Of course, the effects of European colonizers in North America have included forests cleared for agriculture, logging, livestock grazing, long-­­term changes in dominant tree species, forest fragmentation, climate change, introduction of invasive species, and changes in disturbance regimes. The piercing story of human interventions in the complexity of ecosystems is compellingly described for the Blue Mountains of Oregon by Langston (1995).

THE PREVALENCE OF HUMAN VALUES IN FOREST ECOLOGY The widespread decline of ecosystems has led to a great interest in the idea of sustainability since the 1970s (Du Pisani 2006). A general notion of sustainability is that current desirable conditions of an object or process—­be they species composition, structure, recreational opportunities, a ­perception

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

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of wildness, etc.—­continue to exist for the benefit of future generations. In most cases, ­sustainability considers the maintenance into the future of something currently present. For example, Chapin et  al. (1996) defined a sustainable ecosystem as one that “over the normal cycle of disturbance events maintains its characteristic diversity of major functional groups, productivity, soil fertility, and rates of biogeochemical cycling.” Similarly, Turner et al. (2013) defined sustainability as “use of the environment and resources to meet current needs without compromising the ability of a system to provide for future generations.” One definition is strictly ecological and the other incorporates social and economic structures, but both definitions are sufficiently broad and ambiguous to place human values on real units of nature. Because humans differ in the way they perceive the world, placing their values onto specific components of landscape ecosystems ­inevitably results in differences of opinion. In the first edition of this textbook, Stephen H. Spurr elegantly described how the ecosystems we see today are but a snapshot in time and space, and the values humans place on them may also be temporary. In scrutinizing the concepts of native and introduced species, he reasoned (Spurr 1964, p. 157): . . . there is no meaning to the concepts of native and introduced species from a biocentric viewpoint. Characterizing a plant or animal as being exotic or endemic characterizes it only from the standpoint of man’s relationship to it. Actually, all plants and all animals are introduced or exotic from a biocentric standpoint except at the very point in space where the particular gene combination was first put together. Whether the subsequent migratory pattern of that organism took place independent of man or with the help of man is important to man but not to the wilderness ecosystem itself. It is an interest of man to know whether he carried the coconut to a given tropical island or not. To the coconut it is of little importance as to whether it floated by itself in an ocean current or was lodged in the hull of a native dugout canoe which in turn floated on the ocean current. To a maple growing in given spot it is immaterial whether its seed flew there on its own wings or whether it was aided and abetted by the wings of an airplane. A wild cherry is unaffected by concern as to whether its seed was deposited by a seagull who spotted a target below or by a human who brought it thither in a paper bag . . . In short, from the viewpoint of the forest, there is no distinction between native and introduced species .  .  . All were migrants there.

Given what is now known about invasive species (Chapter 21), imagine how most naturalists and ecologists would react if such a sentiment was written as fact in a modern textbook! Indeed, the passage was heavily edited for the second edition (Spurr and Barnes 1973) and eliminated by the third (Spurr and Barnes 1980), primarily because the values of ecologists (and society) began to change. Ecological data from the last 58 years suggest that introduced species can be bad because those that become invasive may outcompete native species, alter nutrient dynamics and/or soil communities, have negative economic implications, or may even lead to wholesale changes in a system. And yet relatively few of these data directly contradict Spurr’s initial sentiments; instead, it is implicit in modern ecological studies that humans value native species over introduced species, find the nutrient dynamics of a near-­­natural ecosystem to be more important than of a heavily altered ecosystem, and abhor disruptions to a well-­­functioning economy. Data over the last six decades have also gradually moved toward acknowledging the importance and inevitability of change in ecological systems (Chapter  17), the major point that Spurr was attempting to make. Spurr might be amused today to see that we have finally accepted ecological change—­as long as systems change as we think they should. Human values are typically not scientific, yet they are often overlain onto scientific inquiry, including that undertaken by forest ecologists. These values are what challenge scientists in incorporating notions of sustainability into scientific understanding.

Concepts of Sustainability  641

HISTORICAL PERSPECTIVE OF SUSTAINABILITY IN FORESTS Ours is not a textbook about the management of forests, but a brief discussion here is of use to illustrate how human values are an intrinsic part of sustainability. In fact, if sustainability’s root concern is providing for future generations, then those concerned with forests may be among the first who explicitly integrated sustainability into their operating models (Westoby 1989), even as those models evolved and changed over decades and centuries. First and foremost, the concept of sustaining natural resources over the long term was present in the traditions of many ancient cultures, such as those of the indigenous peoples of New Zealand, Indonesia, Mali, India, and British Columbia (Gadgil and Berkes 1991). Regarding modern times, a review by Wiersum (1995) notes that sustainability was well expressed in forest management systems in Germany in the eighteenth century, and the philosophy of wise use consistent with future availability of forest resources was well articulated in the 1800s. Notably, the creation of the forestry profession in the United States resulted from concerns about sustainable use of forests in the 1900s. Sustainability in some form therefore has a very long association with forests and forest ecology (Franklin et al. 2018). In North America, one aspect of timber management has been deeply associated with at least one form of sustainability for over a century: the idea of sustained yield. Sustained yield is a Western concept that because harvested trees can be regrown, forests are a renewable resource and can be managed for a continuous flow of resources for future generations. Notably, the term “resource” itself automatically invokes human values, both social and economic. Two general problems, among many, are inherent with the idea of sustained yield. First, one cannot assume that all forests are equally renewable, and as such their yields are not always sustainable. For example, aspen in the Lake States is typically clear-­­cut for pulpwood, sometimes as often as every 20–30 years, because the trees regenerate quickly from root sprouts after cutting and exhibit growth rapid enough to replace themselves within the time frame of our current social and economic systems. Likewise, Douglas-­­­fir forests of the Pacific Northwest are also considered quite renewable after harvesting because they may produce sawlogs within a 50-­­year time frame on suitable sites, again within an appropriate social and economic time frame. These models place the emphasis on the potential product (in this case timber) as the resource of interest, such that clear-­­cutting a 500-­­year-­­old (“old-­­growth”) Douglas-­­fir forest and placing it into a rotation of harvesting every 50 years are appropriate to achieve a sustained yield. A second problem with sustained yield is its focus on a single resource provided by the forest. Using the abovementioned example, outside of the timber resource, transitioning an old-­­growth Douglas-­­fir forest to one that produces 50-­­year-­­old sawlogs does not consider the potential value of the old growth itself, which would not be considered to be renewable. The old-­­growth forest may provide a limited habitat for certain species (both plants and animals), or it may provide recreational opportunities or aesthetic values, or it may create an important protection for water resources not found in younger forests. Forest management in North America broadened to multiple-­­use forest management in the 1960s and 1970s in recognition of the idea that sustainability might also target resources other than timber (Clawson 1978). It also more firmly emphasized the importance of human values in its treatment of sustainability. Multiple-­­use management continued to include sustained yield as one of its multiple uses, but transitioned again in the 1990s to ecosystem management (Grumbine 1994; Brussard et al. 1998). Ecosystem management suffered somewhat from varying definitions and a somewhat ambiguous framework, but generally considered the balance between maintenance of biodiversity and ecosystem processes (including disturbances and the changes they invoked in ecosystems) and appropriate human uses. Its emphasis on “whole systems” included social and economic factors as well as ecological ones and hoped to integrate ecological conservation into socioeconomic models. Social factors, in particular, included concepts crucial to sustainability, such as freedom, justice, and equity. Ecosystem management therefore represented a further emphasis of the human factor into sustainability, and many of its basic tenets are recognizable today in concepts such as ecological forestry (Chapter 14).

642  C h a p t e r 2 3   Sustainability of Forest Ecosystems

ECOSYSTEM SERVICES Central to the concept of sustainability is a clear definition of what exactly is being sustained. Simply put, ecosystem services are the objects and processes of natural ecosystems that are ­beneficial to humans, and they are typically the implicit or explicit targets of sustainability. The idea of ecosystem services was described by Daily et al. (2000), codified by the United Nations’ ­Millennium Ecosystem Assessment (MEA 2005), and has since been discussed by many authors (e.g., Nesbitt et al. 2017; Alix-­­Garcia and Wolff 2014; Levin 2012; Perrings 2007; and many others). Notably, the focus of ecosystem services by design is human well-­­being; the ecosystems supplying the services benefit from protection or conservation only because of what they provide to humans rather than from any intrinsic value they have. It is often argued, however, that protecting ecosystem services contributes to the overall function of the ecosphere and in turn the regional and local ecosystems within it. Indeed, it has been recognized that many ecosystem services are irreplaceable, and their loss would somehow harm the integrity of ecosystems, regardless of direct human benefit (Ekins et al. 2003; Farley 2012; Neumayer 2013). Thus, ecosystem services are considered to be a vital bridge between ecosystems and society, and they justify the actions needed to undergo sustainability (Wu 2013, and references therein). Ecosystem services are often grouped into four broad categories (MEA  2005; Carpenter et  al.  2009): provisioning, regulating, supporting, and cultural. Provisioning services, originally defined as “ecosystem goods” by Daily (1997), include the potential products extractable from an ecosystem such as timber, food, water, or minerals, but also genetic and medicinal resources. Regulating services are less tangible than provisioning services because they describe more abstract ­benefits humans obtain when they maintain important ecosystems processes. Examples of regulating services might include pollination, which is important for fruit production and the persistence of plant species, or trophic dynamics, which are important for control of pests and diseases. Supporting services are those that are vital for the functioning of ecosystems and are necessary for most other ecosystem services to exist, such as soil formation, nutrient cycling, or oxygen production. Turner et  al. (2013) also treat primary productivity and post-­­disturbance forest regeneration as major supporting services on specific landscapes. Supporting services are sometimes difficult to distinguish from regulating services because many regulating services may be supportive at large spatial or temporal scales. Finally, cultural services are non-­­material services obtained from ecosystems such as spiritual enrichment, aesthetic encounters, reflection, or recreation. A detailed treatment of ecosystem services in forests is provided by Solórzano and Páez-­­Acosta (2009); a truncated list of examples is shown in Table 23.1. If ecosystem services are a main target of sustainability, then maximizing those services is typically its goal. However, not all ecosystem services of a forest are able to be maximized at the same time, and trade-­­offs ensue. Perhaps the best example of trade-­­offs in ecosystem services would be plantations or tree farms under a sustained yield scenario, where the provisioning service of timber would be maximized at the expense of biodiversity, wildlife habitat, or aesthetics. Likewise, wilderness areas might maximize cultural services such as recreation, spiritual uses, or aesthetics at the expense of provisioning services when potential resources are not extracted (Kinzig 2009). The point is that most services require their own unique set of ecological attributes such that it is difficult to maximize all or even most simultaneously. Interestingly, many ecologists have argued that biodiversity preservation is the keystone of sustainability, because it augments the maintenance of most other ecosystem services through its heavy influence on ecosystem function (e.g., Cardinale et al. 2012). For example, in one of the earlier explorations of sustainability in forests, Lindenmayer and Franklin (2003) edited a book with chapters written by 13 forest ecologists and managers from around the world, many of whom defined sustainability as the preservation of forest biodiversity and one who defined it as “the

Concepts of Sustainability  643

Table 23.1  Non-­­exhaustive list of examples from forest ecosystems of the four categories of ecosystem services. Type of ecosystem service Provisioning

Regulating

Supporting

Cultural

Ecosystem good or function

Examples from forest ecosystems

Forest products

Timber, fiber

Food

Fish, fruits, graze or browse for livestock, wildlife

Genetic resources

Test and assay organisms, bioprospecting

Biochemical/ medicinal resources

Drugs and pharmaceuticals, dyes, latex, and other resins

Ornamental resources

Diversity of organisms with ornamental potential

Climate regulation

Altering temperature, albedo, evapotranspiration

Hydrological regulation

Regulating volume, quality, and timing of water flow

Disturbance moderation

Dampening of wind disturbance, flood control

Carbon sequestration and storage

Cycling and storage of carbon

Pollination

Pollination of timber/fiber species, fruit production

Trophic dynamics

Pest or disease control, reduction of herbivory on crops

Soil formation and support

Physical support for plants, nutrient cycling, water

Post-­­fire forest regeneration

Maintenance of cover types and habitats

Site productivity

Framework for species composition, soil processes

Recreation activities

Tourism, camping, hunting, hiking

Spiritual inspiration

Use of forests for religious or individual experiences

Aesthetics

Appreciation of scenery

Cultural uses

Background for books, music, film, folklore

grand attempt at forest biodiversity rehabilitation.” Although biodiversity is certainly an important attribute of a sustainable forest, its increase may necessarily be accompanied by trade-­­offs with some other services such as food or timber production and may have less or nothing to do with other services such as carbon storage or erosion control (Kinzig 2009; Reyers et al. 2012). Biodiversity appears to typically be most at odds with provisioning services. Naeem (2009) suggested that the ecosystem functions most closely related to biodiversity are generally supporting services, such as net primary production, nutrient cycling, and other biogeochemical processes, but that these services are not consumed and are least understood by people. Others have argued that biodiversity contributes more to regulating services such as invasion resistance and ecosystem resilience, as well as cultural services (Mace et al. 2012). In any case, some have argued that biodiversity is best conserved when ecosystem services are the target rather than biodiversity itself, because ecosystem services provide the economic and social justification that biodiversity has an instrumental value (Skroch and López-­­Hoffman 2009; Mace et al. 2012; Balvanera et al. 2014).

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TOWARD A DEFINITION OF SUSTAINABILITY If sustainability is built upon a balance between ecological factors and human well-­­being, and ­ecosystems services are its target, then how exactly might we define the term? Unfortunately, inconsistencies in its definition have plagued an already complex concept while its study has increased dramatically over the last three decades. Marshall and Toffel (2005) tabulated at least 100 different definitions of the term. As a matter of comparison, we present three such definitions as the most common and familiar to those who think about sustainability. Sustainability has often been described as a “three-­­legged stool,” with equal-­­sized environmental (integrity of ecosystems), social (justice, equity, and freedom), and economic (material goods and services) legs. In this model, also called the “triple bottom line” or the “people–planet– profit” model (Elkington 2004), all three aspects must be met simultaneously; an emphasis on any one more so than the others will not achieve sustainability. These three components have been described as separate but intersecting, best depicted with a Venn diagram where sustainability represents the limited area where all three components intersect, representing their achievement at the same time (Figure 23.1a). In natural resource management, the idea is to properly account for all three components and the interactions among them. The triple bottom line is increasingly used as a way to plan for sustainability and sustainable development in the business world, with the model often used as a metric of a corporation’s commitment to sustainability. If we accept that the triple bottom line describes a basic model of sustainability, then the likelihood that all three components are not equally emphasized must be considered. If the focus is on the entire system rather than any of its separate components, achievable sustainability is considered to be weak. Weak sustainability allows one component to substitute for another in the ­triple-­­bottom-­­­­line model, as long as the entire system increases or sustains its total capital ­(Figure 23.1b). Further, the model assumes there is no difference between natural capital (ecosystems or environment) and human-­­made capital when providing for the well-­­being of humans, and that natural capital lacks unique properties that cannot be substituted by human-­­made capital. Thus, weak sustainability could be achieved on a heavily logged forest landscape if the harvesting supported economic growth of the region and an appropriate standard of living for its people, because the social and economic spheres of the model would substitute for the environment. Wu (2013) contrasts weak sustainability with absurdly strong sustainability, where the environmental

(a)

Triple Bottom Line

(b) Weak Sustainability

Economy

ty

Society

Ec on

om y

Economy

cie

Environment

Strong Sustainability

So

Sustainability

(c)

Environment

Society

Environment

F I G U R E  2 3 .1   Three models relevant to sustainability: (a) the triple-­­bottom-­­line model outlines the basis of sustainability as the intersections of the three spheres of environment, economy, and society, assuming that all three must be met simultaneously to achieve sustainability; (b) weak sustainability assumes that any of the three components is substitutable for another, so long as the entire system increases or maintains its capital; (c) strong sustainability assumes that the three components are nested within each other such that environment constrains society and economy, and society constrains economy. Source: Wu (2013) / Springer Nature.

Concepts of Sustainability  645

sphere is never substituted even when it could be or perhaps should be. Under such a model, species and ecosystems are protected above all else, even at the cost of human well-­­being (e.g., never clearing another forest for agriculture or cutting trees for fuel even if it means people would starve or go without heat in a local village). Thus, weak sustainability assumes that ecosystems and economic products or resources are equivalent in value and therefore may substitute for one another, while absurdly strong sustainability assumes they are polar opposites. A model of sustainability considered strong therefore lies somewhere between these two ends of the sustainability spectrum. Rather than assuming any of the three components are substitutable by the others as in weak sustainability, strong sustainability assumes that each of the three components compliments the others, because natural capital may not necessarily have an analogous human-­­made substitute. Rather than viewed separate from each other, strong sustainability assumes that the three components are viewed hierarchically, with economy nested within society and society nested within the environment ­(Figure 23.1c). In this model, social and economic activities exist within the context of fundamental ecological processes such as nutrient cycling, climate regulation, and water purification, and economy functions within the context of a specific sociological regime (Franklin et al. 2018). Strong sustainability thus firmly places the ecological/environmental aspect at the center of the paradigm. An example of the differences between weak, strong, and absurdly strong sustainability can be illustrated using the susceptibility of western landscapes to fire. The Colorado Front Range is an area undergoing rapid economic development and extensive exurban sprawl into forested, ­fire-­­prone areas. This expansion of the wildland–urban interface (WUI) poses great problems for federal agencies responsible for protecting private landowners from wildfires burning on adjacent National Forests. Moreover, fire suppression in the twentieth century in the area has made the ponderosa pine and mixed-­­conifer forests overly dense and particularly susceptible to large, severe fires with the potential to heavily impact lives, property, and the landscape itself. This context has led to a great many discussions about the sustainability of continued development and the expansion of the WUI, and arguments have varied precisely based on the type of sustainability model under consideration. Using a weak sustainability model, the continued economic growth and provision of property might substitute for the ecological integrity of the landscape, such that appropriate management would include substantial fire suppression in the wake of further expansion of the WUI. On the other hand, an absurdly strong sustainability model might argue that the environment is not substitutable regardless of the consequences for humans, such that any fires that burn on the landscape should be allowed to burn without care for lives or property as a means of allowing altered forests to heal themselves. A strong sustainability model would probably propose a carefully planned development strategy that minimized environmental impact and mitigated risk to private landowners while allowing the natural disturbance regime to persist. Although it would take region-­­wide cooperation and collaboration, extensive regulation and legislation, and corporate support to achieve strong sustainable development in this manner, it is clear that a strong sustainability model is the most appropriate. Different sustainability models will often lead to different landscapes that vary in the ecosystem services they provide (Wu 2013; Figure 23.2). Wilderness areas, for example, that emphasize the preservation of biodiversity and natural ecosystems provide high natural capital and are most appropriately viewed through an absurdly strong sustainability model. Regulating ecosystem services would dominate in such a landscape, and provisioning services would effectively be absent. At the other end of the spectrum, urban landscapes would constitute a weak sustainability model most closely, having low natural capital and providing mostly cultural and regulating services. Agricultural and semi-­­natural landscapes might follow more of a strong sustainability model because they clearly treat social and economic factors within the context of ecological factors. Agricultural landscapes are generally more altered and thus have less natural capital than ­semi-­natural landscapes with a higher proportion of provisioning services, but neither landscape fully substitutes natural capital with human-­­made capital.

646  C h a p t e r 2 3   Sustainability of Forest Ecosystems Relative Importance or Dominance Sustainability View Environmental Emphasis

Absurdly strong sustainability

Strong sustainability

Nature conservation/preservation

Ecological restoration

Weak sustainability Ecological engineering

Ecosystem Services

Supporting Capital Form

Regulating

Natural Capital

Provisioning

Cultural

Built, Human & Social Capitals

Ecosystem or Landscape Type

Natural

Semi-natural

Agricultural

Urban

F I G U R E  2 3 .2   Varying landscapes and their placement on spectra of sustainability views/models, proportion of ecosystem services categories, and relative amount of natural capital. Source: Wu (2013) / Springer Nature.

The idea that different landscapes follow different sustainability models and provide different ecosystem services suggests that scale and spatial heterogeneity have strong implications for this concept. A mosaic of land use types is always present on modern landscapes, such that achieving strong sustainability at broad scales may require targets of weak sustainability (for example, in urban areas) or absurdly strong sustainability (for example, in wilderness areas) at smaller spatial scales (Wu 2013). Moreover, if landscape pattern influences ecological processes (Chapter 22) and biodiversity, then it likely also influences the provision of ecosystem services, making them patchy across space (and potentially across time). If true, then some patterns are more likely to favor the maintenance of ecosystem services, and thus some patterns are more likely to lead to sustainable landscapes (Wu 2013). Notably, the patchiness, or spatial heterogeneity, of some ecosystem services may be important for sustaining them. In a comparison of two forested landscapes in the western United States (the Greater Yellowstone Ecosystem, a subalpine landscape of the central Rocky Mountains; and the coastal temperate rainforest of the Pacific Northwest), Turner et  al. (2013) found that ­post-­disturbance spatial heterogeneity was important for sustaining the ecosystem services of forest regeneration, primary production, carbon storage, natural hazard regulation, insect and pathogen regulation, timber production, and wildlife habitat. Spatial heterogeneity helps to sustain the supporting service of forest regeneration by determining the location of surviving pre-­­disturbance trees that will provide a seed source for regeneration, and this supporting service will in turn also affect other services such as primary production and carbon storage. Regulation of forest insects and pathogens is influenced by heterogeneity because it determines the spatial patterns of stands of susceptible age and size for insect or pathogen attack, as well as how their placement on a landscape might dampen or enhance an outbreak. The provisioning service of timber is clearly affected by site ­heterogeneity as determined by physiography, soils, and microclimate (Turner et al. 2013).

WHERE DO WE GO FROM HERE?

It is not arguable that we are in an age of ecological crisis. Nearly 20 years ago, Franklin (2003) noted that humans have extensively altered the physical and ecological contexts of forests to the point that they no longer resemble those from which the forests evolved. These include altered

Where Do We Go from Here?  647

­ isturbance regimes, especially fire; changes in global and regional climate and chemical regimes d (i.e., climate change and acid rain); introduction of invasive species; and fragmentation of landscapes that have changed the amount and spatial arrangement of various forest conditions. Each of these topics has been covered in detail in other chapters of this book (Chapters 10 and 20–22) and will not be revisited here, but their increased pervasiveness in modern forests has warranted each a detailed treatment that was not presented in this book’s last edition. Franklin (2003) predicted that these four problems would intensify in the twenty-­­first century, and it seems as though he was correct. So how do we ensure sustainability in forests? In as much as our living within an ecological crisis, this is a question still of great debate whose answer is not an easy or unambiguous one, but one that is clearly related to many of the concepts in this book. In a sense, sustainability of forests is subject to the “land sparing versus land sharing” debate (Phalan et al. 2011). One could simply keep humans out of at least some forests and allow the forests to function without human interference. This land-­­sparing argument suggests that intensively managed planation forests (sometimes called “fiber farms,” usually in temperate regions) be used to supply necessary wood products, such that native forests could be left unaltered by humans and thus sustained (e.g., Bremer and Farley 2010; Paquette and Messier 2010). The adoption of fiber farms serves as a kind of ecological offset that allows larger reserve areas elsewhere on a landscape (Green et al. 2005; Fischer et al. 2008). This type of forest allocation is viewed by some as the easiest and most efficient movement toward forest sustainability because often-­­opposing objectives of stakeholders will conflict less often, such that substitution of human-­­made capital for natural capital will be less common. However, many forest managers counter that such a physical separation of forests into provisioning areas and preserves is likely to result in a failure of forests to fulfill the expectations and needs of society as well as undesirable outcomes for ecosystem function and biodiversity (Franklin  2003). These failures are likely because many forest landscapes have already been heavily altered by humans if only indirectly, and thus there are few or no candidate remnants suitable for preservation. Moreover, the lofty expectations humans have for forests in addition to wood products—­such as biodiversity preservation and watershed protection—­are probably not attainable without active management. Critics of land-­­sparing often follow the land-­­sharing model, which integrates different goals (such as biodiversity and/or ecosystem services and resource extraction) in the same area. Proponents of land-­­sharing in forests are primarily those who favor ecological forestry (Puettmann et  al.  2008; Lindenmayer et  al.  2012; Franklin et al. 2018), discussed in Chapter 14, where complex silvicultural systems (harvesting and regeneration) are developed that address more than simply the need for wood products. Of course, critics of land sharing counter that a focus on multiple goals simultaneously is less effective than a focus on either goal individually. In addition, decreased resource extraction that would accompany ecological forestry would necessitate increased extraction elsewhere (Gabriel et al. 2010). There are certainly no easy answers for forest sustainability today, in part because sustainability is achieved by following very different pathways depending on the landscape. Lindenmayer and Cunningham (2013) summarized key principles for sustainable forests. First, there is an important need for improved planning and monitoring to avoid overcommitment of natural resources. To do so, managers and scientists must be able to identify the critical thresholds beyond which resource use will begin to degrade ecosystem services and biodiversity (Carpenter et al. 2011). Second, landscapes vary in their importance to ecosystem services and biodiversity, and thus it is critical to identify those places with disproportionate contribution to species persistence and ecological processes, such as breeding areas, refugia, or biodiversity hotspots. Research begun immediately after major natural disturbance events is often useful in this regard, particularly in understanding how refugia contribute to post-­­disturbance recovery (e.g., Turner et  al. 2003a). Third, it is important to maintain connectivity, but only when connectivity is desirable for specific processes. Given that connectivity could refer to connected habitat for a species or group

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of species of interest or to connected ecological processes, and that some connectivity could lead to the spread of undesirable disturbances or species (Lindenmayer and Fischer 2013), identifying the type of connectivity important for sustainable forest management leaves much room for future research. Fourth, land-­­use practices may accumulate over space and time in such a way that threatens forest sustainability, and such land uses should be avoided. A classic example would be the great eastern white pine forests logged at the end of the nineteenth century in Michigan and other areas of the Lake States. Wholesale forest clearing and subsequent, repeated slash fires converted a widespread dry-­­­­mesic forest landscape to one dominated by xeric species such as jack pine and xeric oaks much more likely to burn. Finally, land uses typically fall upon a gradient between land sharing and land sparing rather than either end of the spectrum and are extremely context-­­dependent. A final point about the prospects of sustainability draws on our original treatment of forest ecology as a study of landscape ecosystems, which integrates the biotic and physical aspects of whole, volumetric ecosystems. In his treatment of sustainable planning and design for large landscapes, Bailey (2002) reasons that sustaining ecosystems must occur at large spatial scales, and that local management must occur within a strong contextual knowledge of regional ecosystems. The ecological crises we described in the earlier text have at least partially resulted from a mismatch of human development and the limits of the regional ecosystems in which they occur. The sense of place provided by regional ecosystems provides patterns and processes that place bounds or limits on the social and economic components of the triple bottom line of that particular place, and thus recognizing the sense of place is necessary to conserve resources (Bailey  2002). Returning to our retrospective example of eastern white pine harvesting in ­Michigan, a lack of understanding and recognition of the cold winter climate, relatively flat topography, and dry, sandy soils that promote fires limited the level of post-­­harvesting forest regeneration and recovery that might have been expected in that regional ecosystem. Unfortunately, human activities exceeded the regional ecosystem’s ecological limits in a way that has threatened its long-­­term sustainability.

EPILOG: EARTH AS A METAPHOR FOR LIFE Thus far our discussion of sustainability has detailed the problems without suggesting solutions (an action Theodore Roosevelt called “whining”). We have been able to discuss in detail what sustainability is, but have been less able to describe how to achieve it in contemporary forests without a detailed discussion of forest management. Again, the fact that sustainability is defined in part by human values makes it extremely difficult—­and undesirable—­to simply stick to the science. As such, we take the opportunity to end this chapter (and this book) with a final abstract, rather philosophical consideration. In thinking about sustainability, perhaps absurdly strong sustainability in particular, certain things such as biodiversity or ecological integrity have intrinsic value, independent of their value or contribution to human well-­­being. Rare or endangered species should be protected because they have value, regardless of how humans may benefit from them. Wilderness areas are not just valuable for the recreation, aesthetic, or spiritual values they provide humans, but also because they are an end in themselves. Ecosystem services, the major target of sustainability, contrast with this notion of intrinsic value because they are human-­­centered (Colyvan et  al.  2009) and have value only based on what they provide for humans (Reyers et al. 2012). Focusing on the ultimate service that ecosystems provide humans is likely to oversimplify how ecosystems function (Ghazoul 2007). In addition, humans need to adapt to ecosystem function because ecosystems lack

Where Do We Go from Here?  649

intentionality in serving humans (Salles  2011). Sustainability by definition is therefore a ­human-­centered construct, illustrated by the simple fact that we need to label ecosystems with social or economic value to justify protecting them. As described in the earlier text, humans have had an enormous, disproportionate influence on the ecosphere and on ecosystems at every scale. We could easily venture into a discussion of understanding humans’ place in the world. Are we part of ecosystems, or external to them? Should we be stewards of ecosystems, regardless of whether or not we consider ourselves as members? Many (not all) people interested in forest ecology are likely to answer “yes” to both of these questions, but do humans deserve to be the center of focus? Rowe (1998) described “Earth” rather than “organism” as a metaphor for “life”. A majority of ecologists and non-­­scientists alike consider organisms to represent living things, but Rowe argues that this is a reductionist view that turns the vantage point of humans inward toward themselves instead of outward toward their surroundings. Throughout this book, we have emphasized that organisms are but one part of ecosystems, and they cannot realistically be separated from physical factors of the Earth (air, water, soil, and energy). Organisms are parts of wholes, and it is the whole ecosystem that provides life to the organisms. In the past (and to some degree the present), ecological research often tended to gravitate toward the charismatic biota, either as the survival physiology of organisms or as the interactions between them. An ecological viewpoint that places organisms at its center, only later turning to what is around them, lends itself to a human-­­centered focus. When the whole ecosphere, rather than just the biota, is considered to be life, we put spaces at the center of our ecological viewpoint and look at its contents (Rowe 1998). Thus, humans become an integral part of ecosystems, but are not central to them. A viewpoint alternative to one that is human-­­centered is not new; many have articulated the dangers of placing humans at the center of the world. In fact, many of the world’s ecological problems have resulted almost exclusively from an anthropogenic view of the world. Aldo Leopold’s land ethic famously moved the focus from humans to the land, with the idea that land was not a commodity owned by humans but a community that included them (Leopold 1949). Sigmund Freud (not an ecologist) suggested that humans make their biggest scientific strides when they realize they are not the center of the universe. For example, physical science leapt forward when Copernicus speculated that the Earth was not the center of the universe; biology advanced greatly when Darwin suggested that humans actually descended from other animals; and psychology developed quickly from Freud’s theory that a personality is driven in part by the unconscious mind, out of a person’s control (Freud 1917). It is certainly true that an additional scientific leap forward will be needed to find new solutions in this age of ecological crises, perhaps one facilitated by encouraging humans to consider that they are a tenant rather than the steward of ecosystems. How does this philosophy translate to sustainability? It suggests that ecosystems have organisms, not the reverse; thus, ecosystems have humans, not the reverse. Rowe (2000) uses an analogy of the heart in a human body to describe the relationship of humans to ecosystems. The role of the heart is to keep the body healthy, not to oversee the body. Likewise, perhaps the role of humans is to keep ecosystems healthy, but as parts of those wholes rather than as their stewards. Transposing this to ecosystem services, ecosystems do not serve humans; in fact, the reverse should be true. We have no illusions to the fact that this view of the world is awkward and would be difficult if not impossible to instill within our current models of sustainability in forests. At the very least, however, we return to the notion of understanding forest ecology, and of ecology in general, as the study of whole ecosystems that include more than organisms (including humans) alone. If nothing else, that resembles a good start.

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SUGGESTED READINGS Bailey, R.G. (2002). Ecoregion-­­Based Design for Sustainability. New York: Springer 223 pp. Balvanera, P., Siddique, I., Dee, L. et al. (2014). Linking biodiversity and ecosystem services: current uncertainties and the necessary next steps. Bioscience 64: 49–57. Daily, G.C. (ed.) (1997). Nature’s Services: Societal Dependence on Natural Ecosystems. Washington, DC: Island Press 412 pp. Lindenmayer, D.B. and Franklin, J.F. (ed.) (2003). Towards Forest Sustainability. Washington, D.C: Island Press 212 pp.

Mace, G.M., Norris, K., and Fitter, A.H. (2012). Biodiversity and ecosystem services: a multilayered relationship. Trends Ecol. Evol. 27: 19–26. Turner, M.G., Donato, D.C., and Romme, W.H. (2013). Consequences of spatial heterogeneity for ecosystem services in changing forest landscapes: priorities for future research. Landsc. Ecol. 28: 1081–1097. Wu, J. (2013). Landscape sustainability science: ecosystem services and human well-­­being in changing landscapes. Landsc. Ecol. 28: 999–1023.

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Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

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Scientific Names of Trees and Shrubs Acacia

Acacia spp.

Ailanthus; tree-­of-­heaven

Ailanthus altissima (Mill.) Swingle

Alaska-­cedar

Chamaecyparis nootkatensis (D. Don) Spach

Alder, green or Sitka

Alnus crispa (Ait.) Pursh.

Alder, red or Oregon

Alnus rubra (Bong.)

Alder, Sitka

Alnus sinuata (Reg.) Rydb.

Alder, speckled

Alnus rugosa (Du Roi) Spreng.

Apple, common

Malus pumila Miller

Arborvitae

Thuja occidentalis L.

Ash, black

Fraxinus nigra Marsh.

Ash, blue

Fraxinus quadrangulata Michx.

Ash, European

Fraxinus excelsior L.

Ash, red or green

Fraxinus pennsylvanica Marsh.

Ash, white

Fraxinus americana L.

Ash, prickly-­

Zanthoxylum americanum Mill.

Aspen, bigtooth

Populus grandidentata Michx.

Aspen, European

Populus tremula L.

Aspen, trembling or quaking

Populus tremuloides Michx.

Baldcypress

Taxodium distichum (L.) Rich.

Basswood

Tilia americana L.

Beech, American

Fagus grandifolia Ehrh.

Beech, Antarctic

Nothofagus antarctica (Forst.) Oerst.

Beech, blue-­

Carpinus caroliniana Walt.

Beech, European

Fagus sylvatica L.

Bigtree; giant sequoia

Sequoiadendron giganteum (Lindl.) Buchholz

Birch, black or sweet

Betula lenta L.

Birch, bog, swamp, or low

Betula pumila L.

Birch, European or silver

Betula pendula Roth.

Birch, gray

Betula populifolia Marsh.

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

731

732  S c i e n t i f i c Names of   Trees a n d   S h ru bs Birch, paper or white

Betula papyrifera Marsh.

Birch, river or red

Betula nigra L.

Birch, Virginia round-­leaf

Betula uber (Ashe) Fern.

Birch, yellow

Betula alleghaniensis Britton

Bitterbrush

Purshia tridentata (Pursh) DC.

Bittersweet, Asian or Oriental

Celastrus orbiculatus Thunb.

Blackgum

Nyssa sylvatica Marsh.

Blueberry, lowbush

Vaccinium angustifolium Aiton

Boxelder

Acer negundo L.

Buckeye, Ohio

Aesculus glabra Willd.

Buckeye, painted

Aesculus sylvatica Bartr.

Buckeye, red

Aesculus pavia L.

Buckeye, yellow

Aesculus octandra Marsh.

Buckthorn, alder

Rhamnus alnifolia L’Héritier

Buckthorn, common

Rhamnus cathartica L.

Buckthorn, glossy

Frangula alnus Mill

Butternut

Juglans cinerea L.

Catalpa, northern

Catalpa speciosa Warder

Catalpa, southern

Catalpa bignonioides Walter

Ceanothus, redstem

Ceanothus sanguineus Pursh.

Cedar, Alaska-­

Chamaecyparis nootkatensis (D. Don) Spach

Cedar, Atlantic white-­

Chamaecyparis thyoides (L.) B. S. P.

Cedar, eastern red-­

Juniperus virginiana L.

Cedar, incense

Libocedrus decurrens Torr.

Cedar, northern white-­

Thuja occidentalis L.

Cedar, Port Orford-

Chamaecyparis lawsoniana (A. Murr.) Parl.

Cedar, western red-­

Thuja plicata Donn

Cherry, black

Prunus serotina Ehrh.

Cherry, choke

Prunus virginiana L.

Cherry, pin or fire

Prunus pensylvanica L. f.

Chestnut, American

Castanea dentata (Marsh.) Borkh.

Chinquapin, golden

Chrysolepis chrysophylla Douglas ex Hook.

Chinquapin, Ozark

Castanea ozarkensis Ashe

Cottonwood, black

Populus trichocarpa Torr. & Gray

Cottonwood, eastern

Populus deltoides Bartr.

Cottonwood, European; black poplar

Populus nigra L.

Cottonwood, Fremont

Populus fremontii Wats.

Cottonwood, narrowleaf

Populus angustifolia James

Cottonwood, plains

Populus deltoides var. occidentalis Rydb.

Creeping strawberry-­bush

Euonymus obovata Nutt.

Creosote bush

Larrea spp.

Cypress, Arizona

Cupressus arizonica Greens

Scientific Names of Trees and Shrubs  733 Cypress, Mexican

Cupressus lusitanica Miller

Cypress, pond

Taxodium distichum var. imbricarium (Nutt.) Croon (syn = T. ascendens Brong.)

Dawn redwood

Metasequoia glyptostroboides H. H. Hu & Cheng

Dogwood, alternate-­leaf

Cornus alternifolia L.

Dogwood, flowering

Cornus florida L.

Dogwood, gray

Cornus foemina Miller

Dogwood, red-­osier

Cornus stolonifera Michx.

Dogwood, roundleaf

Cornus rugosa Lamarck

Dogwood, silky

Cornus amomum Miller

Douglas-­fir

Pseudotsuga menziesii (Mirb.) Franco

Elm, American

Ulmus americana L.

Elm, rock

Ulmus thomasii Sarg.

Elm, Siberian

Ulmus pumila L.

Elm, slippery

Ulmus rubra Muhl.

Elm, winged

Ulmus alata Michx.

Eucalyptus

Eucalyptus spp.

Fetterbush

Leucothoe spp.

Fir, alpine or subalpine

Abies lasiocarpa (Hook.) Nutt.

Fir, balsam

Abies balsamea (L.) Mill.

Fir, European silver

Abies alba Mill.

Fir, Fraser

Abies fraseri (Pursh.) Poir.

Fir, grand or lowland white

Abies grandis (Dougl.) Lindl.

Fir, noble

Abies procera Rehd.

Fir, Pacific silver

Abies amabilis (Dougl.) Forbes

Fir, red

Abies magnifica A. Murr.

Fir, white

Abies concolor (Gord. & Glend.) Lindl.

Giant sequoia; bigtree

Sequoiadendron giganteum (Lindl.) Buchholz

Ginkgo; maidenhair tree

Ginkgo biloba L.

Gum, black; black tupelo

Nyssa sylvatica Marshall

Hackberry

Celtis occidentalis L.

Hackberry, dwarf

Celtis tenuifolia Nuttall

Hackberry, iguana

Celtis iguanaea (Jacq.) Sarg.

Haw, black

Viburnum prunifolium L.

Haw, dotted

Crataegus punctata Jacquin

Hawthorn

Crataegus spp.

Hazel, beaked

Corylus cornuta Marsh.

Hemlock, eastern

Tsuga canadensis (L.) Carr.

Hemlock, mountain

Tsuga mertensiana (Bong.) Carr.

Hemlock, western

Tsuga heterophylla (Raf.) Sarg.

Hickory, bitternut

Carya cordiformis (Wangenh.) K. Koch

Hickory, mockernut

Carya tomentosa Nutt.

734  S c i e n t i f i c Names of   Trees a n d   S h ru bs Hickory, pignut

Carya glabra (Mill.) Sweet

Hickory, shagbark

Carya ovata (Mill.) K. Koch

Hickory, shellbark

Carya laciniosa (Michaux f.) G. Don

Hickory, water

Carya aquatica (Michx. f.) Nutt.

Hobblebush

Viburnum alnifolium Marsh.

Holly, American

Ilex opaca Ait.

Honeysuckle, Amur

Lonicera maackii (Rupr.) Maxim.

Hop-­hornbeam; ironwood

Ostrya virginiana (Mill.) K. Koch

Hornbeam

Carpinus betulus L.

Horse-­chestnut

Aesculus hippocastanum L.

Ironwood; hop-­hornbeam

Ostrya virginiana (Mill.) K. Koch

Juniper, alligator

Juniperus deppeana Steud.

Juniper, common

Juniperus communis L.

Juniper, ground

Juniperus communis L. var. depressa Pursh.

Juniper, one-­seed

Juniperus monosperma (Engelm.) Sarg.

Juniper, Rocky Mountain

Juniperus scopulorum Sarg.

Juniper, Utah

Juniperus osteosperma (Torr.) Little

Juniper, western

Juniperus occidentalis Hook.

Kentucky coffeetree

Gymnocladus dioicus (L.) K. Koch

Kudzu

Pueraria lobata (Willd.) Ohwi

Larch, eastern; tamarack

Larix laricina (Du Roi) K. Koch

Larch, European

Larix decidua Mill.

Larch, subalpine

Larix lyallii Parl.

Larch, western

Larix occidentalis Nutt.

Laurel, mountain

Kalmia latifolia L.

Leatherleaf

Chamaedaphne calyculata (L.) Moench

Lilac

Syringa vulgaris L.

Locust, black

Robinia pseudoacacia L.

Locust, honey

Gleditsia triacanthos L.

Madrone, Pacific

Arbutus menziesii Pursh.

Magnolia, Fraser

Magnolia fraseri Walt.

Magnolia, southern or evergreen

Magnolia grandiflora L.

Magnolia, sweetbay

Magnolia virginiana L.

Magnolia, umbrella

Magnolia tripetala L.

Maidenhair tree; gingko

Ginkgo biloba L.

Mangrove, black

Avicennia nitida Jacq.

Mangrove, red

Rhizophora mangle L.

Maple, ash-­leaf

Acer negundo L.

Maple, bigleaf

Acer macrophyllum Pursh.

Maple, bigtooth

Acer grandidentatum Nutt.

Maple, black

Acer nigrum Michaux f.

Maple, mountain

Acer spicatum Lam.

Scientific Names of Trees and Shrubs  735 Maple, Norway

Acer platanoides L.

Maple, red

Acer rubrum L.

Maple, silver

Acer saccharinum L.

Maple, striped

Acer pensylvanicum L.

Maple, sugar

Acer saccharum Marsh.

Maple, vine

Acer circinatum Pursh.

Mesquite

Prosopis juliflora (Sw.) DC.

Mexican cypress

Cupressus lusitanica Miller

Mimosa

Albizia julibrissin Durazz.

Mountain-­ash, American

Sorbus americana Marsh.

Mountain-­ash, showy

Sorbus decora (Sarg.) Schneid.

Mountain laurel

Kalmia latifolia L.

Mulberry, red

Morus rubra L.

Mulberry, white

Morus alba L.

Myrtle, California wax

Myrica californica Cham. & Schltdl.

Nannyberry

Viburnum lentago L.

Oak, bear

Quercus ilicifolia Wangenh.

Oak, black

Quercus velutina Lam.

Oak, blackjack

Quercus marilandica Muenchh.

Oak, bur

Quercus macrocarpa Michx.

Oak, California black

Quercus kelloggii Newb.

Oak, California live

Quercus agrifolia Née

Oak, canyon live

Quercus chrysolepis Liebm.

Oak, cherrybark

Quercus pagoda Raf. (syn. Q. falcata var. pagodifolia 111.)

Oak, chestnut or rock chestnut

Quercus prinus L.

Oak, chinquapin or yellow

Quercus muehlenbergii Engelm.

Oak, diamond leaf

Quercus laurifolia Michx.

Oak, dwarf chestnut

Quercus prinoides Willd.

Oak, Emory

Quercus emoryi Torr.

Oak, English, European, or common

Quercus robur L.

Oak, Gambel

Quercus gambelii Nutt.

Oak, holm

Quercus ilex L.

Oak, interior live

Quercus wislizeni A. DC.

Oak, live

Quercus virginiana Mill.

Oak, northern pin

Quercus ellipsoidalis E. J. Hill

Oak, Oregon

Quercus garryana Dougl.

Oak, overcup

Quercus lyrata Walt.

Oak, pin

Quercus palustris Muenchh.

Oak, post

Quercus stellata Wangenh.

Oak, northern red

Quercus rubra L.

Oak, sand live

Quercus geminata Small, [syn. Q. virginiana var. maritima (Michx.) Sarg.]

736  S c i e n t i f i c Names of   Trees a n d   S h ru bs Oak, scarlet

Quercus coccinea Muenchh.

Oak, shingle

Quercus imbricaria Michx.

Oak, southern red

Quercus falcata Michx.

Oak, swamp chestnut

Quercus michauxii Nutt.

Oak, swamp laurel

Quercus laurifolia Michx.

Oak, swamp white

Quercus bicolor Willd.

Oak, turkey

Quercus laevis Walt.

Oak, water

Quercus nigra L.

Oak, white

Quercus alba L.

Oak, willow

Quercus phellos L.

Olive, Russian

Elaeagnus angustifolia L.

Osage-­orange

Madura pomifera (Raf.) C. K. Schneid.

Pacific silver fir

Abies amabilis (Dougl.) Forbes

Palm, cabbage

Sabal palmetto (Walt.) Lodd.

Palmetto, cabbage

Sabal palmetto (Walt.) Lodd.

Palmetto, saw

Serenoa repens (Bartr.) Small

Pawpaw

Asimina triloba (L.) Dunal

Pine, bishop

Pinus muricata D. Don

Pine, black or Austrian

Pinus nigra Arnold

Pine, bristlecone

Pinus aristata Engelm.

Pine, Caribbean

Pinus caribaea Morelet

Pine, digger

Pinus sabiniana Dougl.

Pine, eastern white

Pinus strobus L.

Pine, erectcone or Calabrian

Pinus brutia Ten.

Pine, Jack

Pinus banksiana Lamb.

Pine, Jeffrey

Pinus jeffreyi Grev. & Balf.

Pine, knobcone

Pinus attenuata Lemmon

Pine, limber

Pinus flexilis James

Pine, loblolly

Pinus taeda L.

Pine, lodgepole

Pinus contorta Dougl.

Coastal lodgepole pine

Pinus contorta ssp. contorta

Mendocino White Plains lodgepole pine

Pinus contorta ssp. bolanderi

Rocky Mountain-­Intermountain lodgepole pine

Pinus contorta ssp. latifolia

Sierra–Cascade lodgepole pine

Pinus contorta ssp. murrayana

Pine, longleaf

Pinus palustris Mill.

Pine, Mexican pinyon (piñon)

Pinus cembroides Zucc.

Pine, Mexican white

Pinus ayacahuite Ehrenb.

Pine, Monterey or radiata

Pinus radiata D. Don

Pine, mugo or mountain

Pinus mugo Turra. [P. montana Mill.]

Pine, patula

Pinus patula Schl. & Cham.

Scientific Names of Trees and Shrubs  737 Pine, pinyon (piñon)

Pinus edulis Engelm. or P. monophylla Torr. & Frem. or P. cembroides Zucc. or P. quadrifolia Parl. ex Sudw.

Pine, pitch

Pinus rigida Mill.

Pine, pond

Pinus serotina Michx.

Pine, ponderosa

Pinus ponderosa Laws.

Pine, red

Pinus resinosa Ait.

Pine, sand

Pinus clausa (Chapm.) Vasey

Pine, Scots or Scotch

Pinus sylvestris L.

Pine, shortleaf

Pinus echinata Mill.

Pine, slash

Pinus elliottii Engelm. var. elliottii

Pine, sugar

Pinus lambertiana Dougl.

Pine, stone or Swiss stone

Pinus cembra L.

Pine, table mountain

Pinus pungens Lamb.

Pine, Torrey

Pinus torreyana Parry ex Carr.

Pine, Virginia

Pinus virginiana Mill.

Pine, western white

Pinus monticola Dougl.

Pine, whitebark

Pinus albicaulis Engelm.

Planetree, oriental

Platanus orientalis L.

Plum, Allegheny

Prunus alleghaniensis Porter

Poplar, balsam

Populus balsamifera L.

Poplar, black hybrid

Populus x euramericana

Poplar, European white

Populus alba L.

Poplar, Lombardy

Populus nigra ‘italica’ Muenchh.

Prickly-­ash

Zanthoxylum americanum Mill.

Privet, swamp

Forestiera acuminata (Michx.) Poir.

Redbud, eastern

Cercis canadensis L.

Redwood

Sequoia sempervirens (D. Don) Endl.

Redwood, dawn

Metasequoia glyptostroboides H. H. Hu & Cheng

Rhododendron, rosebay

Rhododendron maximum L.

Rose

Rosa spp.

Sagebrush, big

Artemisia tridentata Nutt.

Saguaro; giant cactus

Cereus giganteus Engelm.

Salmonberry

Rubus spectabilis Pursh.

Sassafras

Sassafras albidum (Nutt.) Nees

Sequoia, giant; bigtree

Sequoiadendron giganteum (Lindl.) Buchholz

Serviceberry, downy

Amelanchier arborea (Michaux f.) Fernald

Serviceberry, smooth

Amelanchier laevis Wieg.

Sheepberry

Viburnum lentago L.

Silver bell; Carolina silverbell

Halesia carolina L.

Snowbrush

Ceanothus velutinus Doug. ex Hook

738  S c i e n t i f i c Names of   Trees a n d   S h ru bs Sourwood

Oxydendrum arboreum (L.) DC.

Spruce, black

Picea mariana (Mill.) B. S. P.

Spruce, Colorado blue

Picea pungens Engelm.

Spruce, Engelmann

Picea engelmannii Parry

Spruce, Norway or European

Picea abies (L.) Karst.

Spruce, red

Picea rubens Sarg.

Spruce, Sitka

Picea sitchensis (Bong.) Carr.

Spruce, white

Picea glauca (Moench) Voss

Sugarberry

Celtis laevigata Willd.

Sumac, smooth

Rhus glabra L.

Sweetgum

Liquidambar styraciflua L.

Sycamore, American

Platanus occidentalis L.

Tallow, Chinese

Triadica sebifera (L.) Small

Tamarack; eastern larch

Larix laricina (Du Roi) K. Koch

Tamarix, five-­stamen

Tamarix pentandra Pall.

Torreya, Florida

Torreya taxifolia Arn.

Tree-­of-­heaven; ailanthus

Ailanthus altissima (Mill.) Swingle

Tuliptree; yellow-­poplar

Liriodendron tulipifera L.

Tupelo, water or swamp

Nyssa aquatica L.

Virginia creeper

Parthenocissus quinquefolia (L.) Planchon

Walnut, black

Juglans nigra L.

Willow, black

Salix nigra Marsh.

Willow, crack

Salix fragilis L.

Willow, peachleaf

Salix amygdaloides Andersson

Willow, sandbar

Salix interior Rowlee

Willow, weeping

Salix babylonica L.

Witch-­hazel

Hamamelis virginiana L.

Yellow-­poplar; tuliptree

Liriodendron tulipifera L.

Yew, Canada

Taxus canadensis Marshall

Yew, Florida

Taxus floridana Nutt.

Yew, Pacific

Taxus brevifolia Nutt.

Index Note: Page numbers followed by “f” and “t” figures and tables, respectively.

A

abiotic environment, 4 aboveground net primary productivity (ANPP), 473, 477f, 506, 635, 637 for alpine, boreal, temperate, and tropical terrestrial ecosystems, 474f relationship with aboveground litter production, 507f soil properties, forest biomass, and, 475–476 terrestrial ecosystems, 473f water balance and, 476f absorption spectra, 133 abstract space, 299 acidity of soil, 210–212 actinobacteria, 498 action spectra, 133 active seed bank, 71t actual evapotranspiration (AET), 474, 474f adaptation to climate change effects on forests, 581–587 of fire in forest trees, 224–227 local, 55 adaptedness, 35 adaptive introgression, 62 adhesion, 203 adult phase, 71–73 advanced regeneration, 425 aerial roots, 104 aggradation, 426 aggrading, 423 aggregates of soil, 201–202, 233 agricultural landscapes, 647 albedo, 124 alfisols, 214 algal crusts, 497 alien non‐native species, 590 alleles, 37 allogenic coexistence, 444 allogenic succession, 416 allometric biomass equations, 470

allopatric speciation, 57 allozymes, 40 alpha diversity, 332, 333–335, 334f alpine tree lines, 302–303 altered fire regimes, 600–601 alumina octrahedra structure, 208, 208f American beech, 444–446, 445f American chestnut trees, 411, 603, 605f American elm, 411, 605 amino acids, 515 andisols, 214 angiosperms, 71n1 animals in forest ecosystems, 269 animal predation effect on seed availability, 288t change in white tailed deer populations, 286f elk browsing, 287f influence of livestock on forest ecosystems, 288–290 anthesis, 74, 75 antifreeze, 159 apical control, 95, 97 apical dominance, 96, 97 apomixis, 39 apparency hypothesis, 270–271 arbuscular mycorrhizae (AM), 504–505, 505f architectural models of trees, 97–102, 98f canopy architecture of individual species, 98–99 patterns of intermittent growth, 99–102 short and long shoots, 98 sylleptic and proleptic shoots, 102 asexual reproduction of woody plants, 70, 71, 388. See also sexual reproduction of woody plants asexual systems, 39 aspect of sloping terrain, 171, 172–173 aspen–birch forests, 563 assimilation, 500 assisted colonization. See assisted species migration assisted gene flow. See assisted population migration assisted genetic migration. See assisted population migration

Forest Ecology, Fifth Edition. Daniel M. Kashian, Donald R. Zak, Burton V. Barnes, and Stephen H. Spurr. © 2023 John Wiley & Sons Ltd. Published 2023 by John Wiley & Sons Ltd.

739

bindex.indd 739

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740  Index assisted long‐distance migration. See assisted species migration assisted population migration, 581 assisted range expansion, 581–582 assisted species migration, 581–583, 582f associations of species, 296 atmospheric CO2, 540 atmospheric deposition, 492–495, 494f, 496f atmospheric water potential, 204 autecology, 12 autogenic succession, 416 available water content, 205 avoiders, 227, 228 Azotobacter, 497

B

balsam poplar, 43 beech bark disease, 599 belowground net primary productivity, 482–485. See also aboveground net primary productivity (ANPP) beta diversity, 332, 335–336 biodiversity, 327, 332 in ecosystem stability, 356–357 preservation, 358, 644–645, 649 value of, 329–331 biodiversity–productivity relationship, 354–356 bioecosystem approach, 4 biological control for invasive species, 610 biological diversity. See biodiversity biological legacies, 385–386, 414, 415t biomass distribution, 467t of forest ecosystems, 466–469 measurement, 469–473 biomass accumulation during ecosystem development, 477, 480f biomass pools in second‐and old‐growth ecosystems, 482t changes in processes controlling, 478f natural disturbance, 481 net ecosystem production, 479–480 biomes, 20–25, 291 distribution of, 23f of Eurasia and North America, 24f biota, 237 biotic community, 3 biotic composition changes, 411 addition of species, 412 elimination of species, 411–412 biotic regions. See biomes birds, 279–281 bitternut hickory, 378–379, 379f black locust, 498, 602

bindex.indd 740

bole, 226 boundary layer of leaf, 151 branch habit and self‐pruning ability, 225–226 brown‐rot fungi, 517 browsers, 269, 288 bud burst, 46, 145, 161–163, 551 building. See aggrading buttresses, 110

C

Cajander’s method, 263 California acorn woodpecker (Melanerpes formicivorus), 279, 279f California and Pacific Northwest mountainous terrain, 176–180 Canadian Forest Ecosystem Classification System (CFEC), 263 canopy closure, 308 canopy packing, 129 canopy structural complexity (CSC), 130 canopy structure, 125, 127f CSC, 130 effects on light, 125 and LAI, 126, 127–129 structural differences in, 126 carbohydrates, 458 carbon (C), 449, 487 biochemical constituents of needles and shoots, 457t budget of 14‐year‐old Scots pine tree, 459t dynamics, 635 emissions for forests, 586f flows, 449 gain, 574 gain hypothesis, 319 light and C allocation, 461 loss, 575–577 and nutrient cycling, 601–602 pools, 575 sequestration, 573 sink, 637 soil nitrogen availability and C allocation, 461–465 source, 637 carbon balance, 449 controlling processes of terrestrial ecosystems, 468f of ecosystems, 465–482 of trees, 450–465 carbon dioxide (CO2), 13, 123, 207, 540 exchange system, 470, 471f fertilization, 549–550 carnivorous bird species, 622 catastrophic and local land movements, 407 cation exchange capacity of soil, 209–210 cellulose, 515

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Index  741 central Appalachians forests and physiography, 180–183 cheatgrass, 600, 601f, 602 chemical control for invasive species, 610 chemical defenses, 270 chestnut blight, 598, 603–606 chlorite, 208 chloroplast DNA markers (cpDNA markers), 62–63 chronosequence, 416 clay mineralogy, 207–209 Clements concept of ecology, 297–298 Clementsian succession, 419–420, 420f climate, 12, 473–475, 539 of broadscale ecosystem, 25, 26 climate‐disturbance–carbon interactions, 580 envelopes, 558 regulation, 647 climate change, 331, 449, 539 adapting to climate change effects on forests, 581–587 effects on forest carbon, 573–581 effects on forest disturbances, 563–573 effects on forest tree, 548–556 effects on tree species distributions, 556–563 global temperature changes, 541f precipitation effects, 545–547 refugia, 585f temperature effects, 542–544 variation in temperature and atmospheric CO2 concentration, 540f climatic classification, 18 ecoclimatic zones of world, 18f vertical zonation, 19f zonal relationships between climate, soil orders, and vegetation, 19t climatic maximum. See climax climax, 297, 419 climatic, 431 monoclimax, 419 pattern of vegetation, 432 physiographic, 431 site, 432 clines of glaucousness, 57 of trees, 41 clone, 39 formation, 70 maintenance or expansion, 71t trembling aspen clones, 91, 92f, 92 vine maple clone, 92, 93f closed‐cone pines, 228–230 cloud deposition, 492 coarse woody debris, 577 co‐dominant crown classes, 309

bindex.indd 741

coevolutionary systems, 42–43 cogongrass, 600, 601f cohesion, 203 cold acclimation process, 156 cold‐air drainage effects on forests, 148–150 cold injury to plants, 153–154 Colorado Front Range, 647 common garden approach, 37 communities, 291, 296 competition and niche differentiation, 305 composition and structure, 349 continuum concept, 298–299 grid of species interactions, 307t grounding, 292–296 interactions among organisms, 307–315 as landscape ecosystem property, 299–300 pattern of co‐occurrence of species on landscape, 300f potential and optimum range of tree species, 306f schools and terminology, 296–298 spatial variation in forest communities, 300–305 type, 297 understory tolerance, 315–324 compensatory dynamics, 356 competition, 305, 308, 600 changing composition in northern hardwood forest, 314t differentiation of trees of pure even‐aged stand, 309f forest community structure and composition, 308–313 and overstory composition, 313–314 tree diameter distributions, 310f in understory, 314–315 competitive exclusion principle, 308 complementarity, 354 compounded disturbances, 409 coniferous trees. See conifers conifers, 73, 75, 115, 135, 222, 548 growth rings, 112 pollination and fertilization in, 74 vegetative reproduction by, 92, 93 connectivity, 622–624 consolidated parent material, 195–197, 196f construction respiration, 456–458 continental relationships, 343–344 continental scales, diversity at, 343–346 continuum concept, 298–299 corridors, 618 cowbirds, 621 crown classes, 309–310 crown dieback, 80 crown fires, 220, 221 cultural control for invasive species, 609 cultural services, 644, 645t

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742  Index

D

damage avoiding fire damage, 224–226 and death, 286–288 heat damage in seedlings, 151–152 principles of wind damage, 401–403 recovering from fire damage, 226 severity, 624 dark respiration, 133, 450–452 de‐acclimation, 159 dead wood, 479–480 deciduousness and temperature, 164 declining. See senescing decomposers, 269 decomposition, 285–286, 575–577 decomposition dynamics, 517 abscised leaf, 519f changes in C and N contents of leaf litter, 522f rate constants, 518t, 520f decurrent form of trees, 96f, 97 deep rooting, 225, 226 definite habitat, 42 degenerating. See senescing deglaciated terrain, primary succession on, 434–437, 435f, 436f degree days, 144–145 de‐hardening, 159 dehydration, 159 deliquescent form of trees. See decurrent form of trees dendrochronology, 115 denitrification, 526–528 density‐dependent competition, 312 deposition velocity, 495 derecho, 401, 573 determinate growth of plants, 98 detritivores, 269 diameter at breast height (DBH), 444, 469–470 differentiation, 335–336 differentiation diversity, 332 diffuse‐porous species, 112 diffusion, 499 digestibility reducers, 270 digital elevation models (DEMs), 172, 183 dioecious condition, 73 direct exploitation of species, 331 discrete forest communities, 300. See also merging forest communities alpine tree lines, 302–303 Coastal California, 301–302 forest–grassland ecotone, 302 marine terraces, 301f disease, 406–407 disturbance(s), 349–351, 381–384, 425 ball‐in‐cup diagrams, 384f biotic composition changes, 411–412

bindex.indd 742

catastrophic and local land movements, 407 as ecosystem process, 384–386 effects on heterogeneity, 627–632 events, 15, 16, 16f fire, 387–400 floodwater and ice storms, 404–405 in forest ecosystems, 387 heterogeneity effects on, 624–627 historical range of variability, 632–634 insects and disease, 406–407 interactions, 409–411, 410f land clearing, 409 on landscapes, 624 linked, 409 lodgepole pine trees, 406f logging, 407–408 refugia, 584 regeneration wave, 405f regimes, 16, 16f, 383 source, 386 wave‐regenerated fir species, 404 by wind, 174, 175, 400–404 diversity, 327. See also biodiversity biological and ecosystem, 327 causes of species, 343–351 conserving ecosystem and biological, 363–364 ecosystem, 327, 342 focal species in conserving, 351–354 forest management and, 357–363 and functioning of ecosystems, 354–357 ground‐cover species, 337–341 index of, 334 landscape ecosystems, 336–337 levels, 331–333 measuring, 331–342 value of species, 328–331 dominant crown classes, 309 dominant unions, 247 dormant/dormancy, 155–156 degree of, 161 and persistent seed bank, 71t transitions in seasonal growth–dormancy cycling in trees, 162f of woody plants, 82–83, 82f Douglas‐fir, 590 forests, 643 genetic differentiation of, 52–54 germination and survival of seeds of, 86f in Pacific Northwest, 394 pollination and fertilization in, 74 regeneration, 555f reproductive cycle of, 75f study of seedling survival of, 85 downgrading. See senescing droughts, 547, 572

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Index  743 drought tolerance, 117 dry deposition, 492, 494–495 dry‐matter accumulation, 388 Dutch elm disease, 598–599, 603–606 dwarf shoots of plants. See short shoots of plants

E

early‐successional species, 85, 86 earlywood formation control, 114–115 Earth as metaphor for life, 650–651 earthworms, 607 eastern North American species, 47–49 eastern redcedar, 281, 281f ecoclimatic zones, 27 ecological/ecology, 3, 4 processes, 634–640 regionalization, 26 responses, 16 separation, 58 species groups, 245–247 system. See ecosystem ecological forestry, 359‐363, 643 ecological land classification (ELC), 262–264 ecological site units (ESUs), 292, 293f economy functions, 647 ecosystem change, 431–432 light and, 141 in southern Appalachians, 373 ecosystems, 3, 4, 7, 228, 642, 651. See also forest ecosystems carbon balance and retention and loss of nutrients, 531–532 complete, 13 delineation, 28 disturbances as ecosystem process, 384–386 diversity, 327, 342 functioning, 354–357 geography, 25 management, 643 and mapping in Baden‐Württemberg, 257–258 and mapping in Michigan, 258–260 non‐exhaustive list, 645t services, 329, 644–645, 650 in southeastern United States, 261–262 in southwestern United States, 262 types, 433 ecosystems, carbon balance of, 465 biomass accumulation during ecosystem development, 477–482 biomass and productivity of, 466–469 climate and productivity, 473–475 measurement of biomass and productivity, 469–473 soil properties, forest biomass, and ANPP, 475–476 ecotone, 293, 301

bindex.indd 743

ecotype of trees, 42 ectomycorrhizal fungi (ECM), 504–505, 505f edaphic characteristics of soil, 12 eddy correlation technique, 472 eddy covariance method, 471 edge effects, 621 edge species, 621 effective depth, 254, 256 elements situated in landscape, 613 elevation of landforms, 170, 344–346 elm–ash–cottonwood, 563 emerald ash borer, 596, 603, 604f endangered species, 353–354 endemics, 353–354 endemism, 353 endurers, 227, 228 enemy release hypothesis, 591 entisols, 214 environmental factors as measure of site, 251 aspect, slope steepness, and site index for black oak, 255f climatic factors, 251–252 land types characteristic of dissected forest terrain, 253f physiographic and soil factors, 252–256 physiographic land classification, 252 productivity curve with site‐index ratio, 255f soil surveys, 256 environmental filters, 595 epicormic branches, 139 epicormic sprouting, light and, 139 epigenetics, 66–67 epigeous seed germination, 83, 84f, 85 epsilon diversity, 332 erosion, 234–236, 234f establishment stage. See stand initiation ethological separation, 58 Eurasian wild boars, 607 evaders, 227, 228 evapotranspiration (ET), 635, 636f, 475 evenness of species, 331–334, 629 evolutionary biology in sexual populations, 38 natural selection, 38–39 sexual and asexual systems, 39 excurrent form of trees, 95–96, 96f exotic non‐native species, 590 exposure time, 151 extensive tree mortality, 603

F

facilitation, 307, 354, 428–431, 429f fascicles, 98 feral swine, 607 fertilization effect, 549 fiber farms, 649

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744  Index field capacity, 204, 205 fine‐grained mica, 208 fine roots, 105–107, 107f, 463 influence of soil N availability, 465f life span, 463 mortality, 464 firebreaks, physiography and, 190–191 fire cycle. See fire rotation fire frequency, 220, 221 fire‐induced opening of cones, 71t fire in forest trees, 220, 230, 387, 565. See also forest trees adaptations, 224–227 avoiding fire damage, 224–226 beneficial effects of, 236 in boreal forest and taiga, 397–398 causes, 220 colonizing burned‐over areas, 226–227 direct effects of, 232–233 exclusion, 399–400 fire‐scar cross section of ponderosa pine, 223f fire regime, 220–224, 225f in forest ecosystems, 387–395 history and behavior, 395–399 indirect effects of, 230–232 key characteristics of forest species, 224 mean annual number, 567f and oak dominance, 446–447 organic matter removal and erosion, 234–236 processes directing forest conversion due to, 568f promoting fire occurrence, 227 recovering from fire damage, 226 roles in landscape ecosystems, 219 strategies of species persistence, 227–230 suppression, 399–400, 634, 647 types, 221 fire in northern Lake States, 395–397 fire in northern Rocky Mountains, 398–399 fire intensity, 220, 221, 222 fire‐resistant live foliage, 226 fire rotation, 220 fire‐triggered dehiscence of fruits and cones, 227 fire–vegetation feedbacks, disruption of, 569–570 fish, 278 fitness–flexibility compromise, 66 flagship species, 351 flammable foliage and bark, 227 flatlands, 183 floodplains, 186–190 Great Plains, 184–185 pine savannas of western Great Lakes Region, 185 till plains of Midwest, 185 southeastern and southern Coastal Plain, 186 flat terrain, 171 flooding/floodwater, 404–405, 628

bindex.indd 744

floodplains, 186 changes in physiography of, 190 forests, 563 landform zonation pattern, 187, 188 Mississippi River floodplain, 187f in outwash plains and outwash over moraine physiography, 189f similar effects, principle of, 187 variation in types of landforms, 189 Florida Keys, 292–293 flowering, 226 mast, 76 of woody plants, 72, 73, 75 flushing, 46 focal species in conserving diversity, 351 endemics and rare and endangered species, 353–354 foundation species, 352 keystone species, 352–353 tree species richness, 353f, 355f foliage retention near ground, 227 forcing, 161 forest, 3, 197 applicability to forest management, 12–13 biomass, 475–476 canopy closure event, 425 direct measurement of forest productivity, 238–239 forest–grassland ecotone, 302 giant and pygmy, 301–302 harvesting and nutrient loss, 532 interior species, 621 projected changes in forest type distributions, 561–563 short‐term changes in flow of water and nutrients, 534f temperature within, 147–148 types, 3, 263, 297 forest carbon carbon gain, 574 carbon loss, 575–577 climate change effects, 573 dynamics, 637–638 feedback among disturbance, climate change, and carbon, 577–581 hypothetical trajectory, 578f management, 585–587 forest community, 3 discrete forest communities, 300–303 merging forest communities, 303–305 numbers of trees per hectare, 311f stand density, 312–313 structure and composition, 308 vertical structure, 310–312 forest disturbances, climate change effects on, 563. See also disturbance(s) fires, 565–570 insects and pathogens, 570–572 wind, 572–573

11-14-2022 20:52:31

Index  745 forest ecology, 3, 641. See also landscape ecology applicability to forest management, 12–13 approach to study of, 11–12 landscape ecosystems, 3, 4–11 prevalence of human values in, 641–642 forest ecosystems, 12, 449. See also ecosystems; long‐ term forest ecosystem and vegetation change biomass and productivity of, 466–469 influence of livestock on, 288–290 instrumentation, 472f mean residence time of organic matter, nitrogen and phosphorus, 514t mycorrhiza, 504–506 nitrogen content of leaf and fine‐root litter, 511t nutrient cycling within, 499 nutrients in forest floor, 511–513 nutrient transport to roots, 499–500 nutrient uptake and assimilation by roots, 500–503 organic matter decomposition and nutrient mineralization, 513–524 plant litter and return of nutrients to forest floor and soil, 506–511 plant root systems, 503f root architecture, 503–504 understory tolerance examples in, 319 forest fragmentation, 614, 619 connectivity, 622–624 distribution of eastern white pine, 619f ecological effects, 619–622 habitat loss on landscape pattern, 620f patches in forest ecology, 618–619 forest management and diversity, 357 ecological forestry, 359–363 effects on diversity, 357–358 in North America, 643 preserving diversity in managed forests, 358–359 forestry‐assisted migration, 583 forest succession, 413, 603–606 change in ecosystems, 431–432 fire and oak dominance, 446–447 following eruption of Mount St. Helens, 437–439 primary succession on deglaciated terrain, 434–437 secondary succession and gap dynamics, 439–446 secondary succession following fire in ponderosa pine forests, 439 forest trees. See fire in forest trees components of phenotypic variation, 35–37 ecological considerations at species level, 57–60 epigenetics, 66–67 evolutionary sequence, 38–39 fitness–flexibility compromise, 66

bindex.indd 745

genecology, 40–57 genetic diversity of woody species, 39–40 growth and mortality, 548–550 hybridization, 60–63 maximum photosynthetic capacity of, 450 niche, 60 phenology, 550–553 polyploidy, 63–65 regeneration, 553–556 resources for growth and development, 195 sources of genetic variation, 37–38 formal ecological theory, 417 fossil charcoal, 220 foundation species, 351–352 Frankia, 498 free‐air CO2 enrichment (FACE), 549 free‐living cyanobacteria, 497 frost hardiness and cold resistance, 156 freeze avoidance by plants, 157, 159 hardening and de‐hardening of twigs, 158f mechanisms of freezing resistance, 158f plants cope with freezing temperatures, 157 regression of killing temperature, 157f thermotropic movements in rhododendrons, 159–161 frost pockets, 148 fuel moisture, 623 functional component of niche, 60

G

gametic incompatibility, 58 gamma diversity, 332 gap dynamics, 439–446 gap‐phase species, 87–88, 423 gap specialists, 444–446 gene flow in plants, 38, 56–57, 59–60 mutation, 37–38 recombination, 37, 38 upregulation, 50 genecology, 40–41 clinal variation of critical night length, 47f examples of genecological differentiation, 46–54 factors affecting differentiation, 56–57 factors eliciting genecological differentiation, 42–46 genecological categories, 42 local genecological differentiation, 55 patterns of genecological differentiation, 41–42 genet, 71 genetic diversity/variation of forest trees of ponderosa pine, 51 sources of, 37–38 of woody species, 39–40 genome sequencing, 59 genomics, 59 genotype, 11, 35

11-14-2022 20:52:31

746  Index genotype–environment interaction, 35, 36 geoecosystems, 4 geographic ecosystems. See landscape ecosystems geographic separation, 57 geometric series, 335 geomorphic processes, 16 geomorphology, 168 germination and establishment, 284–285 of woody plants, 82–83, 82f giant sequoia, 394–395, 396f gibberellic acid application, 73 glacial microrefugia, migration from, 377–379 glacial refugia, 583 glaciated landscapes, 171 glaciation, 344 Gleason concept of ecology, 297–298 global warming, 542. See also climate change glucose, 457 gradient analysis approach, 298 grain, 15 grass stage, 225, 393 gravitational potential of soil water, 204 grazers, 269 Greater Yellowstone Ecosystem, 569, 578 Great Plains ecosystems, 184–185 Great Smoky Mountains, 303 greenhouse effect, 540 green plant chloroplasts, 123 gross photosynthesis, 450 gross primary production (GPP), 467, 468t, 470, 531 ground‐cover species diversity in northern Lower Michigan, 337 alpha diversity indices for, 339t alpha diversity measure for upland ecosystem groups, 341t beta diversity measures for ecosystem groups, 340t ecosystem groups, 337–340, 338f of ecosystems, 335, 336f ecosystem types, 340–341 site conditions of six landscape ecosystem groups, 338t ground fires, 220, 221 grounding communities, 292 Florida Keys, 292–293 interior Alaska, 294 southern Illinois, 294–296 growing season length, 44, 45, 144, 543, 550 growth cessation of trees, 43–46 growth–differentiation balance hypothesis, 271–272 growth‐intercept method, 243 growth resumption of trees, 46 growth rings of trees, 111–112 gusts, 401, 402f gypsy moth. See spongy moth

bindex.indd 746

H

habitat loss and fragmentation, 331 habitat type, 247 heartwood, 111 heat capacity of soils, 146 heat‐induced germination, 227 heat sum. See temperature sum heavy‐seeded species, 81 hemicellulose, 517 hemlock seedlings, 60 hemlock woolly adelgid, 352, 572, 602, 603, 606, 610 herb diversity, 345–346 herbivores, 269 heritability, 36 heterogeneity effects on disturbances, 624–627 heterotrophic respiration, 575–577 high earthworm biomass, 607 Hills’ physiographic approach, 263, 264f historical range of variability (HRV), 632–634 histosols, 214 horizons of soil profile, 198, 198f A horizon, 199 C horizon, 199 O horizon, 198, 199 R horizon, 199 host resistance, 610 human‐caused disturbances, 381 hurricanes, 401, 572 1938 Hurricane, 404, 624–625 broadscale disturbance by, 403–404 generalized pathways, 625f Harvard Tract of Pisgah National Forest, 626f Hurricane Andrew, 403 in New England, 624–627 post‐hurricane damage classes, 626f hybrid bridge hypothesis, 276 hybrid inviability, 59 hybridization, 38, 59, 60–63 hybrid sterility, 59 hybrid weakness, 59 hydrarch succession, 420, 421t hydrogen (H), 487 hypertrophy, 110 hypogeous seed germination, 83, 84f, 85

I

ice‐contact terrain, 171 ice storms, 404–405 igneous rocks, 195, 197t illuminance, 124 illuviation, 199 importance value, 559 inceptisols, 214 indeterminate growth of plants, 98 indicator plants/species, 246, 351

11-14-2022 20:52:31

Index  747 individualistic concept of ecology, 297 inhibition mechanism, 428–431, 429f initial floristic composition (IFC), 422–423 injury of plants, 273–274 insects, 406–407, 570–572, 571f herbivores with distributions in hybrid zone, 276f lower bole of lodgepole pine, 273f performance and simultaneous and sustained plant stress, 275f plant defense against, 272–276 insolation, 124, 124f integrated moisture index, 183, 184f integrated pest management (IPM), 610 intermediate crown classes, 309 intermediate disturbance hypothesis, 349 intermittent growth patterns of trees, 99–102 International Union for Conservation of Nature (IUCN), 358 internodes, 99 interspecific competition, 308 intolerant. See understory intolerant intraspecific competition, 308 introgression, 61, 62f introgressive hybridization. See introgression invaders, 227–228 invasive grasses, 600 invasive pathogens, 598 invasive species, 331, 589–591 characteristic traits of invasive plant species, 591–595 growing conditions, 595f impacts of invasive animals on forests, 606–607 impacts of invasive insects and pathogens on forests, 602–606 impacts of invasive plants on forests, 599–602 non‐plant invasive species in forests, 596–599 performance of native vs. invasive plants, 594f primer of invasive species management in forests, 607–612 trees invasive in North America and native to North America, 591t inventory diversity, 332, 333–335 inversion in mountainous terrain, 148 island biogeography theory, 618 isomorphic substitution, 208–209

J

Jaccard similarity coefficients, 335, 339 jack pine, 501, 632, 634, 634f jack pine barrens, 185 juvenile phase, 71–73, 277

K

kaolinite, 208 keystone species, 351–353

bindex.indd 747

knee roots, 110 kudzu, 592, 592f, 600

L

lammas growth, 102 land clearing, 409 landforms, 168, 170, 186, 433. See also physiographic/ physiography aspect of slope, 172–173 elevation of, 170 levels of landscape ecosystems, 434f level terrain, 170–171 parent material in relation to, 173 position in landscape, 174 position on slope, 172 shape of, 170 slope characteristics, 171 slope inclination, 173 sloping terrain, 171 spatial position of, 175 zonation pattern, 187, 188 landscape, 15, 614 agricultural, 647 change, 616 function, 615 occupied, 563 relationships of soil, 215–216 structure, 614–615 landscape ecology, 15, 613 association of mineralization and nitrification rates, 639t carbon balance, 638f concepts, 614–616 disturbances on landscapes, 624–634 forest fragmentation and connectivity, 614–627 housing density, 616f interactions of landscape patterns and ecological processes, 634–640 post‐fire landscape structure characterization, 615f landscape ecosystems, 3, 4–7 approach, 640 and community, 7 distinguishing and mapping at multiple spatial scales, 25–32 diversity of, 336–337, 342, 342f flow of energy and matter, 8f objects of study, 5f spatial scales of, 15–17, 18–20 structure and function, 7–9 temporal scales, 15–17 three‐dimensional, volumetric nature of, 5f types, 9–11, 10f of upper and lower slopes, 6f vegetation types and biomes, 20–25 landscape‐mosaic view, 619

11-14-2022 20:52:31

748  Index landscape pattern effects on disturbance spread, 627 interactions, 634–640 large disturbances in forests, 628f land‐sparing, 649 landtype association (LTA), 263 lapse rate, 170 latent buds, 74 lateral bud primordia, 74 lateral transfers, 635 latewood formation control, 114–115 latitude, 344–346 layering, 39 leaching, 525–526, 526t leaf absorption, 125, 126f leaf and root litter production, 506–508, 507f, 508t leaf area and productivity, 635 leaf area index (LAI), 126, 127–129, 635, 636f leaf C gain light and, 452–454 photosynthesis, dark respiration, and, 450–452 soil nitrogen availability and, 456, 462f temperature and, 454–455, 454f water and, 455, 455f leaf curling, 160 leaf drooping, 160, 161 Leeuwenberg’s model of trees, 98f legacy creation event, 425 legumes, 497 level terrain, 170–171 lifeboating, 361 light, 124, 452–454 and C allocation, 461 canopy structure and leaf area, 125–130 distribution reaching ecosphere, 124 and ecosystem change, 141 energy, 123 and epicormic sprouting, 139 and growth of trees, 133–136 insolation distribution, 124, 124f photocontrol of plant response, 139–141 photosynthetic light saturation curves, 134f photosynthetic response to changing light intensity, 453f plant interception of radiation, 125 quality beneath forest canopy, 130–133, 131f relative irradiance, 125f respiration, 133 saturation point, 134 and seedling survival and growth, 136–137 and tree morphology and anatomy, 137–138 light compensation point, 133 Light Detection and Ranging (LiDAR), 172 light to far‐red light ratio (R/FR ratio), 140 lignin, 517

bindex.indd 748

lignotubers, 226 lipids, 457 live biomass accumulation of plants, 467 livestock influence on forest ecosystems, 288–290 live vegetation removal effect, 231 loblolly pine, 393 local genecological differentiation, 55 local landscape ecosystems, 30. See also regional landscape ecosystems regionalization, 30 in Upper Michigan, 30–32, 31f locally dispersed seeds, 71t local scales, diversity at, 346–351 location‐specific ecosystems, 413 lodgepole pine forests, 579, 580f, 631, 631f logging, 407–408 long shoots of plants, 98 long‐term forest ecosystem and vegetation change, 367. See also forest ecosystems change before Pleistocene age, 367–368 eastern North America, 370–373, 372f independent migration and similarity of communities, 379–380 last glacial maximum, 370–375 Mt. LeConte, Great Smoky Mountains National Park, 374f patterns of tree genera and species migrations, 375–379 Pleistocene glaciations, 368–369 western North America, 373–375 lumbering, 407

M

MacArthur’s broken‐stick model, 335 macroecosystem, 26t macrolevel landscape ecosystems, 17, 20 macronutrients, 206, 206t, 207 biochemical function of, 207t maintenance respiration, 456–458 mammals, 282–284 mammals, plant defense against, 276–277 managed relocation. See assisted species migration mapping, 25 in Baden‐Württemberg, 257–258 ecosystem units at multiple scales, 26 of landscape ecosystems, 237, 246 Massart’s model of trees, 98f mass flow, 499 mast flowering, 76 mast(ing), 76, 280 matric potential of soil water, 204 matrix, 618, 619f mature forest stage, 426 McClure’s model of trees, 98f mean annual precipitation, 473, 473f

11-14-2022 20:52:31

Index  749 mean annual temperature, 473, 473f merging forest communities, 303. See also discrete forest communities Eastern Deciduous Forest, 303–304 New England, 304 topographic distribution of vegetation types, 304f transitional middle‐aged stands, 305t mesarch succession, 420, 421t mesic forests, 88, 563 mesoecosystem, 26t mesolevel landscape ecosystems, 17 mesophication, 590 metamorphic rocks, 195, 197t metapopulation theory, 618 meteorites, 220 methane (CH4), 540 micelles, 200 microbial abundance, 232 microbial activity, 575 microclimate, 147, 149f microecosystem, 26t microlandforms, 191–194 microlevel landscape ecosystems, 17 microrefugia, 378 microtopography, 237 pit‐and‐mound, 191–193, 193f and regeneration in hardwood swamps, 193–194 migration of forest trees, 367 from glacial microrefugia, 377–379 irregularities and disturbance, 377 routes of tree species in eastern North America, 376f sequence and patterns, 371–373 and similarity of communities, 379–380 Millennium Ecosystem Assessment (MEA), 644 mineral cycling, 285–286 minerals in ash, 233 mineral weathering, 488, 493f in ecosystems, 489–490 nutrient inputs, 491 rates, 489, 492t mini‐rhizotron system, 463, 463f mitigation, 581 mixed fire regime, 220 mixed‐severity fires, 220 modern refugia, 584 modular unit of construction, 99 moisture availability, 554 moisture gradients, 303 molecular sieve, 203 monoclimax, 419 monoecious condition, 73 monomorphic gene, 39 montmorillonite, 208, 209 Morista–Horn coefficient, 336 mosaic ofpatch types, 614

bindex.indd 749

mountainous physiography, 176 comparison of physiographic provinces, 178–179f and forests of central Appalachians, 180–183 interfingering of vegetation zones, 180f mountainous terrain of California and Pacific Northwest, 176–180 physiographic diagram of Sierra Nevada Mountains, 177f multiple‐factor methods of site and ecosystem classification, 256 applications of multi‐factor methods in United States and Canada, 258–264 ecosystem classification and mapping in Baden‐ Württemberg, 257–258 landscape ecosystem types, 259f, 260f, 261f, 262f multiple‐scale approaches, 238 multiple spatial scales, landscape ecosystems at, 25–26. See also spatial scales of landscape ecosystems hierarchical ecosystem classification, 26t local landscape ecosystems, 30–32, 31t regional landscape ecosystems, 26–30 mutualisms in forest ecosystems, 307 nonsymbiotic mutualisms, 307–308 symbiotic mutualisms, 307 mycorrhiza, 307, 504–506

N

national classification systems in United States, 262–263 National Hierarchical Framework of Ecological Units (NHFEU), 262 natural disturbances, 381, 385–386, 481 natural hybridization, 61 naturalized species, 590, 593 natural plant distributions and cold hardiness, 162–164 freezing resistance, 163f natural range of variability. See historical range of variability (HRV) natural selection, 38–39, 57 net ecosystem carbon balance (NECB). See net ecosystem productivity (NEP) net ecosystem productivity (NEP), 469, 479–480, 531, 532f, 637 net photosynthesis, 450, 452, 456 net primary productivity (NPP), 232, 467, 487, 574 allocation, 484t estimates of total plant, construction, and maintenance respiration, 468t hypotheses, 483f net secondary production, 467, 469 New England, merging forest communities in, 304 niche, 60 differentiation, 58, 60, 305 spatial component of, 60 temporal component of, 60

11-14-2022 20:52:31

750  Index nitrification, 523–524, 638 nitrogen (N), 456, 487 availability and belowground net primary productivity, 482–485 availability in forest ecosystems, 523 immobilization and mineralization, 521–523 leaf N concentration, 456f leakage, 606 nitrogen‐fixing actinobacteria, 497 removal, 615 nodes, 99 nongenetic variation, 38 non‐indigenous non‐native species, 590 nonlethal understory fires, 220 non‐plant invasive species in forests, 596 cumulative detections of established forest pests over time, 596f forest insects and diseases, 597t geographical variation, 598f tree pathogens, 599 nonsymbiotic mutualisms, 307–308 North American ash, 596, 598 Norway maple, 592–593 Nostoc, 497 nothospecies, 61 novel ecosystems, 610–612 nurse logs, 193 nutrient dynamics, 638–640 in forest floor, 511–513, 512t mineralization, 513–524 nutrient‐use efficiency of litter, 509, 510f retranslocation, 508–511 storage in boreal, temperate, and tropical forests, 528–529, 529t supply of, 209–210 transport to roots, 499–500 uptake and assimilation by roots, 500–503, 501f nutrient additions to forest ecosystems, 487 accumulation of N in forest ecosystems, 498t atmospheric deposition, 492–495 biological fixation of nitrogen, 495–498 flow of nutrients, 488f mineral weathering, 488–492 nutrient cycling, 487, 647 ecosystem C balance and retention and loss of nutrients, 531–532 within forest ecosystems, 499–524 forest harvesting and nutrient loss, 532–535 nutrient additions to forest ecosystems, 487–498 nutrient loss from forest ecosystems, 525–528 and storage of nutrients in forest ecosystems, 528–531 nutrient loss from forest ecosystems, 525, 532f denitrification, 526–528 leaching, 525–526

bindex.indd 750

streamwater nitrate concentrations, 533f transformation of N by soil microorganisms, 527f nutrition, 274–275, 388

O

oaks black oak, 255f northern red oak, 553f, 590 red oak, 295 at risk, 446–447 old‐growth stage of forest trees, 426 old forest stage. See old‐growth stage of forest trees on‐plant seed storage, 227 open system of plant growth, 99 organic matter, 234–236 decomposition, 8, 513–524 heating and burning of, 233 removal by fire, 231, 234–236 of soil, 212–213 organic soils, 146 orographic precipitation, 170 orthotropic models of trees, 97 osmosis, 203 osmotic potential of soil water, 204 overstory biomass, 466, 476, 638 overstory composition, 313–314 overtopped crown classes, 309 overyielding effect, 356 oxisols, 214 oxygen (O), 487

P

paleogeography, 343–344 Pando clone, 92 parent material, 167, 195 consolidated, 195–196, 196f in relation to landform, 173 unconsolidated, 195–197, 196f patch(es), 613–614 in forest ecology, 618–619 patch‐based ecological theory, 618 patch‐based theory, 618 patch–corridor–matrix model, 618 size, 615 pathogens, 570–572 pathway analysis, 609 peatlands, 575 peds. See aggregates of soil people–planet–profit model, 646 percentage similarity (PS), 335, 339 percent base saturation of soil, 210 periodicity of woody plant seed crops, 76–78, 77f permafrost, 398, 575 permanent wilting point, 204 persistent juveniles, 70, 71t perturbations, 382

11-14-2022 20:52:31

Index  751 phases of tree species, 249 phenological/phenology, 550 escape, 552 of flowering, 56 separation, 58 phenotype, 11, 35 green, 57 plasticity of, 36–37 relationship to genotype and environment, 35–36, 36f phenotypic plasticity. See plasticity of phenotype phloem, 111 photocontrol of plant response, 139–141 photoperiodism, 43, 46, 139 photosensory systems of plants, 123 photosynthate, 78 photosynthesis, 123, 151, 450–452, 548–549, 635 photosynthetically active radiation (PAR), 124, 129, 130 photosynthetic capacity of trees, 450, 451–452t phototropism, 97 phyllosilicate clays, 490 phyllosilicate minerals, 200, 208 building blocks of, 208f clays, 201, 208 physical and chemical properties of, 208t physical control for invasive species, 609 physical legacies, 386 physiognomy, 3 physiographic climax, 431 physiographic/physiography, 12, 20, 167–168, 346–349. See also landforms of broadscale ecosystem, 25, 26 characteristics, 168–169 cross section in Upper Peninsula of Michigan, 215f eastern North American map of deciduous forest regions, 22f and firebreaks, 190–191 flatlands, 183–190 landforms, 168, 170–174, 186 landscape development since deglaciation, 169f effect of local topography, 175f microlandforms and microtopography, 191–194 mountainous, 176–183 multiple roles of, 174–176, 176f regions of North America, 21f setting, 169–170 site types, 263 physiological amplitude of species, 305, 306f physiological optimum range of species, 305, 306f phytochrome, 140–141 phytometers, 244 phytosociology in Europe, 298 pignut hickory, 378–379, 379f pines in New England and Lake States, 389–390 pine savannas of western Great Lakes Region, 185

bindex.indd 751

pinyon ips bark beetle, 572 pioneer species. See early‐successional species pit‐and‐mound microtopography, 191–193, 193f plagiotropic models of trees, 97 plant formation, 296 hybrid zones as reservoirs for insect diversity, 275–276 interception of radiation, 125 physiognomy, 23 succession, 417 plant association, 297 at climax, 250 and habitat types in western United States, 247–250 plant defense, 269 hypothesized relationship of plant investment into defense, 271f against insects, 272–276 investment in, 270–272 against mammals, 276–277 relationship of allocation, 272f plant growth, temperature and, 150 cold injury to plants, 153–154 dormancy, 155–156 external influences and internal interactions, 155f frost hardiness and cold resistance, 156–161 growth resumption, 161–162 heat damage in seedlings, 151–152 phases of temperature regime, 152 photosynthesis role, 151 seasonal pattern of cold resistance, 156f stems, roots and leaves, 150–151 thermoperiod, 152–153 winter chilling and growth resumption, 161–162 plant life history, 277–288 plant litter biochemical constituents of, 514–516, 516f and return of nutrients to forest floor and soil, 506–511 plant species addition of, 412 availability and arrival sequence of, 428 elimination, 411–412 groups of ground cover, 245–247 migrations, 375–379 rescue. See assisted species migration plasticity in leaf anatomy, 138 of phenotype, 36–37 Pleistocene glaciations, 368 maximum extent of continental glaciers in North America, 369f temperature data determined from Vostok ice core, 369f point diversity, 332 point processes, 635 pollen analyses, 375

11-14-2022 20:52:32

752  Index pollination, 644 animal, 277 of woody plants, 75–76 polyclimatic climax theory, 432 polyclimax theory, 431–432 polygenic genes, 38 polymerase chain reaction (PCR), 40 polymorphic gene, 39 polymorphic patterns, 241 polyploids, 63–64 polyploidy, 63–65 ponderosa pine forests, 3, 611 average frequencies of two‐needle fascicles on, 53f disturbance and regeneration, 55 eastern geographic races of, 51f genetic differentiation of, 49–52 interval between fires, 222 landscape ecosystems, 262 monoterpene components in xylem resin of, 54f secondary succession following fire in, 439, 440f western geographic races of, 52f populations, 38n1 diversity in response to genetic differentiation, 57 of modular units, 71 post‐disturbance spatial heterogeneity, 648 post‐dormancy, 155 post‐zygotic barriers, 58t, 59 potential evapotranspiration (PET), 474, 545 prairie landscapes, 167 pre‐dormancy, 155 predominant height, 240 Pre‐European colonization forest communities, 186f preforest stage, 425 press disturbances, 382–383 pre‐zygotic barriers, 57–58, 58t primary productivity, 467, 574–575 primary succession, 414, 420–423. See also secondary succession on deglaciated terrain, 434–437 stages, 421t primer of invasive species management in forests, 607 early intervention strategies, 609 management approaches, 609–610, 611t novel ecosystems and invasive species, 610–612 timeline for invasion, 608f proleptic shoots of plants, 102 proportional diversity, 332 provenance test, 41 provisioning services, 644, 645t pulse disturbances, 382

Q

qualitative defense chemicals, 270 quaternary biogeography, 370 quaternary paleoecology, 370

bindex.indd 752

R

races of trees, 42 local races, 55 ponderosa pine, 51f, 52f rain shadow, 170 ramets, 71, 90, 92 rapid foliage decomposition, 226 rapid juvenile growth, 225 rare species, 353–354 Rauh’s model of trees, 98f recruitment process, 70 red maple, 590 red to far‐red ratio (R/FR ratio), 130 reflectance, 125, 126f refugia, 378, 583–585 regionalization, 25, 258 ecological, 26 hierarchical levels of North America, 27 of local landscape ecosystems, 30 regional landscape ecosystems, 26–28. See also local landscape ecosystems ecoregion map for North America, 27f of Michigan, 28–30, 29f regression tree analysis (RTA), 559–560 regulating services, 644, 645t regulatory control for invasive species, 609 reiteration, 97 relay floristics, 422 reorganization. See stand initiation reproductive control for invasive species, 610 reproductive cycles of woody plants, 74–75 reptiles, 278 resilience of forests, 383, 384f, 569, 573, 583 resistance of forests, 383, 384f resisters, 227, 228 resource‐availability hypothesis, 271 resource‐ratio hypothesis, 428 respiration, 123, 133, 548 retranslocation, 508 return interval, 220 rhizomes, 226 rhododendrons, thermotropic movements in, 159–161 richness of species, 331–332 ring‐porous angiosperms, 112, 113f riparian buffers, 615 riparian zones, 187 risk assessment, 609 riverine ecosystems, 342 root architecture, 503–504 root collar or crown, 226 root development of seedlings, 136 root pressure, 115 roots of trees, 102–103 fine root relations, 105–107 horizontal and vertical root development, 107–108

11-14-2022 20:52:32

Index  753 kinds, forms, and occurrence, 103–105 modification of root systems, 104f periodicity of primary root growth, 108 recovering from fire damage, 226 root development of seedlings, 103t root grafting, 108–110, 109f specialized roots and buttresses, 110 temperature of, 150 rule of tens, 590

S

sand dunes, 302 sapwood, 111 saturation water content, 205 scale, 15. See also spatial scales of landscape ecosystems; temporal scales, landscape ecosystems in scatter‐hoarding, 280 scots pine, 49–50, 463, 464f scrub “islands”, 353, 354 seasonality of fire, 220 secondary chemicals, 270, 282 secondary foundation species, 352 secondary growth of trees, 2, 110, 112–115 secondary minerals, 200 secondary succession, 414, 423–427. See also primary succession fire and oak dominance, 446–447 fire in ponderosa pine forests of Western Montana, 439 and gap dynamics, 439–446 stages, 423t stand development stages, 424f sedimentary rocks, 195, 197t seed bank of woody plants, 82–83, 82f seedbeds, 388 seed dispersal, 278 by birds, 279–281 effectiveness, 280 effects of personality on seed dispersal distance, 283f by fish and reptiles, 278 in forest trees, 56–57 by mammals, 282–284 of woody plants, 80–82 seedlings, 554 of gap‐phase and shade‐tolerant species, 88 heat damage in, 151–152 height and top/root weight ratio, 137t root development of, 136 of species, 85 sugar maple, 88, 90 seed production, 226 of woody plants, 76, 77–78 seed‐source test, 41 self‐thinning. See density‐dependent competition

bindex.indd 753

semi‐natural landscapes, 647 senescing, 423 seral stage, 421 sere, 421 series order habitat types, 249 serotinous cones, 81 severity of fire, 220, 221, 222 sexual reproduction of woody plants, 70, 71, 388 correlational studies, 78, 79f distribution of seeds of Engelmann spruce, 80f effects on vegetative growth, 78–80 establishment following, 83–90 flowering, 72, 73, 75 increasing seed production, 73 maturation and ability to flower, 71–72 periodicity of seed crops, 76–78, 77f pollination, 75–76 post‐establishment development, 90 reproductive cycles, 74–75 seed bank, dormancy, and germination, 82–83, 82f seed dispersal, 80–82 sexual systems, 39 shade‐tolerant species, 88, 129, 590 Shannon–Wiener index, 333–334, 629 shape of landforms, 170 short‐interval fires, 569 shortleaf pine, 393 short shoots of plants, 98 short stature, 227 silica tetrahedra structure, 208, 208f silviculture, 13 similar effects, principle of, 187 simulation model, 579, 623 single‐gene mutations, 38 site, 12 climax, 432 districts, 263 index, 239 regions, 263 site‐index curves, 239f, 240–243 site quality and ecosystem, 237 advantages and limitations, 243 comparisons between species, 243 direct measurement of forest productivity, 238–239 environmental factors as measure of site, 251–256 multiple‐factor methods of site and, 256–264 tree height as measure of site, 239–243 vegetation as indicator of, 244–251 slash pine, 393 slope aspect, 172–173 characteristics, 171 inclination, 171, 173 length, 171 position, 172, 172f

11-14-2022 20:52:32

754  Index sloping terrain, 171 snowpack, 546 soil‐moisture availability, 554 soil N availability, 484f and belowground net primary productivity, 482–485 leaf litter and fine‐root production, 485f soil(s), 195, 346–349 acidity, 209, 210–212 cation exchange capacity of, 209–210, 210t chemical properties of, 206, 207 clay mineralogy, 207–209, 210t color, 202 formation, 197 improvement, 285–286 landscape relationships, 215–216 microbes, 231 nitrogen availability, 456, 461–465, 462f non‐wettability before, during, and after fire, 235f organic matter, 212–213 parent material, 195–197 profile development, 197–199 properties, 200, 475–476 structure, 201–202 supply of nutrients, 209–210 surveys, 256 taxonomy, 213–215 temperature, 145–147 texture, 200–201, 201f, 205, 206f, 210t, 476 water, 202–6 soil‐site studies, 252–256 soil‐water potential, 204 stress effects, 116 solar radiation, 124 interception of, 125, 126 plant interception of, 125 Sorensen similarity coefficients, 335 southeastern and southern Coastal Plain, 186 southern Appalachians ecosystem change in, 373 merging forest communities in, 303–304 southern Illinois, 294–296 southern pines, 393, 563 space‐for‐time substitution, 416 spatial heterogeneity, 613, 648 spatial scale of physiography, 169 spatial scales of landscape ecosystems, 15–17. See also multiple spatial scales, landscape ecosystems at climatic classification, 18–19 physiography, 20–22 spatial variability, 639 specialized roots, 110 speciation, 59 species diversity. See also plant species

bindex.indd 754

average number of plant species in local floras, 330f causes of, 343 at continental and subcontinental scales, 343–346 and disturbance, 350f estimates of relative proportion of groups of organisms, 328f increase of biological diversity over geological time, 329f moist temperate forest trees in Northern Hemisphere, 343t patterns of plant species’ richness in Siskiyou Mountains, 347f plant species with increasing elevation in European Alps, 346f predicted number of species along forest succession, 351f species richness of North American trees, 345f threats to, 331 value of biodiversity, 329–331 spodosols, 214, 216 spongy moth, 386, 602, 603, 606 spring vessel formation, 115–116 sprouting, 226 spruce‐fir forests, 563 spur shoots of plants, 98 stability of ecosystem, 383 stable state of ecosystem, 383 stand‐replacing fires, 220, 221, 387, 638 stand‐replacing wildfires in, 629–632 stand density, 312–313 stand habit, 226 stand initiation, 425 starch, 515 steady‐state stage. See old‐growth stage of forest trees stems of trees, 110–111 analysis, 243 earlywood and latewood formation control, 114–115 elongation, 37 exclusion, 426 periodicity and control of secondary growth, 112–114 recovering from fire damage, 226 temperature of, 150 winter freezing and water transport, 115–116 xylem cells and growth rings, 111–112 stratification, 161, 162 stress tolerance hypothesis, 322 stringers, 222 strong sustainability, 646–648 subcontinental scales, diversity at, 343–346 subordinate unions, 247 succession, 349–351, 367, 388–389, 414, 416 autogenic and allogenic, 416 availability and arrival sequence of species, 428

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Index  755 biological legacies, 414 causes, 418t, 427–431 as ecosystem process, 432–434 end point, 431–432 eruption of Mount St. Helens, 437–439, 438f facilitation, tolerance, and inhibition, 428–431 facilitative and inhibitory effects, 430f key characteristics and regeneration strategies, 427–428 mechanisms, 414–416, 427–431 models, 414–416, 422f, 427–431 pathway, 414–416, 415f, 433f primary and secondary, 414 stages, 420–427 work, 418–431 sudden aspen decline, 393 sugar maple, 39, 444–446, 445f canopy, 130 seedlings, 88, 90, 312, 319 stand in Upper Peninsula, 321f sunflecks, 130–133, 132f supercooling, 159 supporting services, 644, 645t surface backfires, 221 surface fires, 220, 221–222 surface headfires, 221 sustainability, 641–642, 646–648 Earth as metaphor for life, 650–651 ecological crisis, 648–650 ecosystem services, 644–645 of forest ecosystems, 641 in forests, 649 historical perspective of sustainability in forests, 643 landscapes, 648f prevalence of human values in forest ecology, 641–642 sustainable ecosystem, 642 sustained yield, 643 sustaining natural resources, 643 sylleptic shoots of plants, 102 symbiotic mutualisms, 307 sympatric speciation, 57

T

taxonomy of soil, 213–215 temperate forest ecosystem, nitrogen and calcium cycle of, 529–531, 530f temperature, 12, 143, 454–455, 454f, 540 climate effects on, 542 deciduousness and, 164 within forest, 147–148 geographical patterns of, 143–145 mean weekly maxima, minima, and mean temperatures, 147t

bindex.indd 755

natural plant distributions and cold hardiness, 162–164 observed changes in annual average temperature, 543f percentage of land area of contiguous United States, 544f and plant growth, 150–162 at soil surface, 145–147 sum, 46, 46n2, 144 variation with local topography, 148–150 temporal scales, landscape ecosystems in, 15–17 terrain age, 434 thermal conductivity, 146 thermoperiod, 152–153 thermotropic movements in rhododendrons, 159–161 thinning, 388 till plains of Midwest, 185 tolerance mechanism, 428–431, 429f. See also understory tolerance of species Tomlinson’s model of trees, 98f top‐down approach, 432 top height concept, 240 topography, 216. See also microtopography temperature variation with local topography, 148–150 varying leaf phenology related to, 552f tornadoes, 401, 628 total soil‐water potential, 204 toxins, 270 tracheids, 112 transmittance, 125, 126f tree‐fall gaps, 442, 443f tree form, 95 architectural models, 97–102 climate role, 96 decurrent form, 96f, 97 excurrent form, 95–96, 96f woody plant, 95 tree‐line position around globe, 144, 145f tree(s) architecture, 95 genera patterns, 375–379 light and growth of, 133–136 light and tree morphology and anatomy, 137–138 line, 302 pathogens, 599 primary growth, 2 regeneration, 554 roots, 102–110 stems, 110–116 uprooting, 191–193 water deficits and tree growth, 116–117

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756  Index trees, carbon balance of, 450 allocation to structure, storage, and defense, 459–461, 460f construction and maintenance respiration, 456–458 light and C allocation, 461 light and leaf C gain, 452–454 photosynthesis, dark respiration, and leaf C gain, 450–452 soil nitrogen availability and C allocation, 461–465 soil nitrogen availability and leaf C gain, 456 temperature and leaf C gain, 454–455, 454f water and leaf C gain, 455, 455f tree species distributions, climate change effects on, 556 climate envelope richness, 561f future projections of distribution of sugar maple, 559f habitat for current forest types of eastern United States, 564f modeled changes in distance and direction of mean centers, 562f observed range shifts, 556–558 projected changes in, 558–560 tree biomass and seedling densities, 557f tree‐species line, 302 trembling aspen, 65, 390–393 clones, 91, 92f, 92 ecosystem, 127f juvenile sprouts of, 277 triple bottom line, 646, 646f Troll’s model of trees, 98f trophic systems, 269 true dormancy, 155

V

U

water, 202, 455, 455f physical properties of, 203 purification, 647 relations and dynamics of, 12 transport, 115–116 vapor, 540 watershed protection, 649 wave‐regenerated fir species, 404 weak sustainability, 646–647 weather(ing), 195, 208, 209, 539 western pines, 390–393 wet deposition, 492, 495 white‐red‐jack pine forests, 563 white‐rot fungi, 517 whitebark pine, 583 whole genome duplication, 64 widely wind‐dispersed seeds, 71t wilderness areas, 647, 650 wildfires, large, 628 wildland–urban interface (WUI), 647

ultisols, 214 umbrella species, 351 unconsolidated parent material, 195–197, 196f understory intolerant species, 315, 316–317 understory reinitiation, 426 understory tolerance of species, 315, 316–317 competition in, 314–315 environmental factors relating to, 322–323 examples in forest ecosystems, 319 nature, 319–324 photosynthesis and respiration for tolerant and intolerant species, 323t physiological processes, 323–324 ratings of tree species, 317–319 of selected North American forest trees, 318t unicellular bacteria, 497 United States National Vegetation Classification (USNVC), 23, 25 upgrading. See aggrading

bindex.indd 756

variable fire regime, 220 variable‐retention harvest system, 360–363, 360f vascular plant diversity, 350–351 vegetation, 299, 413, 417 applications and limitations of vegetation, 250–251 change or succession, 16 distribution of forest trees, 249f ecological species groups and ecosystem types, 248t height‐over‐age curves of Scots pine, 245f as indicator of site quality, 244 operational site classification, 247–250 plant associations and habitat types, 247–250 reduction of competing, 388 species groups of ground cover, 245–247 types, 20–25 units, 16 vegetative reproduction of woody plants, 71t conceptual model of, 91f by conifers, 92, 93 sprouting, 90–91, 94 survival of single genet, 93–94 trembling aspen clones, 91–92, 92f vine maple clone, 92, 93f vermiculite, 208 vertical sinker roots, 103 vessels, 112 vital attributes concept, 428 volatilization, 233 volcanic eruptions, 628

W

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Index  757 wildlife habitat, 389 Douglas‐fir in Pacific Northwest, 394, 395f fire and ponderosa pine, 392f giant sequoia, 394–395 lodgepole pine regeneration, 390f natural longleaf pine seedlings, 394f pines in New England and Lake States, 389–390 southern pines, 393 western pines and trembling aspen, 390–393 wind, 400, 572–573 broadscale disturbance by hurricanes, 403–404 principles of wind damage, 401–403 widespread and local effects, 401 wind‐pollinated species, 75 windthrow of standing trees, 103, 402f, 403 winter chilling, 161–162 freezing, 115–116 temperatures, 572

bindex.indd 757

woody plant, 69, 275 asexual reproduction, 70, 71 plant life cycle with trees, 70f populations, 62 regenerative strategies, 71t sexual regeneration, 72f sexual reproduction, 70–90 vegetative reproduction, 90–94

X

xerarch succession, 420, 421t xylem cells, 110 small‐diameter, 115 of stem, 111–112

Y

Yellowstone National Park, 628–629 fire, carbon, and climate change in forests of, 579–581

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