Vegetation ecology of Central Europe. Volume II, Ecology of Central European non-forest vegetation : coastal to alpine, natural to man-made habitats 978-3-319-43048-5, 3319430483, 978-3-319-43046-1

Table of contents :
Front Matter ....Pages i-xxxiv
Front Matter ....Pages 1-1
Salt Marshes and Inland Saline Habitats (Christoph Leuschner, Heinz Ellenberg)....Pages 3-61
Sand Dunes and Their Vegetation Series (Christoph Leuschner, Heinz Ellenberg)....Pages 63-115
Mires (Christoph Leuschner, Heinz Ellenberg)....Pages 117-187
Vegetation of Freshwater Habitats (Christoph Leuschner, Heinz Ellenberg)....Pages 189-269
Vegetation of the Alpine and Nival Belts (Christoph Leuschner, Heinz Ellenberg)....Pages 271-431
Front Matter ....Pages 433-433
Dwarf Shrub Heaths and Nardus Grasslands (Christoph Leuschner, Heinz Ellenberg)....Pages 435-494
Nutrient-Poor Dry Grasslands (Christoph Leuschner, Heinz Ellenberg)....Pages 495-596
Agricultural Grassland on Mesic to Wet Soils (Christoph Leuschner, Heinz Ellenberg)....Pages 597-731
Communities on Heavy Metal-Rich Soils (Christoph Leuschner, Heinz Ellenberg)....Pages 733-749
Banks, Shorelines and Muddy Habitats Influenced by Man (Christoph Leuschner, Heinz Ellenberg)....Pages 751-764
Ruderal Communities on Drier Soils (Christoph Leuschner, Heinz Ellenberg)....Pages 765-778
Vegetation of Arable Fields, Gardens and Vineyards (Christoph Leuschner, Heinz Ellenberg)....Pages 779-839
Vegetation of Human Settlements (Christoph Leuschner, Heinz Ellenberg)....Pages 841-860
Syntaxonomic Overview of the Vascular Plant Communities of Central Europe: Non-Forest Formations (Christoph Leuschner, Heinz Ellenberg)....Pages 861-872
Back Matter ....Pages 873-1093

Citation preview

Christoph Leuschner Heinz Ellenberg

Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats Vegetation Ecology of Central Europe Volume II

Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats

Christoph Leuschner  •  Heinz Ellenberg

Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats Vegetation Ecology of Central Europe, Volume II Revised and Extended Version of the 6th German Edition Translated by Laura Sutcliffe

Christoph Leuschner Plant Ecology University of Göttingen Göttingen, Germany

Heinz Ellenberg (deceased) University of Göttingen Göttingen, Germany

Translation of the revised and extended German language edition: Vegetation Mitteleuropas mit den Alpen, by Heinz Ellenberg/Christoph Leuschner, © 2010 by Eugen Ulmer KG, Stuttgart, Germany. All Rights Reserved. ISBN 978-3-319-43046-1    ISBN 978-3-319-43048-5 (eBook) DOI 10.1007/978-3-319-43048-5 Library of Congress Control Number: 2017943125 © Springer International Publishing Switzerland 2017 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. Cover illustration: Forb-rich, extensively used false oat-grass meadow on the Baltic Sea island of Hiddensee Printed on acid-free paper This Springer imprint is published by Springer Nature The registered company is Springer International Publishing AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

In the memory of Heinz Ellenberg (1913–1997)

Preface

With the publication in 1963 of the first edition of his classical textbook Vegetation Mitteleuropas mit den Alpen (Vegetation Ecology of Central Europe), Heinz Ellenberg produced a unique compendium of the diverse vegetation types of Central Europe and their ecology. This region in the heart of Europe with more than 7000 years of continuous human settlement may well, together with Britain, be the best studied region on earth with respect to its vegetation ecology, visible in a myriad of publications on the floristics and plant communities, and the ecology of species, communities and ecosystems. Covering an area of approximately one million km2 in the northern temperate zone, Central Europe harbours a rich variety of landscapes from the Baltic and North Sea coasts in the north to the Alps in the south and from eastern France to eastern Poland. This diverse and well-studied region can provide profound insights into the complex interactions between environment, vegetation and man. Ellenberg’s concept was so successful because he based his analysis on a thorough description and sound classification of the plant communities, as many ecological statements lose value if they cannot be related to a particular community or vegetation type. It is on this foundation that causal relationships between environment, species composition, community dynamics and ecosystem functioning are explored for the main vegetation types. This truly interdisciplinary approach also includes the historical dimension of the vegetation, its recent change under the impact of human land use pressure and climate change, and current conservation issues. In the more than 50 years since the book’s first appearance, a tremendous amount of relevant research has been carried out, so that the task of providing a comprehensive overview of the plant, vegetation and ecosystem ecology of the Central European landscape mosaic has become even more challenging. However, the need for an interdisciplinary synthesis of facts and concepts in plant, vegetation and ecosystem ecology is also increasing. Most current environmental problems such as climate change, land use intensification, eutrophication or acid rain involve multifactorial causation, yet scientific training and research are increasingly narrowed into specialist fields. A broad synthesis of the existing ecological knowledge is vii

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therefore vital to promote sound understanding of the ecosystem as a whole. This was the motivation for me to take on the task of producing an English version of the book using the 6th German edition as a basis (Ellenberg/Leuschner, 2010: Vegetation Mitteleuropas mit den Alpen, Ulmer Verlag). For the English book, most of the text was revised, a wealth of new facts added and more than 400 recent references incorporated (now over 5500 cited publications in total). For convenience, the book was split into two volumes (Vol. I, Ecology of Central European Forests; Vol. II, Ecology of Central European Non-Forest Vegetation), whereby Volume I also contains a general introduction into the physical geography, phytogeography and land use history of Central Europe. The book has greatly profited from the careful work and editing of Laura Sutcliffe who, as a native English speaker trained in ecology in Germany, had the task of translating the text. Many of the frameworks, classifications and other terminologies in the German-speaking ecological literature have developed independently of its Anglo-Saxon counterpart, and we spent many hours discussing how best to translate these. During the process of writing, many colleagues, fellow workers, students and technical assistants have supported me by supplying information, pointing at mistakes, correcting text passages, producing art work and helping to compile the references lists, especially Markus Hauck, Werner Härdtle, Dietrich Hertel, Michael Runge, Norbert Hölzel, Irmgard Blindow, Helge Walentowski, Yasmin Abou Rajab, Bernhard Schuldt, Jonas Glatthorn, Stefan Kaufmann and Ina C. Meier. I am very grateful to Bernd Raufeisen who produced all the figures and Astrid Röben who helped with the reference list. I would like to thank Valeria Rinaudo and Ineke Ravesloot from the Springer Editorial group and the production team around Prasad Gurunadham and S. Madhuriba for the dedication and support they gave this project. I hope that the two volumes of this book provide the reader with a useful and thought-provoking synthesis of the dynamics and functioning of Central European ecosystems with its characteristic vegetation types, habitats and landscapes. Clearly, such a book represents a subjective selection of topics and can cover only a fraction of the relevant publications, which is a severe shortcoming, but I hope that it serves to direct the reader to the relevant further reading. I am always grateful for corrections and any supplementary information. I hope above all that this book will continue to inspire current and future vegetation ecologists and provide a solid information platform for action to protect and value these landscapes. Göttingen, Germany March 2016

Christoph Leuschner

Contents of Volume II

Part I  Natural or Near-Natural Formations 1 Salt Marshes and Inland Saline Habitats.............................................. 3 1.1 The Halophyte Flora of Central Europe.......................................... 3 1.2 Environmental Conditions and Habitat Classification.................... 5 1.2.1 The North Sea Intertidal Mudflats.................................... 5 1.2.2 The Baltic Coast................................................................ 8 1.3 Vegetation........................................................................................ 9 1.3.1 Classification of the North Sea Coastal Vegetation........... 9 1.3.2 The Baltic Sea Coastal Vegetation.................................... 21 1.3.3 Halophyte Vegetation in Inland Saline Habitats............... 31 1.4 Adaptations to the Environment...................................................... 34 1.4.1 Adaptations to  Salinity...................................................... 34 1.4.2 Adaptations to Flooding, Anoxia and Sedimentation....... 39 1.4.3 Adaptations to Low Nutrient Levels and Drought Stress............................................................ 42 1.5 Productivity and Nutrient Cycling.................................................. 43 1.6 Vegetation Dynamics...................................................................... 46 1.6.1 The Genesis of Salt Marshes............................................. 46 1.6.2 Zostera Community Dynamics......................................... 47 1.6.3 Succession in Salt Marsh Communities Caused by Sedimentation............................................................... 49 1.6.4 Succession on Bare Sand Banks....................................... 49 1.6.5 The Colonisation of a Muddy Island in the Baltic Sea................................................................ 52 1.7 Human Influence............................................................................. 53 1.7.1 Effects of Salt Marsh Grazing........................................... 53 1.7.2 Salt Marsh Succession After Abandonment of Grazing......................................................................... 56 1.7.3 Eutrophication of Coastal Waters...................................... 57

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2 Sand Dunes and Their Vegetation Series............................................... 63 2.1 Flora and  Vegetation....................................................................... 63 2.2 Dune Formation and Destruction.................................................... 63 2.2.1 Coastal Dunes of the North Sea........................................ 64 2.2.2 Coastal Dunes of the Baltic Sea........................................ 71 2.3 Vegetation........................................................................................ 72 2.3.1 Coastal Dunes of the North Sea........................................ 72 2.3.2 Baltic Coastal Dunes......................................................... 85 2.3.3 Mobile Dunes Without Vegetation.................................... 90 2.3.4 Vegetation of Inland Dunes............................................... 92 2.4 Adaptations to the Environment...................................................... 98 2.4.1 Life at the Drift Line......................................................... 98 2.4.2 Adaptations to Coverage with Sand, Drought and Heat Stress.................................................................. 100 2.4.3 Adaptations to Low Nutrient Levels................................. 104 2.4.4 Dwarf Plants and Phreatophytes in Dune Slacks.............. 107 2.5 Vegetation Dynamics...................................................................... 108 2.5.1 Nutrient Accumulation and Soil Genesis During Dune Formation.................................................... 110 2.5.2 Changes to the Vegetation Caused by Invasive Species................................................................ 112 2.6 Human Influence............................................................................. 113 2.6.1 The Effects of Eutrophication and Drainage..................... 113 2.6.2 The Effects of Livestock Grazing, Rabbit Browsing and Afforestation............................................................... 114 3 Mires.......................................................................................................... 117 3.1 Flora................................................................................................ 118 3.2 Environmental Conditions and Habitat Classification.................... 120 3.2.1 Peat Formation and Decomposition.................................. 121 3.2.2 Surface Structure and Morphological Classification of Mires...................................................... 122 3.2.3 Macroclimate and Mire Formation................................... 129 3.2.4 Microclimate..................................................................... 131 3.2.5 Water Regimes and Hydrological Mire Types.................. 132 3.2.6 Nutrient Supply and Trophic Mire Types.......................... 137 3.3 Vegetation........................................................................................ 144 3.3.1 Synsystematic Overview................................................... 144 3.3.2 Communities of Raised Bog Hummocks and Wet Heaths (Oxycocco-Sphagnetea)......................... 145 3.3.3 Communities of Hollows in Oligotrophic Mires.............. 151 3.3.4 Acid to Basic Mesotrophic Mires..................................... 153 3.3.5 Calcareous Mesotrophic Fens........................................... 156

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3.4 Adaptations to the Environment...................................................... 159 3.4.1 Adaptations to Drought, Flooding and Anoxic Conditions...................................................... 159 3.4.2 Adaptations to Low Nutrient Levels and the Role of Base Richness.......................................... 162 3.5 Productivity and Cycling of Water and Nutrients........................... 165 3.5.1 Productivity....................................................................... 165 3.5.2 Peat Accumulation and Decomposition............................ 168 3.5.3 Water and Nutrient Cycling.............................................. 169 3.6 Vegetation Dynamics...................................................................... 172 3.6.1 Quaternary Mire Development.......................................... 172 3.6.2 Recent Developmental Processes...................................... 176 3.6.3 Primary Succession in Growing Raised Bogs................... 178 3.6.4 Secondary Succession After Mire Drainage..................... 179 3.7 Human Influence............................................................................. 179 3.7.1 Exploitation by Peat Cutting, Drainage and Cultivation.................................................................. 179 3.7.2 Eutrophication................................................................... 183 3.7.3 Accumulation of Pollutants and Exchange of Trace Gases with the Atmosphere................................ 184 3.7.4 Conservation and Restoration of Mires............................. 185

4 Vegetation of Freshwater Habitats......................................................... 189 4.1 Freshwater Macrophytes and Their Origins.................................... 189 4.2 Environmental Conditions and Habitat Classification.................... 190 4.2.1 Physical Characteristics.................................................... 191 4.2.2 Chemical Characteristics................................................... 194 4.2.3 Ecological Classification of Freshwater Systems............. 198 4.3 Vegetation........................................................................................ 204 4.3.1 The Classification of Aquatic Plant Communities............ 204 4.3.2 Still Water Bodies............................................................. 206 4.3.3 Streams and  Rivers............................................................ 224 4.3.4 Communities of  Springs.................................................... 232 4.4 Adaptations to the Environment...................................................... 235 4.4.1 Photosynthesis in Aquatic Plants...................................... 235 4.4.2 Nutrient Uptake by Aquatic Plants................................... 237 4.4.3 Survival in Hypoxic Sediments......................................... 238 4.4.4 Adaptations to Flowing Water and Wave Action.............. 239 4.4.5 Life Forms and Morphological Adaptations of Aquatic Plants............................................................... 241 4.5 Population Biology and Community Ecology................................ 243 4.5.1 Phenology.......................................................................... 243 4.5.2 Life Cycles of Aquatic Plants........................................... 243 4.5.3 Interspecific Competition Between Aquatic Plant Species..................................................................... 247

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4.6 Productivity and Cycling of Water and Nutrients........................... 249 4.6.1 Productivity....................................................................... 249 4.6.2 Water Cycling.................................................................... 252 4.6.3 Nutrient Cycling................................................................ 253 4.7 Vegetation Dynamics...................................................................... 254 4.7.1 Seasonal and Interannual Fluctuations.............................. 254 4.7.2 Long-Term Dynamics and Succession in Lakes............... 256 4.7.3 Self-Purification of Water Bodies as Secondary Succession......................................................................... 257 4.8 Human Influence............................................................................. 259 4.8.1 Eutrophication of Water Bodies........................................ 259 4.8.2 Acidification...................................................................... 264 4.8.3 Reedbed Dieback.............................................................. 266 4.8.4 Threats to and Conservation of Freshwater Habitats........ 267

5 Vegetation of the Alpine and Nival Belts................................................ 271 5.1 Flora and  Development................................................................... 271 5.2 Environmental Conditions and Habitat Classification.................... 276 5.2.1 The High Mountain Climate............................................. 276 5.2.2 Soils and Nutrient Supply................................................. 291 5.2.3 Soil Moisture Regime....................................................... 294 5.3 Vegetation........................................................................................ 296 5.3.1 Vegetation Zonation in the High Mountains..................... 296 5.3.2 Vegetation Mosaics in the Subalpine Belt and at the Tree Line........................................................... 298 5.3.3 Ecological and Synsystematic Classifications of Alpine Vegetation......................................................... 299 5.3.4 Vegetation Mosaics in the Nival Belt................................ 301 5.3.5 Subalpine-Alpine Grasslands on Carbonate Bedrock (Class Seslerietea albicantis)........................... 304 5.3.6 Carici rupestris-Kobresietea bellardii (Wind-Edge Naked Rush Swards).................................... 314 5.3.7 Subalpine-Alpine Grasslands on Acid Soil (Class Caricetea curvulae).............................................. 316 5.3.8 Subalpine-Alpine Dwarf Shrub Heaths (Class Loiseleurio-Vaccinietea)....................................... 323 5.3.9 Class Salicetea herbaceae and Alliance Arabidion caeruleae......................................................... 328 5.3.10 Subalpine-Alpine Mires, Springs and Flooded Banks................................................................................. 335 5.3.11 Subalpine-Alpine Tall-Herb Communities and  Green Alder Scrub (Class Betulo-Adenostyletea and Betulo-Alnetea viridis)............................................ 344

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5.3.12 The Class Thlaspietea rotundifolii of Carbonate Scree Slopes................................................ 351 5.3.13 The Vegetation of Rocky Areas (Class Asplenietea trichomanis and Others).................... 358 5.3.14 Plant Communities of the Nival Belt................................ 368 5.4 Adaptations to the Environment...................................................... 374 5.4.1 Adaptations to High and Low Temperatures..................... 374 5.4.2 Carbon Gain and Turnover................................................ 379 5.4.3 Nutrient Acquisition by Alpine Plants and Adaptations to Basic and Acidic Soils....................... 385 5.4.4 Characteristic Life Forms in the Alpine and Nival Belts.................................................................. 391 5.4.5 The Structure and Causes of the Alpine Tree Line........... 394 5.5 Population Biology and Community Ecology................................ 403 5.5.1 The Growth and Developmental Rhythms of Alpine Plants................................................................. 403 5.5.2 Diaspore Dispersal and Seedling Establishment............... 405 5.6 Productivity and Cycling of Water and Nutrients........................... 406 5.6.1 Productivity....................................................................... 406 5.6.2 Water and Nutrient Cycling.............................................. 408 5.7 Vegetation Dynamics...................................................................... 411 5.7.1 Primary Succession on Glacier Forelands......................... 411 5.7.2 Changes in the Vegetation as a Result of Climate Change............................................................ 423 5.8 Human Influence............................................................................. 427 5.8.1 Changes in Land Use and Eutrophication at High Elevations............................................................. 427 5.8.2 The Threat to Alpine Vegetation Posed by Tourism and Restoration Approaches............................................. 428

Part II  Partly or Mostly Anthropogenic Formations 6 Dwarf Shrub Heaths and Nardus Grasslands....................................... 435 6.1 Flora and  Development................................................................... 435 6.2 Environmental Conditions and Habitat Classification.................... 440 6.2.1 Climatic Conditions and Water Regime............................ 441 6.2.2 Stand Structure and Microclimate..................................... 442 6.2.3 Soils and Nutrient Supply................................................. 442 6.3 Vegetation........................................................................................ 446 6.3.1 Nardus Grasslands (Order Nardetalia strictae).............. 446 6.3.2 Dwarf Shrub Heaths (Order Vaccinio-Genistetalia)........ 452 6.4 Adaptations to the Environment...................................................... 470 6.4.1 Growth Form and Light Demand...................................... 470 6.4.2 Adaptations to Low Nutrient Availability and Unfavourable Soil Chemical Conditions.................... 471 6.4.3 Adaptations to Drought and Frost..................................... 474

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6.5 Population Biology and Community Ecology................................ 474 6.6 Productivity and Water and Nutrient Cycling................................. 475 6.6.1 Productivity....................................................................... 475 6.6.2 Water and Nutrient Cycling.............................................. 478 6.7 Vegetation Dynamics...................................................................... 480 6.7.1 Fluctuations and ‘Cyclical Succession’............................ 480 6.7.2 Succession Following Disturbance................................... 481 6.7.3 Succession to  Forest.......................................................... 482 6.8 Human Influence............................................................................. 487 6.8.1 Grazing and  Mowing......................................................... 487 6.8.2 Eutrophication and  Acidification...................................... 488 6.8.3 Conservation and Restoration of Heathland and Nardus Grasslands...................................................... 491

7 Nutrient-Poor Dry Grasslands................................................................ 495 7.1 Flora and  Development................................................................... 495 7.1.1 Flora.................................................................................. 495 7.1.2 The Role of the Climate and Humans in the Development of Dry Grasslands............................. 497 7.2 Environmental Conditions and Habitat Classification.................... 499 7.2.1 Aspect and  Microclimate.................................................. 500 7.2.2 Soil Moisture Regime....................................................... 505 7.2.3 Soil Types and Their Nutrient Supply............................... 512 7.3 Vegetation........................................................................................ 514 7.3.1 Classification of the Major Habitat Types......................... 514 7.3.2 Calcareous Dry Grasslands (Class Festuco-Brometea)................................................ 522 7.3.3 Sandy Dry Grasslands and Vegetation of Cliffs and Rocky Debris (Class Koelerio-Corynephoretea)............................................... 538 7.4 Adaptations to the Environment...................................................... 545 7.4.1 Adaptations to  Drought..................................................... 545 7.4.2 Adaptations to Nutrient Shortage...................................... 561 7.4.3 Adaptation to Heat Stress.................................................. 565 7.4.4 Adaptations to Basic and Acidic Soils.............................. 565 7.5 Population Biology and Community Ecology................................ 566 7.5.1 Phenology.......................................................................... 566 7.5.2 Seed Bank, Germination and Dispersal............................ 567 7.5.3 The Influence of Competition on the Species Composition...................................................................... 570 7.5.4 The Causes of Species Richness in Dry Grasslands......... 573 7.6 Productivity and Cycling of Water and Nutrients........................... 576 7.6.1 Productivity....................................................................... 576 7.6.2 Water and Nutrient Cycling.............................................. 578

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7.7 Vegetation Dynamics...................................................................... 581 7.7.1 Primary Succession on Rock Debris and Opencast Mine Areas................................................. 581 7.7.2 Short- and Mid-term Changes in Dry Grasslands............. 582 7.7.3 Secondary Succession in Abandoned Dry Grasslands...... 584 7.7.4 The Development of Steppe-Like Grasslands on Abandoned Fields........................................................ 586 7.8 Human Influence............................................................................. 587 7.8.1 Mown and Grazed Dry Grasslands................................... 587 7.8.2 Eutrophication................................................................... 591 7.8.3 Habitat Fragmentation....................................................... 592 7.8.4 Conservation and Restoration of Dry Grasslands............. 593

8 Agricultural Grassland on Mesic to Wet Soils...................................... 597 8.1 Flora and  Development................................................................... 598 8.1.1 Flora.................................................................................. 598 8.1.2 The Creation of Meadows and Pastures............................ 600 8.2 Environmental Conditions and Habitat Classification.................... 602 8.2.1 Grazing and Mowing as Key Site Factors......................... 602 8.2.2 Stand Structure and Microclimate..................................... 608 8.2.3 Soil Moisture Regime....................................................... 610 8.2.4 Soil Chemical Properties................................................... 619 8.3 The Vegetation of Agricultural Grasslands and Roadside Verges....................................................................... 622 8.3.1 Overview of the Agricultural Grassland Communities of Central Europe....................................... 622 8.3.2 Mesic Meadows................................................................ 625 8.3.3 Wet Meadows.................................................................... 644 8.3.4 Filipendula Riverbank and Similar Tall Forb Communities..................................................................... 656 8.3.5 Grass Verges and Traditional Orchards............................. 657 8.3.6 Pastures and Frequently Mown Grasslands...................... 661 8.3.7 Vegetation of Trampled Ground and Flooded Grassland........................................................................... 666 8.4 Adaptations to the Environment...................................................... 671 8.4.1 Tolerance of Mowing and Trampling................................ 671 8.4.2 Some Ecophysiological Properties of Grassland Species......................................................... 678 8.5 Population Biology and Community Ecology................................ 688 8.5.1 Phenology.......................................................................... 688 8.5.2 Seed Bank, Germination and Dispersal............................ 690 8.6 Productivity and Cycling of Water and Nutrients........................... 693 8.6.1 Productivity....................................................................... 693 8.6.2 Water and Nutrient Cycling.............................................. 701

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8.7 Vegetation Dynamics...................................................................... 706 8.7.1 The Establishment of Meadow Communities and the Importance of Stand History................................ 706 8.7.2 Succession in Abandoned Meadows................................. 707 8.8 Human Influence............................................................................. 709 8.8.1 The Effects of Management Intensification and Fertilisation................................................................. 709 8.8.2 The Effects of Drainage and Irrigation............................. 720 8.8.3 Conservation and Restoration of Meadows...................... 727

9 Communities on Heavy Metal-Rich Soils.............................................. 733 9.1 The Origins and Development of the Heavy Metal Flora............... 734 9.2 Heavy Metal Soils........................................................................... 735 9.3 Vegetation........................................................................................ 738 9.3.1 Vascular Plant Communities............................................. 738 9.3.2 Lichen Vegetation.............................................................. 739 9.4 Adaptations to the Environment...................................................... 740 9.4.1 Heavy Metal Soils with Low Magnesium Concentrations.................................................................. 740 9.4.2 Heavy Metal Soils with High Magnesium Concentrations.................................................................. 745 9.5 Vegetation Dynamics...................................................................... 746 9.6 The Effects of Heavy Metal Deposition on the Vegetation............. 746 10 Banks, Shorelines and Muddy Habitats Influenced by Man............... 751 10.1 Short-Lived Isoëto-Nanojuncetea Communities on Periodically Wet Soils................................................................ 751 10.1.1 Range and Dispersal of the Nanocyperetalia................... 751 10.1.2 Vegetation.......................................................................... 753 10.1.3 Environmental Adaptations and Vegetation Dynamics.......................................................................... 756 10.2 Nitrophilic Bidentetea Communities of Still and Flowing Water Bodies.............................................................. 757 10.2.1 Environmental Conditions and Habitat Classification..................................................................... 757 10.2.2 Vegetation.......................................................................... 760 10.2.3 Environmental Adaptations and Vegetation Dynamics.......................................................................... 762 11 Ruderal Communities on Drier Soils..................................................... 765 11.1 Flora and  Development................................................................... 765 11.2 Environmental Conditions and Habitat Classification.................... 768 11.3 Vegetation........................................................................................ 769 11.3.1 Ruderal Communities of Summer and Winter Annuals........................................................... 769 11.3.2 Communities of Perennial Ruderal Plants........................ 773 11.4 Adaptations to the Environment...................................................... 777

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12 Vegetation of Arable Fields, Gardens and Vineyards........................... 779 12.1 Flora and  Development................................................................... 779 12.1.1 Flora.................................................................................. 779 12.1.2 The Origins of Arable Weeds and Changes in the Segetal Vegetation Since the Neolithic................... 780 12.2 Environmental Conditions and Habitat Classification.................... 783 12.2.1 Microclimate..................................................................... 783 12.2.2 Soil Moisture Regime....................................................... 784 12.2.3 Soil Acidity and Nutrient Supply...................................... 784 12.3 Vegetation........................................................................................ 786 12.3.1 Classification of Arable Communities.............................. 786 12.3.2 Synsystematic Overview................................................... 788 12.4 Adaptations to the Environment...................................................... 794 12.4.1 Arable Plant Functional Types.......................................... 794 12.4.2 Germination Conditions and Light and Temperature Requirements.................................................................... 798 12.4.3 Adaptations to Moisture and Aeration of the Soil............ 800 12.4.4 Adaptations to the Nutrient Regime and the Effects of Fertiliser Application.................................................... 803 12.4.5 The Effects of Cultivation................................................. 809 12.4.6 The Effects of Herbicide Application............................... 811 12.5 Population Biology and Community Ecology................................ 813 12.5.1 Phenology.......................................................................... 813 12.5.2 Diaspore Banks and Dispersal.......................................... 813 12.5.3 Population Dynamics........................................................ 818 12.6 Productivity and Cycling of Water and Nutrients........................... 819 12.6.1 Water Cycling.................................................................... 819 12.6.2 Nutrient Cycling................................................................ 820 12.7 Vegetation Dynamics...................................................................... 822 12.7.1 Interannual Fluctuations and Changes with Crop Rotation............................................................ 822 12.7.2 Secondary Succession on Arable Fallows......................... 823 12.8 Human Influence............................................................................. 829 12.8.1 The Recent Collapse of Arable Weed Populations and Its Causes................................................ 829 12.8.2 Conservation and Restoration of Arable Weed Vegetation................................................................ 836 13 Vegetation of Human Settlements.......................................................... 841 13.1 The Flora of Towns and Villages and Its Origins............................ 841 13.2 Environmental Conditions and Habitat Classification.................... 846 13.3 Vegetation........................................................................................ 847 13.4 Adaptations to the Environment...................................................... 855 13.5 Vegetation Dynamics...................................................................... 858

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14 Syntaxonomic Overview of the Vascular Plant Communities of Central Europe: Non-Forest Formations.......................................... 861 References......................................................................................................... 873 Index.................................................................................................................. 1013

Contents of Volume I

Part I  The Natural Environment and Its History 1 Environmental and Historical Influences on the Vegetation of Central Europe.................................................................................... 3 1.1 The Climate and Phytogeography of Central Europe..................... 3 1.2 An Overview of the Geology and Soils of Central Europe............. 10 1.3 Historical Influences on the Vegetation of Central Europe............. 13 2 Life Forms and Growth Types of Central European Plant Species............................................................................................. 23 2.1 Life Forms....................................................................................... 23 2.2 Endogenous Rhythms..................................................................... 26 2.3 Plant Anatomy and Morphology..................................................... 27 Part II  The Role of Man 3 The Central European Vegetation as the Result of Millennia of Human Activity.............................................................. 31 3.1 Phases of Forest Clearance.............................................................. 31 3.2 The Effects on the Vegetation of Low-Intensity Grazing and Woodland Use............................................................ 40 3.2.1 The Opening Up and Destruction of the Forest................ 40 3.2.2 The Spread of Pasture Weeds............................................ 45 3.2.3 Soil Degradation Through Low-Intensity Grazing........... 47 3.3 From Coppiced Woodlands to Modern Forestry............................. 52 3.3.1 Coppicing With and Without Standards............................ 52 3.3.2 High Forest Management.................................................. 58 3.4 The Development of Arable Cultivation and Arable Weeds........... 60 3.4.1 Pre-industrial Agriculture................................................. 60 3.4.2 The Effects of Technological Advances on Crop Fields and Low-Intensity Pastures............................................... 64

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3.5 The Development of Meadows, Intensive Pastures and Other Grassland........................................................................ 65 3.5.1 Straw and Fodder Meadows.............................................. 65 3.5.2 Continuous and Rotational Grazing.................................. 67 3.5.3 Agricultural Biocide Use, Energy Use and Crop Yield.... 69 3.6 Changes in Landscape Hydrology.................................................. 71 3.6.1 Modifications of River Valley Landscapes........................ 71 3.6.2 The North Sea Dykes and Their Consequences................ 74 3.6.3 The Destruction of Mires, and Attempts to Restore Them................................................................ 77 3.6.4 Increasing Exposure of the Vegetation to Drought........... 79 3.7 Chemical Pollution of the Environment and Its Impact on the Vegetation............................................................................. 79 3.7.1 Long- and Short-Range Effects of Chemical Pollutants... 79 3.7.2 Nutrient Enrichment of Soils and Water Bodies............... 80 3.7.3 Acid Deposition................................................................ 90 3.7.4 Sulphur Dioxide and Ozone Emissions............................ 92 3.7.5 Emissions of Heavy Metals and Other Substances........... 98 3.8 Changes in Game Densities and Their Effect on the Vegetation.... 104 3.9 Introduction of Non-native Plant Species....................................... 105 3.10 Recent Species Losses and Impoverishment of Plant Communities...................................................................... 106 3.11 The Effects of Recent Climate Change on the Vegetation.............. 108

Part III  General Ecology of Central European Forests 4 Abiotic Conditions, Flora, Ecosystem Functions and Recent Human Influence................................................................. 119 4.1 The Flora of Central European Forests........................................... 119 4.2 The Geographic Distribution of Forest Vegetation......................... 119 4.2.1 Zonal, Extrazonal and Azonal Forest Vegetation.............. 119 4.2.2 The Potential Natural Vegetation of Central Europe......... 123 4.2.3 Altitudinal Belts of Forest Vegetation............................... 124 4.2.4 Water and Temperature Limitations of Forest Growth...... 125 4.3 Environmental Conditions and Forest Habitat Classification......... 127 4.3.1 The Climate of the Forest Interior..................................... 127 4.3.2 Soil Water Regime............................................................ 134 4.3.3 Soil Chemical Properties................................................... 141 4.4 Comparative Ecology of Central European Tree Species............... 150 4.4.1 Important Characteristics of Crown Structure.................. 151 4.4.2 Traits Related to Productivity and Stress Tolerance......... 151 4.4.3 Nitrogen Acquisition......................................................... 165 4.4.4 Stress Tolerance................................................................ 166 4.4.5 Litter Quality and Tree Species Effects on the Soil.......... 179 4.4.6 Competitive Abilities of the Tree Species......................... 182 4.4.7 The Effects of Elevation on Tree Growth......................... 188

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4.4.8 The Influence of Climate on Elevational Changes in Tree Species Composition.............................. 191 4.4.9 Forest Cover in Central Europe and the Current Coverage of Major Tree Species....................................... 194 4.5 Forest Floor Plants and Shrubs of the Forest Interior: Ecological Niches and Ecological Grouping.................................. 196 4.5.1 Niches of Forest Shrubs.................................................... 196 4.5.2 Ecology of Forest Floor Plants.......................................... 207 4.5.3 The Ecological Grouping of Herbaceous Plants in Central European Broadleaved Forests.............. 243 4.6 Population Ecology of Forest Floor Plants..................................... 245 4.6.1 Phenology.......................................................................... 245 4.6.2 Life Cycles........................................................................ 249 4.7 Productivity and Cycling of Water and Nutrients........................... 252 4.7.1 The Biomass and Productivity of the Tree Layer.............. 252 4.7.2 The Biomass and Productivity of the Herb Layer............. 266 4.7.3 Ecosystem Carbon Cycling............................................... 270 4.7.4 Water Cycling.................................................................... 272 4.7.5 Nutrient Cycling................................................................ 283 4.8 Vegetation Dynamics...................................................................... 298 4.8.1 Tree Layer Dynamics........................................................ 298 4.8.2 Fluctuations and Succession in the Herb Layer................ 298 4.9 Recent Human Influence................................................................. 300 4.9.1 Forest Damage in the Past and the Present....................... 300 4.9.2 Anthropogenic Changes in Forest Soil Conditions........... 301 4.9.3 Recent Tree Damage and Its Potential Causes.................. 310 4.9.4 Anthropogenic Changes in the Herb Layer and in the Cryptogam and Fungal Flora of Forests........... 324 4.9.5 Conservation and Restoration of Forests.......................... 337

Part IV  Forest and Shrub Formations 5 Beech and Mixed Beech Forests............................................................. 351 5.1 The Classification of Hardwood Broadleaved Forests.................... 351 5.2 The Classification of Beech Forests in Central and Western Europe........................................................................ 356 5.3 Beech Forests on Rendzina and Pararendzina................................. 361 5.3.1 Mesic Limestone Beech Forests (Hordelymo-Fagetum)....................................................... 361 5.3.2 Mull Beech Forests Rich in Wild Garlic........................... 366 5.3.3 Sedge Beech Forests on Dry Slopes (Carici-Fagetum)............................................................... 368 5.3.4 Beech Forests Without a Herb Layer (Fagetum nudum).............................................................. 372 5.3.5 Yew-Beech and Seslerio-Fagetum Forests on Steep Slopes................................................................. 375

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5.3.6 Montane Beech and Fir-Beech Forests............................. 379 5.3.7 Subalpine Sycamore-Beech Forests (Aceri-Fagetum)...... 388 5.4 Beech and Mixed Beech Forests on Moderately Fertile Cambisols............................................................................ 391 5.4.1 The Galio odorati-Fagetum and Related Communities..................................................................... 391 5.4.2 Mixed Beech Forests on Moist Soil.................................. 399 5.4.3 Beech and Mixed Beech Forests Rich in Ferns................ 403 5.4.4 Beech Forests Rich in Festuca altissima........................... 408 5.5 Beech and Oak-Beech Forests on Highly Acidic Soils................... 409 5.5.1 Moder Beech Forests (Luzulo-Fagenion)......................... 409 5.5.2 Climatic and Edaphic Forms of Moder Beech Forests and Oak-Beech Forests......................................... 416 5.5.3 Acid Beech Forests on Limestone..................................... 421 5.6 A Comparison of the Habitats of Beech Forest Communities................................................................................... 422 5.7 Beech Forest Dynamics................................................................... 424 5.7.1 The Inter- and Post-Glacial Development of Beech Forests................................................................ 424 5.7.2 Patch Dynamics of Beech Forests..................................... 429

6 Mixed Broadleaved Forests Poor in Beech Outside of Floodplains or Mires........................................................................... 443 6.1 Maple- and Ash-Rich Mixed Forests.............................................. 443 6.1.1 Habitat Classification of Maple and Ash Forests.............. 443 6.1.2 The Fraxino-Aceretum...................................................... 446 6.1.3 The Aceri-Fraxinetum....................................................... 450 6.1.4 The Carici remotae-Fraxinetum........................................ 452 6.2 Mixed Lime Forests........................................................................ 454 6.2.1 The Asperulo taurinae-Tilietum in the Alps..................... 454 6.2.2 Mixed Tilia cordata Forests Outside of the Alps............. 456 6.2.3 Thermophilic Mixed Large-Leaved Lime-Maple Forests (Aceri platanoidis-Tilietum platyphylli)........................................................................ 457 6.3 An Overview of the Mixed Oak Forests of Central Europe............ 459 6.4 Thermophilic Mixed Oak Forests (Quercetalia pubescentis)........ 461 6.4.1 ‘Relict’ Submediterranean Downy Oak Forests and Continental Steppe Forests......................................... 461 6.4.2 The Quercetalia pubescentis Across  a West-East Climatic and Floristic Gradient..................... 466 6.4.3 The Subcontinental Potentillo-Quercetum........................ 474 6.5 Mixed Oak Forests on Acid Soils................................................... 476 6.5.1 The Betulo-Quercetum and Related Communities in Central Europe........................................ 476 6.5.2 Thermophilic Acid Oak Forests and Sweet Chestnut Coppices in Southern Central Europe............................... 490

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6.6 Oak-Hornbeam Forests (Carpinion betuli)..................................... 494 6.6.1 Thermophilic Subcontinental Oak-Hornbeam Forests (Galio-Carpinetum).............................................. 494 6.6.2 Moist Subatlantic Oak-Hornbeam Forests (Stellario-Carpinetum)...................................................... 497 6.6.3 Beech-Rich Oak-Hornbeam Forests................................. 502 6.6.4 Lime-Hornbeam Forests (Tilio-Carpinetum) Outside the Range of Beech.............................................. 508 6.6.5 A Comparison of the Environmental Conditions in Oak-Hornbeam Forests............................... 518

7 Pure and Mixed Coniferous Forests....................................................... 521 7.1 The Role of Conifers in the Forests of Central Europe................... 521 7.2 The Systematic Classification of Conifer Forest Communities........................................................................ 525 7.3 Silver Fir Forests............................................................................. 526 7.3.1 The Unique Position of Fir Communities in Central European Forests.............................................. 526 7.3.2 Fir Forest Communities of the Alps and Their Foothills............................................................ 529 7.3.3 Fir Forests of Low Mountain Ranges and Lowlands.................................................................... 537 7.4 Spruce Forests................................................................................. 541 7.4.1 The Natural Range and Habitats of Spruce Forests in Central Europe.................................................. 541 7.4.2 The Systematic Classification of Spruce-Rich Conifer Forests.................................................................. 545 7.4.3 Montane and Subalpine Spruce Forests............................ 549 7.4.4 The Role of Spruce in Lowland Areas.............................. 556 7.4.5 Environmental Conditions in Various Spruce Forest Communities.......................................................... 558 7.5 Subalpine Larch-Swiss Stone Pine Forests and Larch Forests............................................................................ 560 7.5.1 Environmental Conditions of Larch and Swiss Stone Pine Forests in the Central Alps.............................. 560 7.5.2 Larch-Swiss Stone Pine Forests in the Alps and the Tatra...................................................................... 565 7.5.3 Larch Forests of the Southern Alps and  Non-Alpine Larch Stands.................................................. 570 7.6 Mountain Pine Stands Outside of Mires......................................... 571 7.6.1 Erect Mountain Pine Communities................................... 571 7.6.2 Dwarf Mountain Pine Scrub Under Different Environmental Conditions................................................. 574 7.7 Pine Forests Outside of Mires and Floodplains.............................. 581 7.7.1 Central European Scots Pine Forests: Variation with Environmental Conditions......................... 581

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7.7.2 Scots Pine and Black Pine Communities in the Alps......................................................................... 585 7.7.3 A Comparison of Pine and Mixed Oak Forests in the Pleistocene Lowlands.............................................. 590 7.8 Conifer Forest Dynamics................................................................ 598 7.8.1 Conifer Regeneration........................................................ 598 7.8.2 Stand Dynamics................................................................ 600

8 Forest Plantations and Clearings............................................................ 607 8.1 Plantation Communities in Comparison to Semi-natural Forest Communities........................................................................ 607 8.1.1 Types of Plantation Vegetation.......................................... 607 8.1.2 Conifer Monocultures in Broadleaved Forest Habitats.................................................................. 622 8.1.3 The Vegetation of Clearings and Burnt Areas................... 626 9 Woody Vegetation of Floodplains and Swamps..................................... 633 9.1 Flora and  Origins............................................................................ 633 9.2 Habitat Conditions and Classification............................................. 636 9.2.1 Floodplain Morphology and Local Climate...................... 636 9.2.2 Soil Chemistry and Nutrient Supply................................. 640 9.2.3 Discharge Regime, Flooding Frequency and Soil Moisture.............................................................. 644 9.2.4 Stagnant and Flowing Groundwater.................................. 648 9.3 Vegetation........................................................................................ 652 9.3.1 Woody Vegetation of Floodplains and Riverbanks........... 652 9.3.2 Swamp and Mire Forests................................................... 688 9.4 Adaptations to the Environment...................................................... 700 9.4.1 Flood Tolerance of Floodplain Species............................. 700 9.4.2 Summer Drought Stress in Floodplains............................ 705 9.4.3 Willows as Characteristic Species of Floodplains and Swamps...................................................................... 706 9.5 Population Biology and Community Ecology................................ 708 9.5.1 Phenology.......................................................................... 708 9.5.2 River Valleys as Migration Routes for  Mountain Species.............................................................. 709 9.5.3 Regeneration and Population Dynamics in Floodplain Forests......................................................... 712 9.6 Productivity and Cycling of Water and Nutrients........................... 715 9.6.1 Forest Structure, Biomass and Productivity...................... 715 9.6.2 Water and Nutrient Cycling.............................................. 716 9.7 Vegetation Dynamics...................................................................... 718 9.7.1 Dynamics of Floodplain Vegetation.................................. 718 9.7.2 Succession Following Disturbance................................... 721 9.8 Human Influence............................................................................. 722

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9.8.1 Exploitation, Drainage and Destruction of Floodplain Forests......................................................... 722 9.8.2 Conservation and Restoration of Floodplain Forests........ 726

10 Epiphyte Vegetation................................................................................. 729 10.1 Tree Bark as an Epiphyte Substrate................................................ 729 10.2 Epiphytic Algal, Lichen and Bryophyte Communities................... 731 10.2.1 Alga-Rich Epiphyte Communities.................................... 732 10.2.2 Lichen-Rich Epiphyte Communities................................. 733 10.2.3 Bryophyte-Rich Epiphyte Communities........................... 734 10.3 Adaptations to the Environment...................................................... 737 10.3.1 Important Ecological Properties of Epiphytic Cryptogams....................................................................... 737 10.3.2 Carbon Assimilation as a Function of Moisture, Light Intensity and Temperature....................................... 738 10.3.3 Chemical and Physical Properties of the Substrate........... 740 10.3.4 The Effects of Toxic Substances and the Role of Epiphytes as Indicators................................................. 741 10.3.5 The Importance of Stand Structure and Stand Age........... 742 10.4 Recent Changes in Epiphyte Communities..................................... 743 11 Forest Edges, Scrub, Hedges and Their Herb Communities............... 747 11.1 Flora and  Development................................................................... 747 11.2 Environmental Conditions and Habitat Classification.................... 750 11.3 Vegetation........................................................................................ 756 11.3.1 Forest Edges, Scrub and Hedges....................................... 756 11.3.2 Herb Fringe Communities................................................. 763 11.4 Adaptations to the Environment, Population Biology and Vegetation Dynamics................................................................ 768 11.5 Human Influence............................................................................. 771 11.5.1 Decline and Destruction of Hedges.................................. 771 11.5.2 The Importance of Hedges for Agriculture and Agricultural Landscapes............................................. 773 12 Syntaxonomic Overview of the Vascular Plant Communities of Central Europe: Forest and Scrub Formations............................................................................. 775 References......................................................................................................... 781 Index.................................................................................................................. 891

Directions for Use

This two-volume work aims to give readers an introduction to the vegetation and ecology of the wide variety of plant communities that can be found in Central Europe. Volume I covers all types of forest and scrub vegetation, be it natural or man-made. This volume (Vol. II) covers all types of non-forest vegetation in Central Europe, from the mostly natural coastal, mire, freshwater and alpine formations to the broad array of man-made habitats including managed grasslands, heathlands and arable fields, as well as ruderal and urban ecosystems. Volume II consists of 14 chapters and includes a summarising chapter (14) on the synsystematic classification of these formations. The chapters are arranged in sequence from the natural non-forest formations to the purely anthropogenic ones, starting with coastal ecosystems (Chaps. 1 and 2: salt marshes and dunes), followed by semiaquatic ecosystems (Chaps. 3 and 4: mires and freshwater systems) and alpine and nival ecosystems (Chap. 5). The subsequent group of anthropogenic systems starts with heathlands (Chap. 6) and managed dry or mesic to wet grasslands (Chaps. 7 and 8) and proceeds to the vegetation of heavy metal-rich soils (Chap. 9) and ruderal wet or mesic to dry habitats (Chaps. 10 and 11), ending up with the heavily disturbed vegetation (Chaps. 12 and 13: vegetation of arable land and urban areas). The main chapters on the formations were in most cases structured into the following eight subchapters: 1 . Flora and development 2. Environmental conditions and habitat classification 3. Vegetation 4. Adaptations to the environment 5. Population biology and community ecology 6. Productivity and water and nutrient cycling 7. Vegetation dynamics 8. Human influence The climatic, geological and pedological characteristics of Central Europe and the history of human impact are discussed in detail in Chaps. 1, 2, and 3 of Vol. I and are not repeated in this volume.

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Directions for Use

This book is based on the Central European system of floristically defined (phytosociological) vegetation types, which allows us to analyse the spectrum of natural and man-made ecosystems and gives insights into their ecology and dynamics. The vegetation classification system most widely used in Central Europe is the one introduced by Josias Braun-Blanquet around a hundred years ago, which was subsequently modified and combined with numerical vegetation analysis. It uses diagnostic species (‘character species’ and ‘differential species’) to identify combinations of plant species that co-occur on a regular basis and thus can be considered as community types (syntaxa). Analogous to the taxonomic system of classification, syntaxa are organised in hierarchical levels of relatedness. This book does not attempt to give a complete syntaxonomic overview of the Central European vegetation but instead refers to existing syntheses, notably the work of Oberdorfer et al. (1987– 1992) and Oberdorfer (ed., 1992–1998), Pott (1995), Dierschke (1996 et seqq.) and Rennwald (2000) for Germany and that of Chytrý (2007 et seqq.) for Czech Republic, of Schaminée et al. (1995 et seqq.) for the Netherlands and of Mucina et al. (1993) and Willner and Grabherr (2007) for Austria. Here, only the widespread community types that are frequently referred to in the literature (associations, together with alliances and orders) will be presented. To highlight the most important associations, their names are underlined when mentioned for the first time in the text. An overview of the most important higher syntaxonomic units (classes, orders, alliances) is given in Chaps. 14 (this volume: non-forest vegetation) and 12 (in Vol. I: forest and scrub vegetation). The coverage of species in a relevé is indicated in the vegetation tables by the numbers 1–5 and the symbol  +  (+  = rare, 1 =  6000 g dm m−2 year−1, Long and Woolhouse 1979), which are comparable to the highest agricultural yields. This productivity is aided by the ability of Spartina to use different sources of C. In addition to the uptake of gaseous CO2 through its leaves, Spartina also stores CO2 in the aerenchyma of its stem, and can also carry out the dark reaction of photosynthesis in the roots fixing dissolved bicarbonate from the sediment with the enzyme phosphoenolpyruvate carboxylase (PEPCO) (Hwang and Morris 1992). It is interesting to note that the photosynthetic apparatus of S. anglica and S. alterniflora is relatively insensitive to salt, and even needs low Na concentrations to function, making it more productive in saltwater than in freshwater (Longstreth and Strain 1977; Drake 1989). The halophyte flora of the Pannonian region includes two other native C4 plants (the Poaceae Crypsis aculeata and the Chenopodiaceae Camphorosma annua). Effects of Submergence  Flooding not only influences the access of the plant to carbon dioxide, but also creates reducing conditions and toxicity of iron, manganese and sulphide in the sediment, particularly when this is fine-grained. The S2− and Mn2+ concentrations are generally, but not always, greater in the lower mudflats covered by Salicornia communities than in the upper marshes. Havill et al. (1985) measured concentrations of up to 400 μM sulphide in the sediment in a Salicornia community. This gradient from lower to upper marsh is linked to a higher sensitivity of the plants of upper salt marshes to elevated sulphide, Fe2+ and Mn2+ concentrations. For example, Salicornia is relatively tolerant of Mn2+ (up to 10 mM) and S2− (up to 50 μM; Cooper 1984; van Diggelen et al. 1987). High sulphide tolerance has also been recorded for other species of the lower salt marshes, such as Aster tripolium (Havill et al. 1985; Rozema et al. 1985a). A well developed aerenchyma in the roots reduces the toxic effects of certain ions in the reduced medium, as it can release oxygen into the rhizosphere, oxidizing the immediate surroundings of the root (see Sect. 4.4.3). Nevertheless, anoxia also causes reductions in growth in the plants of the lower marsh, such as in Spartina, via the effect of S2− (King et  al. 1982). These observations have led vegetation ecologists to hypothesise that the clear vertical zonation of the vegetation along the North Sea coast is not only caused by the salinity gradient, but also by the variation in tolerance of the halophytes to high sulphide, iron and manganese concentrations (Adam 1990). Cooper (1982) predicted that the species of the upper marsh zone, such as Festuca and Juncus, suffer a reduction in growth due to flooding and salinity, whist flooding promotes the growth of Puccinellia (saltmarsh grass) and Salicornia (glasswort) of the lower marsh zone. Rozema et al. (1985a) tested this in a two-factorial experiment with 15 halophytes from the upper and lower salt marshes of the North Sea coast with the conditions high or low salinity (746 or 42 mM) and with or without flooding. They

1.4  Adaptations to the Environment

41

Fig. 1.21  Relative tolerance of 15 salt marsh plants to flooding (permanent water level 5 cm above the soil surface, (a)), salinity (750 mosmol kg−1 NaCl, (b)) and a combination of the two (c) under experimental conditions in a greenhouse. The tolerance index expresses the growth rate of a plant under each treatment relative to its growth without this stress factor. Values >1 mean an increase, values 50 cm above mean high tide). By the end of the observation period, Elymus athericus dominated on almost all of the domes that were over 25 cm above mean high tide. As seagulls often nest here, numerous ruderal plants had also established (see Fig. 1.25).

52

1  Salt Marshes and Inland Saline Habitats

In the lee of this sandy salt marsh complex, true mudflat salt marshes and large stands of Spartina similar to those in the dyked coastal areas form. The first plants of Spartina anglica appeared on Mellum in 1954 and proceeded to cover large areas of the island, causing fine-grained mud to accumulate over the sand, i.e. acting as marsh pioneers. The edges of the channels that formed were stabilised by sea rush (Juncus maritimus), whilst wet depressions in the salt marsh that had become less saline from collecting rainwater were colonised by Schoenus ferrugineus. On the barrier islands, accretion is typically higher in the lower marshes than in the higher marshes which leads to the development of large areas suitable for the colonisation by intermediate succession species such as Puccinellia, Limonium, Triglochin and Plantago (Bakker 2014). On Mellum, brackish water reedbed communities of Bolboschoenus and/or Phragmites developed where rhizomes had washed up around the edges of these depressions. They were flooded by water from the North Sea during high tide, demonstrating the salt tolerance of these species. The example of the vegetation development on Mellum clearly shows that natural salt marsh development results in a heterogeneous microrelief with a mosaic of halophyte communities. A recent analysis of the vegetation dynamics on the island of Trischen (Schleswig-­ Holstein) reveals that four decades of natural island development have increased the overall diversity of vegetation types on the island, while plot-level species richness generally declined mostly due to the spread of Elymus athericus and Festuca rubra (Stock et al. 2014). On smaller Wadden Sea islands, seabird colonies are likely to have an important influence on plant species composition by promoting the ­establishment of annual, ruderal and cosmopolitan species while native species decline (Ellis 2005). The widespread dyking of the German and Dutch Wadden Sea coast led to succession in the newly-formed polders, which was studied e.g. by Joenje (1978) and Wolfram et al. (1998).

1.6.5  The Colonisation of a Muddy Island in the Baltic Sea Conditions similar to those in the intertidal mudflats were temporarily provided in the Baltic in 1945 on the newly created muddy island of Bock to the southwest of Hiddensee (west of Rügen). Voderberg and Fröde (1958, 1967) recorded this primary succession in the mesohaline littoral of the Baltic. By 1946, the initially bare mud was covered by Salicornia plants. They remained longest in the wet zone at the edge of the water, but were replaced by the Puccinellietum maritimae or species of the Juncetum gerardii in higher-lying areas. The Puccinellietum also covered an area of the island of Bock that was unusually large for the Baltic region. This may have been due to the shallow sloping relief and the high nutrient availability in the newly deposited mud. The seeds washed up onto the island led initially to a random distribution of species on all levels of the island. Fertile soils were then quickly dominated by highly

1.7  Human Influence

53

competitive species, so that Voderberg (1955) found most halophyte communities to have developed their typical species composition within 10 years of the island’s creation. Voderberg (1955) also recorded woody species above the supralittoral, demonstrating their natural role in coastal habitats. Where the tidal range is low and the soil is not permanently saline, forest can develop directly behind the reedbed zone linking land and sea.

1.7  Human Influence 1.7.1  Effects of Salt Marsh Grazing Humans have used the grassy vegetation in the marshes of the North Sea coast for over 2500 years, mainly through grazing with cattle and sheep. By the 1980s, well over half of the 360 km2 of salt marsh along the Dutch, German and Danish Wadden Sea was intensively grazed, and in Schleswig Holstein this was over 90 % (Stock et  al. 2005). It is only since 1985 with the creation of the Wadden Sea National Parks that larger areas have been left to natural succession. Grazing means not only the removal of nutrients, but also permanently influences the species composition and dominance relationships in the salt marsh. Bos et al. (2002) compared grazed and ungrazed plots on the Dutch and Danish coasts of the Wadden Sea, and found that 23 of 30 common species were clearly promoted by grazing, and only 4 grazing-sensitive species declined. These species, Atriplex portulacoides, Elymus athericus, Artemisia maritima and Aster tripolium, were usually strongly suppressed or grew only at a greater distance from the dyke. Sheep graze very selective in particular on Aster. In contrast, the species particularly promoted by grazing included grazing-tolerant graminoids with rhizomes or above-­ ground stolons (e.g. Festuca rubra, Puccinellia maritima and Juncus gerardii), low-growing and light-demanding species (e.g. Glaux maritima, Armeria maritima) and annual species of disturbed ground (Salicornia and Suaeda species). Less palatable or tough species also profit from grazing, as they are usually avoided by livestock (e.g. Juncus gerardii, Adam 1978). Species that mainly reproduce generatively are usually disadvantaged by grazing (Hutchings and Russel 1989), although this is somewhat compensated for by the dispersal of seeds via winter storms. Bakker (1989) investigated the seeds that had been washed up at the drift line and found particularly large numbers of seeds of Elymus athericus, Festuca rubra and Juncus gerardii, and these species were almost always present in the seed bank (see Table 1.6). In fact, the majority of common species in the established vegetation were also present in the seed bank (Bakker 1989). Moderate grazing intensity, particularly by sheep, increases the shoot density and causes a dense mat of roots to form. Species well adapted to this type of grazing include particularly the grasses such as Puccinellia maritima, and low-growing rosette plants like Plantago maritima and Armeria maritima. According to Erchinger et al. (1996), these species increase the shear strength of the salt marsh sediment.

1  Salt Marshes and Inland Saline Habitats

54

Table 1.6  The influence of grazing abandonment (Ab), mowing in August (Mo) and grazing (Gr) on the coverage of species (in %) in three salt marsh communities on the island of Schiermonnikoog (Dutch Wadden Sea) at the end of a 10-year experiment

Management Year (19..) Species no. Limonium vulgare Artemisia maritima Puccinellia maritima Triglochin maritima Glaux maritima Juncus gerardii Agrostis stolonifera Festuca rubra Atriplex prostrata Plantago maritima Armeria maritima Elymus pycnanthus Salicornia europaea Spergularia maritima Suaeda maritima Spergularia salina Atriplex portulacoides Sagina procumbens Bellis perennis Rumex acetosella Poa annua

Festuca rubra ssp. littoralis community Juncetum gerardii Agropyretum littoralis 0 Ab Mo Gr Sb 0 Ab Mo Gr Sb 0 Ab Mo Gr Sb 71 81 81 81 81 71 81 81 81 71 81 81 81 10 5 11 11 15 8 10 12 10 7 5 3 11 13 19 3 4 8 2 • • 6 2 1 • 1 1 1 2 1 1 6 4 1 8 1 2 3 6 3 30 1 4

2 50 1

1 1 80 1 1 1 1 1

1 1 30 20 30 4 4

• • ● ● ● • •

1 1

8 6 35 2 3 6

4 1 4 1 2 40 2

1

2 1 70 1 2 4 1 1

12 2 4 1 1 1

● ● • 1 1 • 3

1 2 70 1 1 1 1 30 70 8

1

1 1

● ● • •

1 8

●

2

•

1 •

1

12 20 1

1 1

1 1 1

● • •

● • • •

Coverage at the start of the experiment (1971) = 0. The occurrence in the seed bank (Sb) of the abandoned site is also shown (1977/78 and 1982): ● frequent, • infrequent. From Bakker (1989), species list has been shortened

However, grazing also has a clear negative effect on the sedimentation rate, as these ‘golf course’ swards are less effective at trapping mud particles than the tall vegetation of ungrazed salt marshes (Dierßen et al. 1994; Bakker 2014; Nolte et al. 2015). Ungrazed salt marshes are therefore more likely to contribute to a positive sediment balance to compensate for future sea level rises than grazed marshes (Dijkema et al. 1990). Current mean surface elevation increase is around 6 mm per year in the salt marshes of Schleswig-Holstein (Suchrow et al. 2012). Not only sheep and cattle, but also wildfowl such as brent geese graze the salt marshes in large numbers. Winter goose grazing causes a reduction in productivity in the following summer in Puccinellia swards in Schleswig-Holstein (Bredemeier et al. 1999), and not a stimulation of growth as postulated by Hik and Jefferies (1990).

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55

Fig. 1.26  Aerial photograph of a grazed lower saltmarsh with artificial drainage system near Husum in the German Wadden Sea (Schleswig-Holstein). The grey patches are Artemisietum maritimae, the yellow-brown areas mostly the Festuca rubra subsp. littoralis-community and the green areas Puccinellietum maritimae

Cattle and sheep grazing causes compaction of the sediment and therefore a reduction in oxygenation, so that the soils of the grazed upper salt marshes become more similar to the hydromorphic soils of the lower marshes. This explains why grazing usually leads to a spatial shift in the Puccinellietum towards the foot of the dyke, i.e. a reduction in area of the upper salt marsh zone. Grazing can also increase the denitrification rate (Jensen 1989; Svendsen 1992) and change the salinity of the soil. Schmeisky (1974) found an increase in salinity due to grazing, as the patches of bare earth created experienced higher rates of evaporation and salt accumulated at the surface due to capillary rise. In contrast, Bakker (1989) observed the salinity to decrease following grazing. The grazing pressure by geese, hare and other natural herbivores is particularly high in the initial and intermediate stages of salt marsh succession, while late-successional plants such as Elymus are avoided (Bakker 2014). Horse grazing may help to reduce Elymus athericus as was observed in the Dutch Wadden Sea. The grazing of salt marshes usually occurs along with marsh drainage, achieved using a network of drainage ditches (see Fig. 1.26). This probably influences the salt marsh ecosystem more than the grazing itself. The majority of grazing-sensitive species manage to recolonise salt marshes within 5–10 years after grazing has been abandoned. Yet drained marshes lack the rich vegetation mosaic characteristic for natural salt marshes with undisturbed hydrology (see the example of the island of Mellum in Sect. 1.6.4). This results in a reduction in the diversity of different micro-

56

1  Salt Marshes and Inland Saline Habitats

habitats in the marsh and therefore lower plant species richness (Kiehl 1997). Stands dominated by Elymus athericus, E. × oliveri and Festuca rubra, which are of little value for nature conservation, often form on areas of abandoned pasture, where dense networks of drainage ditches have led to a homogenisation of the hydrological regime.

1.7.2  Salt Marsh Succession After Abandonment of Grazing Since the mid-1980s, conservation efforts have reduced or stopped the grazing in many salt marsh areas of the German and Dutch North Sea Coast. This has initiated secondary succession processes over an area of nearly 100 km2 of salt marsh. In the intensively grazed mudflat salt marshes of the mainland, this abandonment of grazing led to a striking change in the vegetation (Kiehl et al. 1997; Heinze et al. 1999; Esselink et  al. 2000; Bos et  al. 2002; Schröder et  al. 2002; Wanner et  al. 2014; Rupprecht et al. 2015). In the lower salt marsh, large areas of Puccinellietum were first replaced by sea aster stands, and then by homogenous Halimionetum scrub, so that the Puccinellietum only remained in the lower regions around the mean high tide mark. In the upper marsh zone, Elymus athericus and its hybrids were promoted in the areas receiving high sediment loads rich in N (see also Bakker 2014). Festuca rubra and Artemisia maritima also increased in number, whilst Juncus gerardii declined (see Table 1.6). These changes also affected the grazing geese, which used the salt marshes of the Wadden Sea in autumn and winter in huge numbers. They avoid ungrazed marshland, as they do not eat Atriplex and Elymus species (Bos et  al. 2002). In the upper marsh, the plot-level plant diversity under low-intensity grazing (1.5–3 sheep per ha) was therefore generally higher than under high-intensity grazing, but also higher than under no grazing at all (Berg et al. 1997; Wanner et al. 2014). The dense stands of E. athericus reduced small-scale plant species richness (Wanner et al. 2014). In contrast to the sheep-grazed marshes of Schleswig-Holstein, Bakker et  al. (2003) found higher species richness under intensive cattle grazing than no-grazing areas in all studied scales (0.01–2500 m2) in the mainland salt marshes of the Dutch Wadden Sea, 30 years after grazing was abandoned in part of the region. If the drainage of the marsh is also reduced or stopped, then Spartina, Salicornia and Suaeda become more widespread in the lower marsh and the heterogeneity of habitats increases through the formation of depressions and the silting up of ditches (Esselink et al. 2000), increasing plant diversity at larger scales. Overall, abandonment of grazing leads to the expansion of the upper salt marsh because most upper marsh species are less grazing-tolerant than lower marsh species (Bakker 1989). Consequently, the area of the lower marsh has decreased, with a corresponding reduction in halophyte vegetation. Salt marshes on the sandy Wadden Sea islands were always less intensively grazed than those seawards of the dykes. The cessation of grazing on the islands Terschelling and Schiermonnikoog and in Skallingen led over 25 years to a reduc-

1.7  Human Influence

57

tion in the species richness of 4 m2 plots (see Fig. 1.11). As mentioned in Sect. 1.7.1, of the 30 common species, only 4 were promoted by the lack of grazing, whilst 23 declined (Bos et al. 2002). Low-level disturbance (grazing at 0.5–1.5 cattle or sheep per ha) therefore also promoted species diversity here, supporting the intermediate disturbance hypothesis of Connell (1978). From a nature conservation perspective, a large-scale mosaic of different management regimes and extended regions with natural drainage would be most favourable for the phytodiversity of the salt marshes of the Wadden Sea and would benefit many, but not all, vertebrates and invertebrates (Bakker 2014). Geese and many waders, for example, prefer the short, grazed-down salt marsh swards resulting from grazing with more than 1 cattle or sheep per ha. The basic successional diagram created by Dijkema (1983) gives the impression that the cessation of grazing would lead to a directional salt marsh succession with a clearly defined climax community (right-hand side of Fig. 1.24). This does not, however, occur in reality, as demonstrated by a quantitative analysis of the results of repeated mapping of a salt marsh on the North Frisian holm Oland, carried out by Leuschner and Riemer (unpublished) in 1987 and 1998. The 70 ha area had not been grazed since the 1980s, and drainage was stopped in 1986. Figure 1.27 shows that the succession here is not unidirectional, but rather the nine most frequent salt marsh communities are linked by multiple developmental paths. Transitional forms between two communities can develop in both directions, as visible in the Puccinellietum and the Festuca litoralis community. This is probably due to the fact that sedimentation processes and increased waterlogging can take place at very small scales and thus lead to neighbouring vegetation patches developing in opposite directions. The most important change during the 11 years between the mapping periods was the expansion of the Agropyretum littoralis in the upper marsh zone, the Halimionetum portulacoidis in the lower marsh zone and the Salicornietum strictae on the mudflats. The Festuca litoralis community decreased in extent, as did the Puccinellietum and the Spartinetum. The secondary succession after the abandonment of mesohaline brackish water bays and estuaries like the Dollart bay proceeded somewhat differently. Phragmites australis dominated the lower salt marsh after the cessation of grazing, whilst Elymus repens (and not E. athericus as in the euhaline dyked marshes) dominated the upper marshes (Esselink et al. 2000).

1.7.3  Eutrophication of Coastal Waters Baltic Coast  The majority of lagoons and backwaters along the German and Polish Baltic coast received much higher inputs of nitrate and phosphate in the period from 1970 to around 2000 than ever before (Hübel et al. 1998; Hübel and Dahlke 1999; Schiewer 2008a). This led to excessive growth of phytoplankton and epiphytic algae, the decomposition of which consumes oxygen and produces hydrogen sulphide on the bed of the lagoon. In the 1970s, the plankton communities still went through a characteristic sequence of blooms of flagellates, diatoms, green algae and

58

1  Salt Marshes and Inland Saline Habitats

Fig. 1.27  Schematic diagram of the succession of the salt marsh vegetation on the holm Oland (Wadden Sea of Schleswig-Holstein) for the period between 1987 and 1998. The numbers show the area (in ha) that changed from one vegetation unit to another over the 11 years between the two mapping periods. Arrows in both directions (e.g. between Agropyretum and Puccinellietum) show that the Puccinellietum mainly changed into Agropyretum, but also small areas of Agropyretum were replaced by Puccinellietum (From Leuschner unpublished; Riemer unpublished)

cyanobacteria, but today the coastal lagoons are dominated by green algae and cyanobacteria of nano- and pico-plankton size (Schiewer 2008a). The increased density of planktonic algae reduces the light penetration and the depth at which macrophytes can grow, whilst epiphytic algae have also increased (Schramm 1996). The shallow lagoons of Mecklenburg-Vorpommern were characterised by stonewort communities into the 1970s, whilst they were still in an oligotrophic or mesotrophic state (Lindner 1978). In the Bay of Gdańsk, the Fucus vesiculosus, Zostera marina and red algae (Furcellaria) stands that were widespread in the 1950s have now been almost completely replaced by the eutrophic species Zannichellia palustris, Ectocarpus and Pilayella littoralis (a thread-like brown alga) (Kruk-Dowgiallo 1991; Kruk-Dowgiallo and Szaniawska 2008). The losses have been similarly dramatic in the West Pomeranian lagoons, such as the Bay of Greifswald, where Zostera, Potamogeton and green algae stands covered only 3 % of the area in 1988,

1.7  Human Influence

59

compared to 50 % in the 1930s (Messner and v. Oertzen 1991; Schiewer 2008c). The excessive nutrient inputs caused a catastrophic decline in submerged macrophytes in the 1980s. The species that initially profited from these changes were Potamogeton pectinatus, Myriophyllum spicatum and Ruppia spiralis, and in eutrophic to polytrophic conditions Cladophora, Pilayella and Enteromorpha species (Hübel and Dahlke 1999, Blümel et al. 2002, Schiewer 2008b). The open waters of the western Baltic also saw a reduction over the last 30 years in Fucus vesiculosus, Laminaria and Furcellaria stands, replaced in parts by red algae (Breuer and Schramm 1988). The extent of Zostera marina stands in southern Denmark declined by 50 % between 1900 and 1980, and their depth range had decreased by 4–5 m. The eutrophication of many brackish water lagoons on the Baltic coast led to a transition from macrophyte dominance to phytoplankton dominance, laying the sediment surface bare. This occurred, for example, in the Darss-Zingster Bodden chain and the Greifswalder Bodden west and east of the island of Rügen (Blindow and Meyer 2015). As in the Wadden Sea, the earlier P limitation of plant growth was replaced by N and light limitation, promoting potentially toxic N2-binding cyanobacteria. The latter are dangerous not only for the fauna, but can also damage macrophytes like Zannichellia palustris via allelopathic effects (van Vierssen and Prins 1985; Lampe 1996). The increase in plankton biomass furthermore leads to a ­build-­up of detritus, and thereby the growth of mud layers on the bed of the water body that can become unstable without consolidating effects of the macrophytes. In this condition, the lagoons and backwaters had largely lost their previous function as ‘settling basins’ for the freshwater inflow from the mainland before it reaches the Baltic Sea. Since the early 1990s, improved wastewater treatment has reduced the nutrient loads, but many coastal lagoons have not yet been recolonised by submerged macrophytes as the water remained turbid with high concentrations of chlorophyll (Blindow and Meyer 2015). In lagoons with remaining cover of submerged macrophytes, the species composition has shifted with eutrophication from small charophytes to larger, more competitive angiosperms such as Potamogeton pectinatus (I. Blindow, pers. comm.). Gocke et  al. (2003) have summarised the change in primary producers in the Schlei in the course of its rapid eutrophication (see Fig.  1.28). In this brackish Schleswig-Holstein inlet, the submerged vascular plants were first colonised by epiphytic algae such as Pilayella, Enteromorpha and Ceramium, until they died from lack of light. The increase in phytoplankton density was accompanied by a phase during which free-floating algal mats covered the sediment and the banks, leading to a decrease in the area of reedbed. Wadden Sea  Three major human impacts have greatly altered the ecosystem functioning in the Wadden Sea over the last 50 years: (i) more than 40 years of high anthropogenic nutrient loads (Eriksson et al. 2010), (ii) the dramatic loss of seagrass beds in most of the Wadden Sea since the 1930s, and (iii) the removal of large parts of the intertidal mussel beds and high-density cockle-beds since the 1990s through shellfish dredging and bottom trawling. Nutrient inputs from the major rivers and the North Sea led to large-scale eutrophication of the Wadden Sea ecosystems in the

60

1  Salt Marshes and Inland Saline Habitats

Fig. 1.28  Changes in dominant primary producer groups during the eutrophication of the Schlei, a brackish water inlet of the Baltic coast of Schleswig-Holstein (Modifed from Gocke et al. 2003)

Fig. 1.29  Decrease of nitrogen and phosphorus loads since the late 1970s transported to the coastal North Sea by the rivers Elbe, Weser, Ems, Rhine and Maas and other minor streams (modified from Quality Status Report Wadden Sea 2009)

1970s–1990s (see Fig. 3.43 in Volume I), where microalgal production at first was significantly stimulated by high P loads (de Jonge et  al. 1993; Brockmann et  al. 1994). For example, from the 1970s to the 1990s, the production of epibenthic and pelagic microalgae has roughly doubled in the Wadden Sea around Sylt (Gätje and Reise 1998) with smaller flagellates increasing the most (Philippart et al. 2007). The nitrate and phosphate concentrations in the seawater peaked in the 1980s and 1990s and significantly decreased thereafter (see Fig. 1.29), resulting in a general decline in pelagic production since around the 1990s (Philippart et al. 2007). This nutrient pulse means that the growth of algae in the Wadden Sea is generally no longer limited by P, but by silicate for the diatoms and nitrogen for the rest of the phytobenthos and phytoplankton (Reise et al. 1998). Nitrogen limitation is assumed to occur at

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61

Fig. 1.30  Change in the state of the western Wadden Sea in the last 50 years from a system with low human impact and low external nutrient input, high sediment stability and dominance of benthic filter feeders to a largely disturbed system with widespread absence of benthic filter feeders but dominance of benthic deposit-feeders and low sediment stability as caused by high nutrient levels and human engineering works (schematic, after Eriksson et al. 2010)

N:P ratios in the water under 16 (Brockmann et al. 1994). Eutrophication has also increased the nutrient contents of the salt marsh sediments. However, eutrophication is not the only stressor to the Wadden Sea ecosystems in recent time. The fundamental change which took place in this environment in the past 50 years was summarised by Eriksson et  al. (2010), as shown in Fig.  1.30. Accordingly, the tidal flats have shifted from a historical state dominated by sediment-­stabilising ecosystem engineers (seagrass meadows and mussel beds) to a modern state dominated by sediment-destabilising macrofauna (mainly p­ olychaetes). Thus, dredging, trawling and eutrophication have greatly disrupted the internal regulation functions of the native soft-bottom communities and have exposed the system to regulation largely by external drivers, i.e. high inputs of nutrients and organic matter from the North Sea and high turbidity due to human disturbance. Correspondingly, birds feeding on marine worms increased in the Dutch Wadden Sea at the expense of species mainly feeding on mussels (van Roomen et al. 2005).

Chapter 2

Sand Dunes and Their Vegetation Series

2.1  Flora and Vegetation Dunes are very variable and highly dynamic habitats supporting a large number of plant communities. Depending on the age of the dune and its distance to the sea, it is dominated by plants adapted to high salinity, being buried by sand, drought or lack of nutrients. The herbaceous dune vegetation alone includes no fewer than seven different vegetation classes, which are present in the form of characteristic dune communities. These are the Cakiletea maritimae (shoreline communities), Ammophiletea arenariae (dune communities), Koelerio-Corynephoretea (open swards of sandy and rocky ground), Calluno-Ulicetea (matgrass swards and dwarf shrub heaths), Isoëto-Nanojuncetea (short-lived dwarf-sedge pioneer communities), Scheuchzerio-Caricetea nigrae (small-sedge fen and hollow communities) and Oxycocco-Sphagnetea (raised bog and wet heath communities). The resulting flora is very species-rich for Central European standards. On the 25 or so islands off the Dutch, German and Danish North Sea coast alone, around 900 vascular plants, 300 bryophyte and 300 lichen species were recorded (Dijkema and Wolff 1983), of which many occur in the dunes. Many of these taxa have inland origins, but have colonised coastal habitats with morphologically and physiologically adapted genotypes (Akeroyd 1997). We will discuss this plant diversity first in the context of the beach vegetation, then moving to the white and grey dunes and finally to the older dune scrub.

2.2  Dune Formation and Destruction Dunes can only form where the wind blows over areas of bare sand, constantly transporting and depositing particles on its slopes (see Fig. 2.1). This occurs at wind speeds of around 4 m s−1 (4 on the Beaufort scale, Bagnold 1954). Today, coastal © Springer International Publishing Switzerland 2017 C. Leuschner, H. Ellenberg, Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats, DOI 10.1007/978-3-319-43048-5_2

63

64

2  Sand Dunes and Their Vegetation Series

Fig. 2.1  White dunes with Ammophila arenaria on the Baltic coast. In the foreground on the left: a young, fast-growing dune, in the background on the right: older dunes partially destroyed by a storm. These have been stabilised by the roots of Ammophila

dunes cover around 480 km2 in the Netherlands, 96 km2 in Germany and 350 km2 in Poland (Doody 1991). Before the forest recolonised Central Europe after the last ice age, these expanses of sand were also found in inland areas, mainly along the large glacial valleys of the Rhine, Elbe, Oder and Vistula. These inland dunes are, however, no longer active, and creation and destruction of dunes now occurs almost exclusively in coastal areas.

2.2.1  Coastal Dunes of the North Sea 2.2.1.1  Coastal Dynamics The North Sea coast of Germany and the Netherlands is a sandy coast with currents flowing parallel to the coastline towards the north (or east). The swash transports moderately large-grained sand (0.25–1 mm diameter) to above the mean high tide line on the beach and mixes it with bits of sea shell, plant and animal remains and other flotsam. The finer particles are not deposited directly on the beach, but in the mudflats formed in quieter bays and the lee sides of islands. The North, West and East Frisian islands are part of a barrier system of sandbanks, dune islands and dune chains just off the coast, stretching from the Rhine estuary to Denmark. The main factor determining the coastal morphology is the tidal range. If it is less than 1.5 m, then a graded shoreline forms with a contiguous berm and long foredunes and dune slacks parallel to the shore, as in western Holland

2.2  Dune Formation and Destruction

65

Fig. 2.2  Schematic diagrams of the three island types that form on the tidal coast of the southeastern North Sea under different currents and tidal ranges (see text) (From Doing 1983)

and western Jutland. If the tidal range is larger, meaning a greater destructive power of the water, then barrier islands are created with a characteristic hooked form and a dynamic dune landscape. Such islands are found in the West and East Frisian islands between Texel and Wangerooge and the North Frisian and Danish islands between Amrum and Fanø (Behre 1995). If the tidal range exceeds 2.9 m (e.g. in the Weser estuary it is around 3.6 m), then open mudflats with small, very dynamic sand and vegetation patches form, as the ebb tide has a particularly high erosive power (Pott et al. 1999). This coastal type is found in the estuaries of the Ems, Weser, Elbe and Eider. Doing (1983) differentiated three morphogenetic types of barrier islands in the Wadden Sea, showing characteristic zonation of dune and salt marsh habitats (see Fig. 2.2). Type 1 islands have a centre of Pleistocene moraine deposits (‘geest’), which is surrounded by a series of dunes bent landwards at both ends (e.g. Sylt, Amrum and Texel). Type 2 islands do not have this geest centre, and the seaward dunes and the mudflats behind them stretch from northwest to southeast. These include the majority of West and East Frisian islands. Type 3 islands are relatively small islands close to river mouths, and are unstable due to the large tidal range and fast currents. These islands are characterised by horseshoe-shaped dunes that are open to the east, such as on Borkum, Memmert or Trischen. The dune belts of the southern North Sea coast can also be classified according to the lime content of the deposited sand. Lime-rich sands (3–10 % CaCO3) are found on the Dutch mainland coast south of Bergen aan Zee, and moderately lime-­ rich (0.5–1 %) on the primary dunes on Texel, Schiermonnikoog, Borkum, Juist and Baltrum. Lime-poor sands (0.1–0.5 %) are found on the rest of the West and East Frisian islands and almost lime-free sands (< 0.1 %) in North Frisia. The grain size

66

2  Sand Dunes and Their Vegetation Series

also varies with region: compared to the dunes of the North Frisian islands Amrum and Sylt that are dominated by grain sizes of 0.2–0.5 mm, the sand of the East and West Frisian islands is finer, with a maximum diameter of 0.1–0.2 mm (Heykena 1965). Finer sands have a higher water storage capacity and usually higher levels of plant-available potassium, magnesium and phosphorus (Wright 1956). If humans do not disturb the natural coastal formations, then the west side of the islands will experience erosion and the formation of dune scarps. The sand from the west side feeds the growth of the east side of the island (Flemming and Davis 1994). The majority of islands on the southern North Sea coast are therefore moving slowly eastwards (see Fig. 2.3). The East Frisian island of Langeoog has moved 2 km in 2000 years. Some islands have disappeared completely over the last century, such as the island Jordsand near Sylt, whilst others have visibly grown, such as Mellum in the Jade estuary and Norderoogsand in North Frisia (see Sect. 1.6.4). 2.2.1.2  Dune Creation Sunshine and wind cause the top layer of sand to quickly dry out, allowing it to be blown away by the wind. On stormy summer days, it can be stripped down to the edge of the capillary zone of the salty groundwater. Usually, the dry sand slides in a thin layer over the corrugated surface of the beach. Even small barriers on the beach promote the deposition of sand in their wind shadows (shadow dunes, see Fig. 2.4a). If the wind is blowing parallel to the coast, then barchan (or transverse) dunes can form, i.e. bare, sickle-shaped dunes such as are characteristic for the desert. Van Dieren (1934) observed them on the Dutch coast as small, unstable dunes up to 1 m high that are only vaguely similar to the massive shifting dunes of the Sahara or the Curonian Spit in the Baltic (see also Arens 1994). The sand only becomes stable in the wind shadow of larger barriers, such as behind washed up flotsam, or behind any plant that is able to colonise here. However, these embryo dunes are rarely higher than a few centimetres. The formation of permanent dunes in Central Europe is linked to the grasses Elymus farctus (sand couchgrass), Leymus arenarius (sea lyme grass), Ammophila arenaria (marram grass) and × Calammophila baltica (hybrid marram), which is why it is described as biogenic dune formation. The effectiveness of the grasses as windbreaks is shown by the measurements of Lux (1964) in primary dunes on Sylt (see Table 2.1). The wind speed on the crest of the dune is reduced by two-thirds between Ammophila tussocks, and by four-fifths in the dense stands on the lee side of the dune. These dune pioneers are typical psammophytes, i.e. plants adapted to moving sand and drought. They can colonise beach areas wherever the deposited sand is mixed with a little organic material and the soil moisture is sufficient, e.g. after flooding or rain. These conditions are often found in the deposited organic material at the drift line left by high winter storm tides. This nutrient-rich substrate is not only the basis for the colonisation of short-lived drift line communities (see Sect. 2.3.1.1), but also ­provides good conditions for the germination of the perennial grasses that lead to the formation of primary dunes (Heykena 1965).

2.2  Dune Formation and Destruction

67

Fig. 2.3  The alternation between dune formation and destruction on the North Sea island of Juist between 1650 and 1971 (according to a local historical map). The submerged churches show the south- and eastwards movement of the island. The coastal protection measures carried out in the twentieth century increased and stabilised not only the dunes but also the rest of the island exposed during low tide (LT). A freshwater lake protected by a dyke now marks the place where the island was once divided in two by seawater. Marshes formed only in areas protected from the sea, and in some cases have since been covered by dunes

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Fig. 2.4  A diagram of dune development and the formation of parabolic dunes on a coastline at right angles to the prevailing wind (From Paul 1944/1953) Individual primary dunes that form on the beach (A) combine (B). Changing wind directions and storm tides smooth out irregularities, so that a continuous foredune ridge forms (C). Blowouts (D) occur here that enlarge and drift inland in the form of parabolic dunes (E) Table 2.1  Relative wind speed on a beach and on white dunes of various expositions

Measuring point 50 cm above soil surface 10 cm above soil surface

Exposed shore 100

Marram grass stands (Ammophiletum) On leeward slope On windward In dense slope On crest Exposed vegetation 103 81 69 75

On leeward dune foot 73

85

60

50

33

42

18

From measurements by Lux (1964) on Sylt, in % of the highest average on the flat beach

Many different plant communities are involved in the formation and development of dunes. If the beach is undisturbed by human activity, they arrange themselves in a very characteristic sequence of zones. This zonation is mostly caused by succession, and is generally the same on both the North and Baltic Sea coasts, as the differences in tides and salinity of the seawater do not have much of an influence on the dunes. As shown in Fig. 2.5, the bare surface of the beach face usually contains patches of drift line material supporting nitrophytic annuals. Further landwards, Elymus farctus and Leymus arenarius form low primary dunes. These are then colonised by Ammophila arenaria, transforming into mobile (or white) dunes with steep summits, which combine into a dune ridge reaching heights of over 10 m. If the coastline remains stable and is not eroded, then the constant supply of sand will

2.2  Dune Formation and Destruction

69

Fig. 2.5  Vegetation succession and soil formation in coastal dunes along the North Sea coast Top: From Ellenberg, Vegetation Mitteleuropas, 4th edition: Without the influence of humans and livestock, at least the dune slack (9) and the dune heath (10, 11, with Calluna), but probably also the whole grey dune complex (7), would be covered with forest (12, 13) Middle: The same sequence on the island of Spiekeroog (From Gerlach 1993) (1) Sandy beach, (2) Honckenyo-Elymetum farcti (drift line), (3) Elymo-Ammophiletum typicum (primary dune), (4) E.-A. festucetosum arenariae (white dune); (5) the Hippophae form of (4), (6) Tortulo-Phleetum (grey dune), (7) Hieracio-Empetretum polypodietosum, Rubus caesius form (in parts grey dune with Empetrum scrub, the beginnings of podzolisation), (8) Hippophaeo-­ Sambucetum nigrae (scrub community), (9) Salix repens subsp. argentea community (creeping willow dune slack with a podzolised gley soil), (10) like (7), typical form (old dune heath with podzolised Cambisols), (11) Violo-Corynephoretum (on newly mobile acid grey dunes), (12, 13) Carici arenariae-Betuletum pubescentis (dune birch swamp forest with Alnus glutinosa) Bottom: Net nitrogen mineralisation in the upper soil (0–30 cm) from March to November 1987 and nitrogen pools that accumulated in the soil in the course of succession (From Gerlach 1993)

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allow these initial white dunes to grow into tall ridges, stabilised by Ammophila or by planting by humans. Further landwards where the sand is largely stable, humus-rich grey dunes form. Their vegetation is dominated by low-growing grasses and many forb species, as well as bryophytes and lichens. These are followed by older dune heaths with low woody vegetation (‘brown dunes’). The damper north or lee sides of the white dunes are often covered with scrub, which also grow out of the slack up to the summit of the dune. The North Sea coastal dunes are today only forested in a few areas, such as the western Netherlands, on some West and East Frisian islands and in small areas of the North Frisian islands (e.g. the south of Sylt). The higher the white dunes grow, the greater their exposure to storm winds. Here and there, trough- or saucer-shaped blowouts form that can quickly extend (see Fig. 2.4d, e). The trailing lobes become re-stabilised with vegetation, but the sand at the nose of the blowout migrates too quickly to allow plants to establish. The prevailing westerly and southwesterly winds push the blowouts into long troughs with parabolic walls. Groups of these parabolic dunes of different ages and sizes can migrate together through the dune field, combining with or destroying each other as they move (see Fig. 2.4e). Further inland, these dunes can grow to heights of 30 m or more, and can be much taller than the initial white dune ridge at the edge of the beach (Fig. 2.6). Under natural conditions, the lobes of the parabolic dunes remain covered in vegetation, which sooner or later spreads to stabilise the entire dune when it stops moving. The plant communities of the parabolic dune fields are more similar to those of grey dunes, or other older stages, than to those of the white dunes. Naturally, these would eventually be covered by forest. The blowout path of a parabolic dune develops into a secondary dune slack that, in contrast to the primary dune slacks parallel to the coast, runs in the direction of the prevailing winds. There are many exceptions to the normal sequence of dune formations and their vegetation, and often one or even several stages do not occur. Dune series are frequently incomplete on abrasion coasts, such as the west side of the East Frisian islands and in many places on the German and Polish Baltic coast (Piotrowska 1988). White dunes, grey dunes and even forested dunes are found here close to the water’s edge, as the current erosion rates are higher than they were when the dunes were formed. Dune landscapes become more variable through the formation of wide transverse valleys produced by the incursion of seawater or deep blowouts. The sand can be stripped from these slacks down to the capillary zone of the groundwater. If this occurs during a dry summer, i.e. during low groundwater levels, then the depressions will later periodically fill with water, producing mires. Larger incursions are known as washovers. These natural disturbances produce a heterogeneous mosaic of vegetation in dune pools, dune slack mires and sand pockets. Whilst dunes are produced by the sculpting power of the wind, berm ridges are the product of the breaking waves, particularly during storm surges which deposit gravel and sand parallel to the water’s edge in long banks (see Fig. 2.14).

2.2  Dune Formation and Destruction

71

Fig. 2.6  Leeward slope of a large mobile dune on the Curonian Spit in 2009 (Kaliningrad, Russia). The drifting sand blankets the pine and alder forests at the base of the dune and moves into the lagoon. The Scots pine plantation was established to halt dune movement. In recent years the dunes have been allowed to move freely within the National Park. The grey vegetation belt in the middle of the photo is mostly Festuca polesiaca grassland in immobile grey dunes

2.2.2  Coastal Dunes of the Baltic Sea The majority of the German and Polish Baltic coast is being eroded. Most of the coastal dunes have developed in areas with sand accumulation and the formation of spits. In the western Baltic (Schleswig-Holstein and the west of Mecklenburg-­ Vorpommern), the dunes are mostly less than 1  m in height, because the winds mainly blow seawards from inland and therefore do not transport much sand. Further east, in Vorpommern (northeastern Germany) and particularly in eastern Poland, white dunes can reach much larger sizes, such as those on Zingst and the 13 different spits between Wolin and the Vistula Lagoon (Piotrowska 1988). This is particularly the case for the once large mobile dunes, caused by human vegetation clearance destabilising the sand. Large-scale vegetation destruction can cause the sand to form huge mobile dunes free of vegetation, as for example in the Łeba Spit in the Polish Słowiński National Park and on the Curonian Spit in Kaliningrad and Lithuania (see Fig.  2.6). Many parabolic dune fields were probably also formed partly due to human activity: during the sixteenth and seventeenth century, as the

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overuse of the coastal dunes reached its peak; these became a real problem in the Netherlands and Denmark (Warming 1907). Many areas of the Baltic coast have multiple rows of similar shaped low walls of sand, on which the vegetation suggests increasing age. These berm ridges are evidence for the progressive movement of the beach, either due to a raising of the land from tectonic activity or from the deposition of large amounts of sand, e.g. on the Darß peninsula in Mecklenburg-Vorpommern (Fukarek 1961).

2.3  Vegetation 2.3.1  Coastal Dunes of the North Sea The heterogeneity in environmental conditions in the dunes and on the beach supports a large variety of plant communities over a relatively small area. We will restrict ourselves in the discussion of these communities to the major stages of natural dune formation and their later development. Detailed overviews of the dune vegetation on the Central European North Sea coast are given by Doing (1983), Westhoff (1991), Westhoff and van Oosten (1991), on the Dutch coast by Schaminee et  al. (1998), and on the German coast by Klement (1953), Heykena (1965), Wiemann and Domke (1967), Pedersen (1983), Preising et  al. (1990b), Hobohm (1993), Pott (1995), Fromke (1996), Biermann (1999) and Petersen (2000). 2.3.1.1  The Drift Line Especially the winter surges deposit small walls of sand and organic material, which in some years can form irregular belts at the high-tide mark on the beach. The remains of algae and eelgrass, mixed with seeds and bits of rhizome, form a short-­ lived but fertile substrate for communities of annual drift line species. These stands are very species-poor, with usually fewer than six species in a stand (Lee and Ignaciuk 1985). These drift lines are usually destroyed again during the winter storms, and in the following summer new stands will form from deposited seeds in another part of the beach. Tansley (1939) referred to these drift line communities as ‘migrating permanent communities’. These unstable habitats are usually colonised on North Sea beaches by the Cakiletum maritimae, Atriplicetum littoralis or Atriplex prostrata community (see Fig.  2.8). These are often accompanied by Honckenya peploides, Salsola kali and Elymus repens. Since the 1930s, Crambe maritima (sea kale) has also joined these communities, having arrived from other coastal areas of Europe and spread in the Dutch Wadden Sea. The increasing frequency of Atriplex is certainly due to the eutrophication of the North Sea and Wadden Sea (Dierßen 1996a). Tüxen (1950a) described the vegetation class as the Cakiletea maritimae after the Brassicaceae Cakile maritima (European sea rocket), which makes up large areas of coastal drift line.

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73

Seawater delivers the seeds of a wide range of nitrophytic plants, which mainly belong to inland ruderal vegetation or even be neophytes (e.g. sunflower and tomato plants). Some garden weeds originate from drift line communities, such as Senecio vulgaris and Tripleurospermum perforatum. Old drift line communities are ­sometimes colonised by perennial species of salt marshes or grassland, if they are not destroyed by storm surges (Heykena 1965). On the North Sea coast, this often occurs with Elymus species. 2.3.1.2  Primary Dunes and White Dunes The most important pioneer of dune formation is the salt-tolerant Elymus farctus (sand couch-grass), in the wind shadow of which primary dunes up to 1 m in height can form. These stands contain not only E. farctus itself, but also its hybrids with Elymus athericus (Elymus × obtusiusculus) and Elymus repens (E. × laxus, von Glahn 1987). The Agropyretum juncei (named after an earlier synonym for Elymus farctus) is a species-poor pioneer community often inundated by storm surges (see Table 2.2). At the bottom of dunes, it is frequently joined by the slightly succulent, nitrophilic Honckenya peploides, which can form dense carpets (Honckenya peploides community). In areas that contain more organic material, dense tussocks of the nitrophilic grass Leymus (syn. Elymus) arenarius form, which is much more effective at collecting sand than the thin tufts of Elymus farctus. Leymus arenarius also grows well on low foredunes and in semi-ruderal habitats (Fig. 2.7). Other species can also colonise sandy drift lines and promote the formation of low primary dunes, such as Cakile maritima, Salsola kali, Puccinellia maritima and Agrostis stolonifera (Heykena 1965). The Elymo-Ammophiletum is characteristic for the mobile white dunes, the name of which derives from the lack of carbon to darken the white sand (also sometimes called yellow dunes, see Table 2.2). The nutrient-demanding and salt-tolerant Leymus arenarius is characteristic for the transitional area between primary and mobile white dunes, which is at an early stage of soil formation. Ammophila arenaria, in contrast, dominates the more stable white dunes (see Fig. 2.9). It is often accompanied by a hybrid between A. arenaria and Calamagrostis epigejos (× Calammophila baltica). This is frequently planted to improve coastal protection due to its rapid growth. Further characteristic species for this community are Eryngium maritimum and Sonchus arvensis subsp. uliginosus. Despite the high nitrogen supply in the white dunes (see Sect. 2.5.1), the Elymo-Ammophiletum is typically species-poor as the sand is too unstable for most species, and it is only towards the secondary grey dunes that the species density noticeably increases. The plant communities of the mobile primary and white dunes are assigned to two phytosociological classes. In Central Europe, the class of the Ammophiletea contains a single order Ammophiletalia with the alliance Ammophilion. This contains the dune communities of the Agropyretum juncei and the Elymo-­ Ammophiletum. The more northerly class of the Honckenyo-Elymetea with the

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Fig. 2.7  A primary dune on the North Sea coast near St. Peter (Schleswig-­ Holstein) with Leymus arenarius

Fig. 2.8  The species of the Atriplicetum littoralis (driftline community) only root in the well-­ aerated and nitrogen-rich topsoil, i.e. up to around 8 cm depth (From Fukarek 1961) From left to right: Cakile maritima, Chenopodium glaucum, Atriplex hastata var. oppositifolia (centre), Salsola kali, Atriplex littoralis. The growth can be much denser or much sparser than is depicted here

order Honckenyo-Elymetalia and the alliance Honckenyo-Elymion (=  Agropyro-Rumicion Nordh. 1940) contains the Leymetum arenariae, the Crambeetum maritimae and the Honckenya peploides community. However, many authors do not separate these two classes (e.g. Pott 1995), as both contain communities that are ecologically very similar in Central Europe.

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75

Table 2.2  The vegetation of primary (P), white (W) and grey dunes (Gr) of the East Frisian islands Dune formation: Community no.: Graminoids: Cp Elymus farctus A Elymus arenarius Cw Calammophila baltica Cw Ammophila arenaria Festuca rubra subsp. arenaria Corynephorus canescens Carex arenaria F Phleum arenarium G Koeleria glauca Luzula campestris Aira praecox Other vascular plants: Cp Honckenya peploides Sonchus arvensis subsp. uliginosus Cw Eryngium maritimum Cw Hierac. umbellatum fo. linariifol. A Oenothera parviflora Cw Lathyrus maritimus Hypochoeris radicata F Jasione montana Viola canina var. Cg Hierac. umb. fo. armeriaefolia Silene otites F Myosotis stricta F Saxifraga tridactylites G Vicia lathyroides Erophila verna Galium mollugo Cg Viola tricolor subsp. curtisii F Cerastium semidecandrum Cg Lotus corniculatus fo. crassifol. F Sedum acre F Trifolium arvense Rumex acetosella 4 Galium verum fo. litorale Rubus caesius Bryophytes and lichens: Tortula ruralis subsp. ruralis Coelocaulon aculeatum

P 1

W 2

5 3

1 3 3 5

3

4 2 1

1 3 3 2 2

3

3 1 5 5 4 2

3 1 3 3 3 3 3 2

Gr 4

5

EIV M R

S

4 4 4 5 2 5 5 4

6 6 4 4 4 2 3 3 3 4 2

7 7 5 7 5 3 2 7 8 3 2

7 1 1 1 1 0 1 1 0 0 0

4 4 3 4 3 3 2 2 4 5 5 4 5 3 2

2 3 4 2 1 2 1 1 2 3 4 4 5 3 2 3 3 3

6 5 4 4 3 4 5 3 4 4 2 3 2 2 3 5 3 3 4 2 3 4 4 5

7 7 7 4 7 7 4 3 3 4 7 6 7 3 5 7 6 6 7 5 2 2 7 8

5 3 2 2 1 1 1 0 0 0 0 0 0 0 0 0 1 0 0 1 0 0 0 0

5 3

2 2

2 3

6 5

– –

5 5 5 5 5 5 2

(continued)

2  Sand Dunes and Their Vegetation Series

76 Table 2.2 (continued) Dune formation: Community no.: Peltigera rufescens Brachythecium albicans F Cladonia furcata Cladonia fimbriata

P 1

W 2

3

Gr 4 2 2

5 3 4 3 3

EIV M 3 2 5 5

R 8 5 4 4

S – – – –

From data in Tüxen (1937) and others Ecological indicator values from Ellenberg et al. (1992) Cp, Cw, Cg local character species of the primary, white and grey dunes, A order character species of the Ammophiletalia, G alliance character species of the Koelerion glaucae, F order ­character species of the sandy dry grasslands. Some species with low constancy values have been omitted

No. 1: No. 2 and 3: 2: 3: No. 4 and 5: 4: 5:

Primary dunes: wheat grass dunes (Atriplici-Elymetum arenariae), White dunes: Marram grass dunes (Elymo-Ammophiletum), typical, i.e. relatively young (E.-A. typicum), red fescue-marram grass dune (E.-A. festucetosum arenariae). Grey dunes: grey hairgrass dunes (‘Corynephoretum maritimum’, now Violo-Corynephoretum) typical, i.e. still somewhat mobile sand (C. M. typicum), rich in cryptogams, immobile and humus-rich (‚subass. of Brachythecium albicans’).

mM 5.4

mR mS 6.6 3.9

4.3 3.9

6.6 5.2

1.6 1.0

3.1 3.0

5.1 5.0

0.5 0.4

No. 1 is influenced by the groundwater (mM 5.4); the higher dunes are drier and colonised by xerotolerant species (mM 4.3–3.0). The increasing loss of base cations is visible in the mR values of the plant communities. The salinity is low even in the primary dunes (mS 3.9) and decreases over the white dunes (mS 1.6–1.0) to almost zero in the grey dunes

2.3.1.3  Grey Dunes and Dune Heaths Grey Dunes  Directly behind the zone of maximum sand deposition, and often in the lee of the white dunes, the microclimate rapidly changes. The habitat changes gradually into grey dunes with decreasing sand deposition and increasing soil maturity, as visible in the changing colour of the sand from white to grey to black. The species of the Elymo-Ammophiletum do not disappear immediately, but are gradually replaced by the denser and species-richer Elymo-Ammophiletum festucetosum arenariae. This community is dominated by Festuca rubra subsp. arenaria and other low-growing graminoid species such as Corynephorus canescens and Carex arenaria, as well as mesic forbs such as Hieracium umbellatum, Lathyrus maritimus and Oenothera ammophila (see Table 2.2 and Figs 2.10 and 2.11). These species indicate the lower disturbance from wind and the movement of sand in the lee of the white dunes, and its structure resembles that of dry grasslands in the class Koelerio-Corynephoretea (see Sect. 7.3.2).

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77

Fig. 2.9  A diagram of the development of Ammophila dunes. Ammophila forms a new layer of roots every time it is covered with sand (From Paul 1944/1953)

Fig. 2.10  Grey dune with Violo-Corynephoretum on the East Frisian island Spiekeroog with Empetrum scrub (bush dune) in the foreground. The dominant grass species are Corynephorus canescens (yellow tussocks) and Festuca rubra subsp. arenaria

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Fig. 2.11  A cross-section through a grey dune community on Darß (Elymo-Ammophiletum festucetosum) (German Baltic Sea coast) (From Fukarek 1961) From left to right: Calamagrostis epigejos, Ammocalamagrostis baltica, Hieracium umbellatum var. stenophyllum, Festuca rubra subsp. arenaria, A. b., Carex arenaria, Galium verum, A. b., Viola tricolor subsp. curtisii. The position of the rhizomes shows a previous period of sand deposition

The lower rates of sand deposition cause Ammophila arenaria to gradually weaken and stop flowering, allowing the grey dune alliance Koelerion albescentis (cf. Sect. 7.3.3.2) to take over. As long as the sand still contains a little lime or occasionally receives fresh sand from the beach, these open stands can support calcareous species such as Koeleria arenaria or Phleum arenarium, which depend on sparsely-­vegetated calcareous sands as they are easily outcompeted on acidic sediments. A characteristic community of the calcareous grey dunes is the TortuloPhleetum arenarii, which is often closely linked to the Elymo-Ammophiletum, and characterised by Phleum arenarium (sand timothy) and Tortula ruraliformis (star moss). All types of intermediate communities between the Koelerion albescentis and Corynephorion can be found in the coastal dunes of the West and East Frisian islands, but the lime-poor North Frisian islands do not support the TortuloPhleetum arenarii. Grey dunes contain a number of communities of nutrient-poor sandy grasslands (Festuco-Sedetalia acris), of which only the most important will be discussed here. These include the Festuco-Galietum veri, and the similar Agrostio-Poetum humilis typically containing Festuca filiformis (syn. F. tenuifolia), Poa pratensis, Agrostis capillaris and Galium verum (see Fig. 2.11). Two further (monodominant) communities within this order are the Carex arenaria community and the

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79

Fig. 2.12  A cross-section through a lichen-rich Corynephoretum canescentis cladonietosum on Darß (German Baltic Sea coast) (From Fukarek 1961) From left to right: Polytrichum piliferum, Corynephorus canescens, Cladonia arbuscula, C. c., Cladonia foliacea, P. p. Lichens can dominate older stands of the same community once sand deposition ceases. Note the deep and finely branched roots

Campylopus introflexus community, both of which will be discussed in more detail in Sects. 2.5.2 and 2.6.1. The dominant plant community of the grey dunes on the North Sea coast is the Violo dunensis-Corynephoretum canescentis, a sparsely vegetated pioneer community sometimes rich in mosses or lichens and usually found on south-facing slopes on nutrient-poor, acidic sands (see Table 2.2). This community is the oceanic equivalent of the Spergulo-Corynephoretum canescentis of inland dunes. In the western Wadden Sea, it develops under increasing decalcification from the Tortulo-Phleetum arenarii or, particularly in the lime-poor dunes of the northern Wadden Sea, from the Elymo-Ammophiletum festucetosum. The Violo-Corynephoretum colonises very lime-poor drifted sand and stabilises it after disturbances. It is therefore mainly found in the blowouts of older dunes, but also occurs in fragments wherever the sandy soil has been disturbed. In the pioneer form on open sand drifts, the stands are dominated by Corynephorus canescens and Carex arenaria (Violo-Corynephoretum typicum), whilst in its lichen-rich form it contains high densities of Cladonia, Cetraria and Coelocaulon species (Violo-Corynephoretum cladonietosum, Pott 1995; see Fig.  2.12). A detailed phytosociological description of the Violo-­

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2  Sand Dunes and Their Vegetation Series

Corynephoretum of the southern and eastern North Sea coast can be found in Biermann (1999). Instead of the fast-growing rhizome geophytes of the white and primary dunes, both the Tortulo-Phleetum arenarii and the Violo-Corynephoretum are dominated by hemicryptophytes that bind the top layer of sand (see Fig. 2.12). Deep-rooting species, especially of the Asteraceae, such as Leontodon saxatilis, also thrive alongside these shallow-rooted species. All species of these two communities are generally scleromorphic due mainly to the lack of nitrogen and of water. During dry periods, the soil of the south-facing slopes of grey dunes can dry to wilting point (Ernst 1991). The wind has a large influence even on stable dunes, especially as any damage to the sparse vegetation can cause the loose sand to become mobile again. If these gaps in the vegetation are frequently created, then they are usually filled by Corynephorus canescens. Cryptogams only establish in coastal dunes where they are protected by the grassy communities of the grey dunes. Neither lichens nor bryophytes can withstand being covered with sand faster than they can grow. However, the orthotropic (vertically growing) and denser mosses are better at growing through the sand than other moss species. It has been known for a long time that mosses can bind the sand with their rhizoids (Warming 1907). Younger grey dunes often support large colonies of star moss (Tortula ruralis, particularly the subsp. calcicolens; Koppe 1969; see Table 2.2), and more decalcified dunes usually contain Racomitrium canescens. Both play a large role in the formation of humus in the dunes. The lichen communities here are surprisingly species-rich; there are around 300 species in the dune ­habitats of Denmark alone (Alstrup in Biermann 1999). Lichens and bryophytes often form an almost continuous layer below the sparse phanerogams that is often quite striking, for example in the patchwork of white Cladonia alcicornis and black-­ brown Coelocaulon aculeatum thalli. The occurrence of several otherwise epiphytic lichens on the soil surface is also somewhat surprising, e.g. Platismatia glauca, Ramalina farinacea, Usnea subfloridana (Westhoff 1991), and in weakly acidic dunes also Hypogymnia physodes and Evernia prunastri. These cryptogam communities will be discussed in more detail in Sect. 2.3.4.2. Dune Heaths  Largely stable dunes are often colonised by Empetrum nigrum (crowberry), usually on the north-facing slope, marking the transition from grey to decalcified fixed dunes (or dune heaths) (see Fig. 2.10). These are known in German as brown dunes, due to the colouration of the soil by the accumulation of the iron(III) hydroxide Goethite (FeO(OH)), although this often does not occur in iron-poor dune sands. Empetrum tolerates being covered by sand (Lötschert 1968; Barendregt 1982) and according to Heykena (1965), even grows better under regular sand deposition. In contrast, it is quite sensitive to the exposure of its roots by erosion. Its berries are transported by birds, but it is mainly through vegetative reproduction that it creates large carpets with their own, humid microclimate. This dwarf shrub has a boreal-atlantic range and is relatively sensitive to drought, which is probably why it grows mainly on north-facing slopes in the West and East Frisian islands and thrives

2.3 Vegetation

81

in the shade of trees. In the cooler North Frisian islands, Empetrum is also found on south-facing slopes. Empetrum is more sensitive to trampling than the other Ericaceae species found in dunes, i.e. Calluna vulgaris and Erica tetralix (Hylgaard and Liddle 1981). Disturbance from grazing livestock, high densities of rabbits and fire have all caused the populations of Empetrum to decline on the North Sea islands over the last few centuries. The Hieracio-Empetretum (crowberry heaths, related to the Carici-Empetretum and Polypodio-Empetrum communities) are, together with the Violo-­ Corynephoretum, the major vegetation types of the grey dunes and dune heaths. Alongside the dominant Empetrum, constant species of this community include Carex arenaria, Salix repens and Polypodium vulgare. The Hieracio-Empetretum forms from the Violo-Corynephoretum via the establishment of young Empetrum plants or by the lateral spreading of existing communities. If the deposition of sand ceases, then Empetrum can be outcompeted by Calluna vulgaris (Heykena 1965). Stable sand with higher humus content supports the Genisto-Callunetum, which often develops from earlier Hieracio-Empetretum stands. 2.3.1.4  Dune Scrub and Wooded Dunes On the moderately calcareous West and East Frisian islands, the Elymo-­ Ammophiletum often directly borders on a sand-accumulating scrub community. This community of the Salici-Hippophaetum rhamnoides (alliance Salicion arenariae, see Chap. 12 in Volume I: no. 6.6.1.1) mainly colonises the lee sides of the white dune ridges (see the descriptive vegetation maps in Tüxen 1956c; Wiemann and Domke 1967; Petersen et al. 2003). In these sheltered spots, Hippophae rhamnoides (sea buckthorn) can reach heights of up to 2 m, whilst Salix repens subsp. dunensis (syn. S. arenaria) rarely grows above knee height. Salix repens is, however, more resistant to wind and sand deposition than Hippophae, and grows even on the highest dune summits, which it also helps to form through its stabilising influence. Its branches often show signs of wind shear, and can die off in dry years, especially as the wind contains salt when it blows off the sea. As shown by Paul (1953) on the Curonian Spit, Salix repens tends to germinate most frequently on damp, bare sand. In low dune fields, these conditions are only found in the freshly created dune slacks. The colonisation of Salix is also promoted at the foot of larger dunes, where groundwater can build up from the seepage of rainwater. Juncus balticus, which Braun-Blanquet and De Leuuw (1936) mention as frequently co-occurring with Salix repens scrub, also establishes mainly in such areas of water-saturated mobile sand on the Curonian Spit. As Salix repens can easily grow through deposited sand, it can form dunes of up to 10 m in height, making it hard to imagine that it originally established close to the groundwater. This would support the view of Paul (1953) that the Salici-Hippophaetum rhamnoides is not a stage in normal dune succession (in contrast to Braun-Blanquet and De Leeuw 1936).

82

2  Sand Dunes and Their Vegetation Series

Young Salici-Hippophaetum communities develop on sands that are still somewhat calcareous, at least below the surface. Its herbaceous species are similar to those of the grey dunes. The stable sand is, however, quickly decalcified and the semi-shade of the scrub is colonised by acidophytic forest mosses such as Dicranum scoparium, Hylocomium splendens, and particularly Hypnum cupressiforme as well as recently also Campylopus introflexus. Larger acid-tolerant plants also establish, such as Polypodium vulgare, Pyrola rotundifolia subsp. maritima, Veronica officinalis and Hieracium umbellatum, leading finally to the development of the Polypodio-Salicetum arenariae with forest-like undergrowth (see Fig.  2.16). Epilobium angustifolium and Sambucus nigra grow through ageing sea-buckthorn bushes to form the Hippophae-Sambucetum nigrae (Pott 1995), which is rich in nitrophytes such as Rubus caesius, Urtica dioica, Galium aparine and Solanum dulcamara. Several dry, south-facing slopes of grey dunes and dune heaths on the islands of the North Sea are colonised by the grazing-sensitive Roso pimpinellifoliae-­ Salicetum arenariae, with the white-flowering Rosa pimpinellifolia (burnet rose) as a pioneer species of dune scrub. In damp dune slacks, species of Pyrola (wintergreen) join Salix repens in the Pyrolo-Salicetum repentis (see Fig. 2.5). The presence of mosses such as Calliergonella cuspidata, Drepanocladus species and indicators of damp conditions such as Hydrocotyle show that the groundwater is only 45–50 cm below and sometimes even closer to the surface of the humus-rich to peaty sand. These scrub communities merge with the communities of dune slacks (see Sect. 2.3.1.5). As the East Frisian islands have migrated relatively quickly from west to east over recent centuries, their dunes are relatively young (Tüxen 1956c; Gessner 1957; Streif 1990; see Fig. 2.3). In addition, these islands were always relatively densely populated for their size, so that the development from dune scrub to true forest was only very rarely able to occur. Dutch dunes, in contrast, support numerous communities containing scrub and tree species (Westhoff 1952; van der Maarel et  al. 1985; Westhoff and van Oosten 1991). Alongside the previously mentioned Hippophae and Sambucus, the scrub communities here also contain Crataegus monogyna and Ligustrum vulgare (Pott et al. 1999). The progression of dune scrub and forests is especially visible in the calcareous dunes of the western Netherlands and on some East Frisian islands (van der Maarel et  al. 1985; Isermann and Cordes 1992). The rapid succession is certainly not just the result of the lack of livestock grazing, but also of the spread of N-fixing Hippophae and atmospheric N deposition. The pioneer tree species in dune ecosystems are often Betula pubescens and B. pendula, Populus tremula, and in some places also Sorbus aucuparia. Quercus petraea and particularly Q. robur follow later, although Q. robur can also act as a pioneer tree species (Lawesson and Wind 2002). The succession on the West and East Frisian islands tends to develop first into grassy pioneer scrub and pioneer birch and birch-oak forests (Betula pubescens community, Empetro-Betuletum carpaticae, Quercus robur community; Pott et al. 1999), from which many of the char-

2.3 Vegetation

83

acter species of the Quercion roboris are missing. According to Lawesson and Wind (2002), Scots pine is also naturally present in the Atlantic dune forests of the North Sea. In the damp dune slacks, tall-growing willow scrub of Salix cinerea with S. repens establishes, as well as black alder swamp forests if the ground is nutrient-­ rich enough (Stortelder et al. 1999). It was previously doubted that dunes could naturally become wooded, and indeed the dunes of the North Frisian islands are still almost treeless today (Neuhaus 1990). Nevertheless, after numerous successful afforestations and the spread of neophyte tree species (e.g. Prunus serotina in North Frisia), it is now rarely disputed that tree growth is possible despite the strong winds. Even on the Norwegian skerrys, birch forms small woods close to the wind-whipped, salt-sprayed shoreline (Wassen 1965). Natural afforestation often begins in the damp depressions and slacks of the coastal dunes. The colonisation by trees on the dry dunes of the North Sea is probably mainly limited by the lack of seeds, as well as the threat of being covered by sand and the lack of water. Also browsing by deer and rabbits can play a role, as some islands have high densities of these animals. 2.3.1.5  Dune Slacks If the slacks between the lobes of the dunes are eroded down to the water table, then a mosaic of vegetation types form mainly consisting of phreatophytes, i.e. deep-­ rooted plants that are mostly or entirely dependent on groundwater. Dune slack communities are often rich in vascular plants and bryophytes and contain numerous rare and endangered species (Grootjans et al. 1995). Noest et al. (1989) found 346 species (including bryophytes) in the dune slacks of Voorne in Rhine estuary, and densities of over 40 species per m2 have been counted (van der Maarel and Leertouwer 1967). The high species richness is partly caused by the repeated disturbance by inundation (‘intermediate disturbance hypothesis’, Connell 1978), and partly by the small-scale variation in water supply, producing many ecological niches within a small area (Lammerts et al. 1995). Boorman et al. (1997) list around 60 higher plants that are characteristic for Northwestern European dune slacks, including many small Cyperaceae and Juncaceae (see Fig.  2.13). Only some of these are obligate phreatophytes (see Sect. 2.4.4), whilst the majority are facultative phreatophytes or belong to the neighbouring dry dunes. Most species of these communities are low-growing and fairly uncompetitive. The type of plant community found in the dune slacks is mainly determined by the level of the groundwater and its seasonal fluctuations. Dune slacks close to the beach often contain brackish water, whilst those further away are influenced by the freshwater that seeps from the dunes. The annual variation in precipitation leads to the periodic drying out of the smallest pools, and causes fluctuations in the species composition (van der Laan 1979). Further influential factors are the age of the substrate in the slack, its lime content and the nutrient content of the soil. Although floating plants, reed and scrub communities can also play a role here, we will focus

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2  Sand Dunes and Their Vegetation Series

Fig. 2.13  Dune slack community with Radiola linoides, Carex oederi (in the centre), Potentilla anserina and dwarf Phragmites australis in the Listland of Sylt (North Frisia, Germany)

particularly on the low-growing aquatic communities of the class IsoëtoNanojuncetea and Littorelletea, as they are so characteristic for damp dune slacks. Pioneer communities on bare, calcareous muddy soils in brackish to freshwater dune slacks include the Samolo-Littorelletum, with the helophytes Ranunculus flammula, Littorella uniflora, Mentha aquatica, Carex oederi, Hydrocotyle vulgaris and Juncus anceps. Periodically wet, acidic soils also support the very rare Cicendietum filiformis with Cicendia filiformis, Radiola linoides, Anagallis minima and Juncus pygmaeus (Petersen 1999). The latter community forms annual low-­ growing stands on nutrient- and humus-poor sandy and clayey soils, sometimes accompanied by Isolepis setacea and Sagina procumbens or with Juncus bulbosus stands. The swamp plants are often interspersed with species of drier habitats, such as Carex arenaria or Leontodon saxatilis. These pioneer communities quickly disappear in the absence of regular disturbance, as they are simply overgrown by more productive vegetation. Regular removal of the vegetation and top layer of soil is one of the few ways to protect this endangered community. If disturbance ceases, then the sparse pioneer vegetation of muddy soils can develop into communities of the Scheuchzerio-Caricetea nigrae (Westhoff 1991). In lime-poor, acid dune slack swamps, this is the Caricetum trinervi-nigrae, which, apart from the eponymous Carex trinervis also contains species such as Carex nigra, Calliergon, Drepanocladus and Sphagnum species and Eriophorum

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85

angustifolium, as well as Salix repens. In base-rich dune slacks, the vegetation develops towards calcareous mire communities that may contain the rare, subatlantic-­prealpine orchid Liparis loeselii. The Junco baltici-Schoenetum nigricantis community is often exposed to brackish water if the water levels fluctuate considerably. Apart from the tussocks of Schoenus nigricans, this community contains numerous rare species such as Dactylorhiza incarnata, Epipactis palustris, Pedicularis palustris, Eleocharis quinqueflora, Juncus balticus and characteristic mosses (e.g. Pellia endiviifolia, Aneura pinguis, Bryum pseudotriquetrum, Westhoff 1991). A lowering of the water table and/or acidification causes both communities to develop into damp Empetro-Ericetum heaths or taller Salix repens and S. cinerea scrub. Empetrum, Erica and Salix form dense stands of the Empetro-Ericetum, which sometimes also contain Drosera rotundifolia and Lycopodiella inundata if there are patches of open ground (Neuhaus 1990). With average groundwater levels of 30 cm below the surface, these communities are often flooded in winter (Petersen et al. 2001). The vegetation of brackish and freshwater dune pools of the western Wadden Sea are described in more detail by Westhoff (1991) and Westhoff and van Oosten (1991). 2.3.1.6  Coastal Cliffs In Central Europe, coastal cliffs only occur in the Triassic Buntsandstein formation of the North Sea island of Helgoland and the chalk cliffs of Rügen (Mecklenburg-­ Vorpommern) and Møn (southern Denmark). Particularly characteristic for the Helgoland cliffs is the Cardario drabae-Brassicetum oleraceae community with Brassica oleracea subsp. oleracea, the wild form of the domestic cabbage, and the neophytic Lepidium draba (whitetop). This community is found on coarse rocky debris with little fine earth, and is a permanent pioneer community containing a mixture of ruderal plants and halotolerant coastal taxa (Dierschke and Walbrun 1986; Walbrun 1988).

2.3.2  Baltic Coastal Dunes Along the German and Polish Baltic coast, the landscape alternates between abrasion coasts with scarps and areas of accretion with sandbanks, berms and spits with dune formation. The vegetation of the drift line, primary dunes and white dunes is basically similar to that of the North Sea dunes (see Fig. 2.16). However, like in the salt marshes of the Baltic coast, several typical salt-tolerant species gradually disappear towards the east. The Atriplicetum littoralis and the Agropyretum juncei, for example, are only present in fragments on the Polish Łeba Spit (Wojterski 1978). The Crambe maritima community occurs particularly on shingle beaches (Eigner 1973, Fig. 2.14). The low salinity on the Baltic coast means that Ammophila, and

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2  Sand Dunes and Their Vegetation Series

Fig. 2.14  Shore on the southern Danish island Aeroe in the Baltic Sea with a berm ridge of gravel which is colonised by Crambe maritima (on the crest), Rumex obtusifolius (in the foreground), Beta vulgaris, Sonchus arvensis and several ruderal species

further east also Leymus arenarius, play an important role in dune formation (see Fig. 2.7). × Calammophila baltica is also more important in the mobile white dunes of the southern Baltic coast than on the North Sea coast, and this hybrid is twice as common in the Elymo-Ammophiletum than its parent species Ammophila. The dune communities also contain several subcontinental species such as Petasites spurius, Helichrysum arenarium, Artemisia campestris and Chondrilla juncea (Möller 1975; Isermann 1997). The grey dunes are often colonised by subatlantic communities such as the Violo-Corynephoretum, Festuco-Galietum veri, Thymo-Festucetum ovinae and Carex arenaria community (see Fig. 2.15). These are largely similar to those on the North Sea coast (see Table 2.3), but the basiphilic Tortulo-Phleetum is largely absent from the Baltic coast. In contrast to the North Sea, the Baltic grey dunes also support various sandy steppe communities with many subcontinental and continental species (Hundt 1985). The communities belonging to the subcontinental-­continental sand steppes (Festuco-Sedetalia acris), including the Helichryso-Jasionetum, are mainly found in the east and gradually disappear towards the west in the dunes of Mecklenburg-Vorpommern. Isermann (1997) lists the characteristic species of these flower-rich communities as Cardaminopsis arenosa, Chondrilla juncea, Erigeron acris, Artemisia campestris, Helichrysum arenarium and Festuca polesica. The latter dominates many eastern grey dunes in place of the suboceanic Corynephorus. In contrast to these markedly continental grey dune communities, the dwarf shrub-rich dune heaths of the Baltic coast have a more oceanic character.

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Fig. 2.15  Extended grey dune complex on the Curonian Spit (near Kaliningrad) with a large mobile dune on the horizon. Instead of Corynephorus, which colonises most dunes in oceanic western Central Europe, the open grassland in this subcontinental to continental region is dominated by Festuca polesica

Table 2.3 Plant communities of white (W) and grey dunes (Gr) of the Baltic coast of Mecklenburg-Vorpommern Dune formation Community no.: No. of relevés Grasses and grass-like plants Ammophila arenaria × Calammophila baltica Elymus × obtusiusculus Elymus farctus Leymus arenarius Festuca rubra ssp. arenaria Festuca rubra ssp. rubra Agrostis capillaris Corynephorus canescens Carex arenaria Anthoxanthum odoratum Bromus hordeaceus Aira praecox Holcus lanatus Deschampsia flexuosa Festuca ovina (agg.)

W 1 519

Gr 2 447

Gr 3 341

Gr 4 244

52 71 12 24 59 41 8

9

5

19

6 5 15 98 63 9 2 2 1 9 10

3 7 64 35 31 27 20 14 21 18 51

14 12 27 82 52 22 7 5 8 10 9

18 42 2

3

(continued)

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2  Sand Dunes and Their Vegetation Series

Table 2.3 (continued) Dune formation Community no.: No. of relevés Other vascular plants Lathyrus maritimus Cakile maritima baltica Salsola kali kali Honckenya peploides Lactuca tatarica Cerastium semidecandrum Sedum acre Trifolium arvense Hypochaeris radicata Artemisia campestris Rumex acetosella Jasione montana Helichrysum arenarium Teesdalia nudicaulis Plantago lanceolata Spergula morisonii Hieracium pilosella Achillea millefolium Ornithopus perpusillus Hieracium umbellatum Galium mollugo Conyza canadensis Bryophytes and lichens Ceratodon purpureus Polytrichum piliferum Dicranum scoparium Cephaloziella divaricata Cladonia aculeata Cladonia arbuscula Cladonia fimbriata Cladonia foliacea Cladonia furcata ssp. furcata Cladonia macilenta Cladonia pyxidata Cladonia scabriuscula Cladonia subulata Brachythecium albicans

W 1 519 11 13 10 14 13 7 2 12 12 9 5 2

3 24

13 4 1

4

2 1 4 6

Gr 2 447

Gr 3 341

Gr 4 244

8 2 2 24 4 28 18 2 14 2 39 12 2 4 10 4 4

27 16 31 54 21 62 26 16 14 31 5 48 50 14 16 24 12

37 23 23 55 39 40 73 53 18 9 4 41 18 16 30 21 18

31 46 20 28 32 33 10 28 20 29 28 6 24 5

26 28 20 6 8 6 8 5 11 1 7 6 6 16

49 30 22 23 12 15 28 12 24 15 32 21 22 24

From Berg et al. (2004) Species with low constancy were omitted No. 1: white dunes, Elymo-Ammophiletum arenariae, No. 2: grey dunes, Corniculario aculeatae-Corynephoretum, similar to Violo-Corynephoretum, No. 3: grey dunes, Thymo pulegioidis-Festucetum ovinae, No. 4: grey dunes, Helichryso arenarii-Jasionetum

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89

Fig. 2.16  A transect through a dune series with Empetrum-pine forests on the Polish Baltic coast (Modified from Wojterski 1964b) In contrast to the North Sea dunes (Fig. 2.5), the dune vegetation contains fewer species and is less variable. Pine quickly establishes and leads to the development of forest with a subcontinental character. Closer to the coast, more oceanic species (such as Erica tetralix and Myrica gale) still play a surprisingly large role. Calluna and Vaccinium species are also frequent in these open pine forests. The destruction of dune forests like these led to the formation of the tall mobile dunes of the Curonian and Łeba Spits (see Fig. 2.6)

As is the case on the North Sea coast, the Hieracio-Empetretum and the GenistoCallunetum dominate the majority of the area (see Fig. 2.16), although the latter disappears east of Gdansk with the increasingly continental climate (Steffen 1931: cf. Sect. 6.3.2.1.2). Dune forest pioneers on the Baltic coast are mainly Betula pendula and B. pubescens, Pinus sylvestris and Populus tremula, as well as Quercus robur and Sorbus aucuparia in places. Alnus glutinosa also occurs in damp dune slacks (Paul 1953; Wojterski 1964b; Hundt 1985; Isermann 1997). Pinus sylvestris has been planted in many places since the beginning of the nineteenth century, but it would probably also be naturally present in the dunes (Pyrolo-Pinetum, Empetro-Pinetum; Hundt 1985; Piotrowska 1988; see Fig. 2.16). Dune slacks are usually colonised by Betula pubescens (Paul 1953), whilst stunted, windshorn pines reach the ridges of seaward older dunes (see Fig. 2.6). East of the Oder estuary, pine forests are even present as the zonal vegetation on the sometimes very large areas of old dunes on the Polish coast (Wojterski 1964a, b), although many stands have developed from plantations. Usually, these dune pine forests are rich in Empetrum, which distinguishes them from the pine forests further inland. Further information on the vegetation of the dunes and beaches of the German and Polish Baltic coast can be found in the overviews of Raabe (1950) and Möller (1975) for Schleswig-Holstein, of Libbert (1940), Fröde (1950), Fukarek (1961), Passarge (1962), Passarge and Passarge (1973), Isermann (1996, 1997) and Berg

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et  al. (2001, 2004) for Mecklenburg-Vorpommern, and of Steffen (1931), Paul (1944, 1953), Piotrowska and Celinski (1965) and Piotrowska (1988) for Poland.

2.3.3  Mobile Dunes Without Vegetation 2.3.3.1  Development of Mobile Dunes During the Middle Ages and the early modern age, forest clearance and grazing spread in the dune landscapes of the North and Baltic Sea coasts to such an extent that it led to the development of desert-like mobile dunes. Large shifting dunes still exist in Kaliningrad and on the Łeba Spit in Poland (see Fig. 2.6), as well as on a much smaller scale e.g. on Sylt and in Jutland. The Curonian Spit, the largest area of mobile dunes in Europe, was studied in detail by Paul (1944, 1953). The chain of sand dunes on the Curonian Spit is 80 km long and 400–800 m wide, and has been unstable for the last 200–300 years. These bare dunes are on average 35  m high, reaching a maximum of 70 m, whilst vegetated dunes only exceed 25  m in exceptional cases. If the dunes are not stabilised, then they drift 3–12 m per year eastwards, leaving a trail of fairly flat sand several metres above sea level. These huge dunes shift via the transport of dry, bare sand from the west-facing 4–12° slopes to the east-facing slopes at angles of 30–35°. Contiguous forest covered the majority of the Curonian Spit until well into the seventeenth century (see Fig. 2.16). It grew on old parabolic dune fields, of which a few are still visible that haven’t been covered by the mobile dunes. Evidence of this forest can still be seen in the humus horizon on the windward side of the mobile dunes. These old forest soils are still visible even on the afforested slopes in the clearly better growth of pine. Usually, several forest soil layers are present in irregular sequences and separated by light-coloured sand layers up to several metres thick. They are witnesses to the historical changes in dune landscapes, and can e.g. be seen in the Polish Słowiński National Park and the Russian part of the Curonian Spit near Levkoje. By the first half of the nineteenth century, the Curonian Spit was just a huge field of sand. The mobile dunes swallowed villages and cropland, and even threatened to block up the harbour of Klaipėda (Memel) and fill the Curonian lagoon. In order to prevent the further movement of sand, the seaward edge was planted with Ammophila to create a continuous immobile foredune. Ammophila did not, however, spread, and neither did Leymus arenarius. Both remained mainly on the coast, i.e. in areas receiving regular sand deposition. It is only when larger amounts of sand are deposited that Ammophila also grows further inland as, for example, found by Paul (1944) in places on the ridges of mobile dunes. It is also well established in places on the lower but still impressive mobile dunes in the Słowiński National Park. Planted Ammophila generally does not grow well on the nutrient-poor sand, so that it only acts as a mechanical barrier to the movement of the sand.

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91

The low-growing grasses of the Corynephorus and Festuca psammophila communities that are found all over the trails left by the mobile dunes, do not, in contrast, tolerate large volumes of sand deposition, and were equally unable to stabilise the sides of the dunes. In order to finally restrain the drifting sand, it was therefore necessary to secure the windward dune slopes with brushwood hedges, with Scots pine or mountain pine planted in the holes. In the meantime, the majority of the dune area has been stabilised in this way, and even the trails have been almost completely forested. This reafforestation can be seen as further evidence that neither wind, blown sand nor salt spray make tree growth impossible on the coasts. 2.3.3.2  Vegetation Succession on Parallel Ridges Now under nature conservation management, some of the large dunes on the Baltic coast are still migrating eastwards (e.g. at a rate of 6 m per year on the Łeba Spit). This has caused a system of dune ridges to form on the western edge of the Curonian Spit. This 300–900 m wide zone is not affected by the mobile dunes, and contains around 20 parallel ridges of sand of 0.5–2 m in height. Their vegetation stabilises the sand that is deposited there by easterly winds from the large dunes. Paul (1944) was able to determine the ages of these ridges using various methods, so that they provide a unique opportunity to measure the speed of grey dune succession. The development of a Festuca rubra arenaria stage requires between 3 and 15 years, whilst the Violo-Corynephoretum takes around 25–30 years to reach its optimum. Algae often initially cover the surface of the sand as a green crust, whilst bryophytes and lichens are rarely pioneer species here. It is only after a few decades, when the sand is completely stable, that lichens can cover over 25 % of the area in places. Gaps in the vegetation caused by mechanical damage promote Carex arenaria with its long rhizomes during all stages of succession (see Sect. 2.6.1). Grazing or strong browsing by wild animals can lead to the formation of a Festuca ovina community. However, this species-rich community is only found on the oldest (i.e. over 75 years) ridges. Overall, succession therefore occurs very slowly on the nutrient-poor drifted sand. Under the higher nutrient conditions of the beach, in contrast, it only takes a few decades for an area to develop from bare sandy beach to white dune with an optimum Elymo-Ammophiletum community (according to the observations by Reinke 1903; Warming 1907; Tüxen and Böckelmann 1957 and others; see Sect. 2.5). Paul (1944) found almost no evidence of young woody forest pioneers on the ridges. However, it should be noted that pine seedlings mainly germinate in moss and lichen vegetation, which was rare here. Cryptogams can in fact be seen as early indicators of the development of a lichen-pine forest, which is a natural climax community on older dunes in the east of Central and Northern Europe (see Sect. 2.3.1.4). Similar ridges left by large dunes can also be found on the Łeba Spit and on Sylt, although much smaller and less clearly recognizable.

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2  Sand Dunes and Their Vegetation Series

2.3.4  Vegetation of Inland Dunes 2.3.4.1  Formation and Distribution of Inland Dunes Old coastal dunes further from the shore that have been cut off from the supply of sand from the sea should be treated as essentially inland dunes. In particular, parabolic dunes that have migrated several hundreds of metres inland can in some cases be of the same age and support very similar vegetation to the sand dunes further inland. Nevertheless, coastal and inland dunes differ in their floristic composition. Even the oldest decalcified coastal dunes that no longer receive new sand still contain a few species that occur exclusively in coastal areas. These are mainly the many species, subspecies, varieties and forms with the scientific names ‘maritima’, ‘litoralis’ and similar (several examples are given in Table 2.2). The seeds of these coastal plants are dispersed in such large numbers close to the coast that there are always a few individuals present, even when only a small proportion germinate. Only Ammophila is no longer an indicator of coastal areas, as it has been so frequently planted in inland dunes. It does not, however, grow well here, even in the calcareous sands of the Upper Rhine plain. The majority of inland dunes developed during the Weichselian or the early postglacial period, or at least the substrate comes from the period when the large talsand plains formed during and after the Saalian and Weichselian glaciations which were only sparsely vegetated. The sands formed in the glacial valleys of the northern plains were always lime- and nutrient-poor. Larger inland dune landscapes generally occur on the eastern and northeastern edges of the Pleistocene river valleys, e.g. along the Rhine (Campine and Veluwe) as well as along the Ems, Weser, Aller, Elbe, Oder, Noteć and Vistula (Pyritz 1972; see Fig. 2.17). Nature reserves that still have shifting sand can be found e.g. east of Warsaw. Inland blown sands come from the wide meltwater rivers, which once deposited sand during flooding events each summer. The prevailing westerly and southwesterly winds then caused the deposited coarse sand to blow across the dry terraces and spread over the neighbouring moraines, whilst the fine sand and silt was deposited further away as sandy loess. Many of these old sandy areas are largely flat, and have not grown into dune ridges with the help of stabilising vegetation. The most notable inland dunes found today mostly developed in the Middle Ages, when humans unintentionally destabilised the sand, mainly via deforestation and sheep grazing. Even in the early twentieth century, the formation of sand drifts on sheep drove-roads could still be observed in Emsland (northwestern Germany) and other heathland areas. There are still mobile sand dunes in the Lüneburg Heath near Wilsede (south of Hamburg), although on a very small scale. In the wide Upper Rhine plains, there are also large-scale dune fields, particularly in the central and northern parts near Heidelberg and between Mainz and Darmstadt (e.g. the ‘Mainz Sands’). In contrast to the dunes of the lowlands formed by the Saalian (penultimate) and Weichselian glaciations in northern Central

2.3 Vegetation

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Fig. 2.17  Inland dune at the edge of the Elbe valley in eastern Lower Saxony (Klein Schmöhlen, northern Germany). An open Spergulo-­ Corynephoretum grassland rich in lichens is visible in the foreground. During the last centuries, most inland dunes of Pleistocene origin have been planted with pine forests

Europe, the sands here are richer in lime, as the bedrock is more calcareous and has not been as thoroughly leached. This contrast is clearly visible in the vegetation, so that the two regions should be discussed separately. Compared to the coastal dunes of the East Frisian islands or even with the northwest German inland dunes, the dune vegetation of the Rhine plain is more grassland-like and species-rich. We will therefore leave the discussion of these communities until Sect. 7.3.3.2. The Morava plain near Vienna is also a dune-rich landscape with relatively young and calcareous sand like the Rhine plain, but containing more species of southeastern Europe (Fischer 2004). In contrast, the inland dune vegetation of northeast Germany is similar to that of the northwest, but lacks several atlantic species and is closer in character to the more nutrient-rich and continental dune communities such as have been recorded near Warsaw (Juraszek 1928). As it is not possible to go into all forms of inland dunes here, we will concentrate on those in northwest Germany. 2.3.4.2  Corynephorus Communities and Fruticose Lichen Layers Grey Dune Grasslands  The sands of the inland dunes in the Pleistocene lowlands are decalcified down to the subsoil and are moderately to strongly acidic. Their pH (H2O) values are between 3.2 and 4.6, i.e. in the same range as acid forests and

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2  Sand Dunes and Their Vegetation Series

heaths (Krieger 1937). These quartz-rich dune soils were never particularly rich in nitrogen or base cations, and their lack of silt and clay means that they are well drained. If unshaded, their surface will dry out very quickly and is easily eroded. At most, around 25 Central European phanerogam species are adapted to such a combination of unfavourable growth conditions. Still, the first colonisers of bare sand in these habitats are vascular plants, rather than lichens or bryophytes. The slow-­ growing cryptogams have few demands on the water and nutrient supply, but require stable substrates to develop, such as rocky outcrops, boulders, exposed mor humus or sand protected from the wind. They grow poorly or not at all in exposed areas where the wind removes or deposits large amounts of sand. Tüxen (1928) was one of the first to recognise this in his overview of succession on northwest German inland dunes. Blown sand that is not too strongly affected by the wind supports the germination of Corynephorus canescens, the seeds of which are anemochorous, if there has been a longer period of rain in late summer or spring and the sand is moist on the surface. The young, grey-green, brush-like tufts of Corynephorus collect the drifting sand into small hummocks, which they fill with tillers. At the same time, they send their roots in all directions down to depths of over 50 cm (see Fig. 2.11), providing both a good anchorage and water and nutrients from a large volume of soil. After 3 years of growth, Corynephorus communities can become so dense that they even support the growth of species sensitive to sand deposition. Corynephorus then begins to decline, not only due to the increasing competition, but also because it, like Ammophila, simply grows better under higher rates of sand deposition (see Sect. 2.4.3). The initial phase of development of the Spergulo-Corynephoretum in subatlantic inland areas supports several further phanerogams, namely the spring therophytes Spergula morisonii and Teesdalia nudicaulis. These small, shallow-rooting plants are quick to produce seeds and are mainly found in the submediterranean-atlantic coastal areas, giving the Corynephorion communities of inland northwest Germany an oceanic character (see Table 2.4). Perennial species in this community include Carex arenaria (sand sedge) and Rumex acetosella (sheep’s sorrel), the seeds of which are also wind-dispersed and profit from sand deposition for their establishment. Carex arenaria can establish at the edges of gaps in the vegetation, sending its rhizomes down to 50 cm below the surface. The substrate here rarely dries out but is still loose and well aerated, compared to the surface layers that often completely dry out and are hostile to the establishment of plants, especially in the warm summer months (Tidmarsh 1939). Only a few rosette plants survive in these areas, such as Hieracium pilosella. Carex arenaria also grows well in dune habitats close to rivers, as it is tolerant of flooding (Moog 1989). Corynephorus canescens is only found on loose, sandy, neutral to acidic soils. Paul and Richard (1968) found that it would grow without competition on calcareous sand or even clay, but even in culture it grows best on acid sands. Rumex acetosella can also grow well on both acid and calcareous soils, and is only forced onto nutrient-poor acid sands by competition in more fertile habitats. In New Zealand,

2.3 Vegetation

95

Table 2.4  Corynephorus communities and fruticose lichen communities on inland dunes in northwestern Germany. From tables in Tüxen (1937); ecological indicator values (M, R, N) from Ellenberg et al. (1992)a Community no.: Grasses and sedges:    Corynephorus canescens    Carex arenaria    Agrostis canina Da Festuca ovina s. str.? Other vascular plants: Cs Spergula morisonii Cs Teesdalia nudicaulis    Hypochoeris glabra    Jasione montana Cs Filago minima    Rumex acetosella Da Hypochoeris radicata Da Hieracium pilosella    Calluna vulgaris Bryophytes and lichensb    Polytrichum piliferum    Coelocaulon aculeatum Cc C. muricatum    Cladonia mitis    C. portentosa Cc C. uncialis    C. gracilis Cc C. fimbriata

1

2

3

M

R

N

5 4 1

5 2 4 3

3 1 2 3

2 3 5 5

3 2 3 3

2 2 2 1

5 3 2 2 1 4

4 5

3 1

3 3 5 4 4

2

2

3 3 3 3 2 4 5 4 5

5 1 3 3 4 2 4 5 1

2 1 1 2 1 2 3 2 1

4 5 2 3 4 4 4 4

2 3 6 5 5 5 5 5

2 5 5 5 5 2 3 4

– 2 2 1 1 2 1 1

5 3

2 1

3 2

The mosses and lichens were not included in the calculations of the average EIV scores, as the criteria for assigning indicator values differ from those used in vascular plants

No. 1 + 2: 1: 2: No. 3:

mM

mR

mN

3.0

2.4

1.8

3.5

2.7

2.0

>2.9

17)

Cladonia uncialis

Cladonia cervicornis subsp. verticillata

  Sphagnum cuspidatum   S. papillosum, S. tenellum   Cladopodiella fluitans L Cephalozia planiceps var. sphagnorum   Sphagnum pulchrum L Kurzia pauciflora L Cephalozia bicuspidata var. lammersiana   Sphagnum balticum   S. compactum Another 8 species

Several additional species

All of the species listed here have much wider physiological amplitudes than the conditions found in raised bog habitats. They are simply less competitive and therefore restricted to raised bogs, where most other species are unable to survive. Species promoted by drainage and/or burning (e.g. Molinia caerulea) are not included here b Previously much more common a

3.2  Environmental Conditions and Habitat Classification As mires are produced by plant communities, it is important to understand their species composition. However, good phytosociological characterisations are only available for mires in natural states, and even these are imperfect. The majority of mires support numerous communities of different synsystematic levels, which form mosaics or belts of vegetation that can be combined into higher units. We will therefore first provide an overview of the ecological mire types in which these mosaics occur.

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3.2.1  Peat Formation and Decomposition Mires are characterised by the accumulation of peat, i.e. partially decayed organic material that is saturated with water and decomposes very slowly due to the lack of oxygen. The build-up of peat is therefore not due to the high productivity of the vegetation, but rather because production and mineralisation occur at different rates. Peat forms either in the course of the terrestrialisation of lakes, or the paludification (i.e. wetting) of terrestrial habitats. Limnic peat formed during terrestrialisation processes contains not only organic matter from dead plankton, aquatic macrophytes and patches of moss, but also clay and sand. Depending on whether the sediments formed in oligotrophic lakes rich in humic acids or in eutrophic or hypertrophic lakes, the early stages of peat formation can take the form of dy, gyttja or sapropel. Calcareous oligotrophic lakes produce lime gyttja (see Sect. 4.2.2), although such clear-water lakes are now rare in Central Europe. Terrestrial soils undergo paludification when the inflow of groundwater increases or the evapotranspiration rate decreases, for example after the death or clear-cutting of a forest. In contrast to limnic peat, these semiaquatic or terrestrial peats contain lower proportions of mineral particles. The type of peat formed depends on the dominant peat-forming species, e.g. tall- or small-sedges, reeds, Sphagnum or other mosses, and the amount of oxygen available during its formation (Succow and Joosten 2001). The peat layers in mires show a characteristic vertical stratification into the upper acrotelm (peat formation horizon) and the lower catotelm (peat accumulation horizon), caused by the transition from aerobic to anaerobic conditions. The acrotelm contains a layer of living vegetation with photosynthetically active Sphagnum moss and vascular plants (euphotic zone, usually only a few centimetres thick) as well as the lower peat layers that consist mainly of increasingly decomposed and compressed remains of Sphagnum (aphotic zone). Decomposition is mainly aerobic here, as there is sufficient oxygen. Ivanov (1981) defines the lower limit of the acrotelm as the lowest groundwater level in the peat over several years of observations. The much denser catotelm is constantly saturated with water and has a decomposition rate around a hundred times slower than that of the acrotelm. Oxygen is rarely available and then only in the uppermost horizons. Microbial decomposition of the plant remains, which are sometimes several thousands of years old, is therefore solely anaerobic via lactic acid, ethanol or acetate fermentation and methanogenesis. The reducing conditions are visible in the presence of hydrogen sulphide and the black-coloured iron sulphide (cf. Fig. 3.28). It is not only the diffusion of oxygen that is hindered in the deep and dense peat bogs, but also the permeability to water is 10,000 times lower than in surface horizons, so that deeper peat layers only have low movement rates of water (Ivanov 1975). The deeper catotelm is a cold stenothermic habitat, which is frozen in boreal regions. Raised bog-peat in northwestern Germany can be divided into an upper white peat layer of little-­decomposed moss and a deeper, strongly decomposed black peat layer stained by humic substances. The former contains more fulvic acids, and the latter more humic acids.

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3.2.2  S  urface Structure and Morphological Classification of Mires Mires have been differentiated into ‘raised bogs’ and ‘fens’ since the very beginning of their exploitation, and the differences between the two are clearly visible, particularly in the northwestern lowlands of Central Europe and the foothills of the Alps (see Figs. 3.2 and 3.3). This division of mires is based on their morphological characteristics and ignores developmental, hydrological and chemical factors. It is therefore not always useful, particularly when regarding the ecology of these habitats. The surface of a fen follows that of the groundwater, so is mostly horizontal. In contrast, both lowland and upland raised bogs are formed of cushions of Sphagnum growing above the level of the groundwater. The overall surface of the bog forms a clear dome shape rising out of the surrounding area and is fed from rain water and snow. A raised bog is thus easier to drain than a fen, which can only collect and not release water to its surrounding areas. Fens are mainly fed from the groundwater, and so can form even in the driest areas of Central Europe, such as in the rain shadow of the Harz Mountains at under 500 mm precipitation per year.

Fig. 3.2  The view from the sloping edge of a raised bog in southern Bavaria (with Pinus mugo, Betula pubescens, dwarf shrubs and Rhynchospora alba in the foreground) to the central plateau of the bog

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Fig. 3.3  Semi-schematic cross-section through a raised bog close to the Alps with adjacent fen south of the Chiemsee lake (Bavaria) (Modified from Leiningen 1907). K–L = raised bog with mountain pine and birch; L–M = sloping edge with birch; N–O = spruce swamp forest; O–Q = transition to alder swamp forest; R–S = lakeshore vegetation (N–S = fen in the broad sense, i.e. including swamp forests)

Transition mires are complexes that are morphologically fens, but often found close to raised bogs and represent intermediate stages between the two in terms of their development and hydrology. In terms of water supply, they are a mixture of ombrogenous (fed from rainwater) and soligenous (fed from groundwater), as is typical for many aapa mires in the boreal zone. In these mires, the elevated ­hummocks are mainly fed by rainwater and the hollows by groundwater (Dierßen and Dierßen 2001). The detailed study by Jensen (1961) of mires in the Upper Harz documents the ecological gradient from mesotrophic minerotrophic fen to oligotrophic ombrotrophic raised bog, with transition mire habitats in between (see Fig. 3.4). The term transition mire (‘Zwischenmoor’) is used inconsistently in the literature. Some authors use it to describe acid, oligotrophic small-sedge communities adjacent to raised bogs belonging to the Scheuchzerietalia (Paul and Lutz 1941; Grosse-Brauckmann 1996; see Sect. 3.3.1). Pott (1995) and Philippi (in Oberdorfer 1977b) link it particularly to the alliance Caricion lasiocarpae. In contrast, Succow (1988) and Succow and Joosten (2001) speak of transition mires in a much wider sense, namely for all mesotrophic mires that have nitrogen levels somewhere between oligotrophic and eutrophic mires. The term is therefore variously used for intermediate habitats with respect to development, hydrology or nutrient status. To avoid ambiguity, this term should not be used (Sjörs 1950; Jensen 1961). A meaningful morphological classification of Central European raised bogs includes the following types, based on descriptions in Osvald (1925); Overbeck (1975); Jenik and Soukopova (1992); Dierßen and Dierßen (2001) and Steiner (2005). In order of increasingly favourable conditions for tree growth, these are: 1. Blanket bogs 2. Flat bogs 3. True raised bogs 4. Kermi bogs 5. Bog woodland ‘True’ raised bogs are naturally the most common type of bog in Central Europe, especially in flat, beech-dominated forest regions. In lowland areas, these true raised

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Fig. 3.4  Diagram of a soli-ombrogenous sloping mire with fen, transition mire and raised bog complexes in the upper montane spruce belt of the Harz Mountains (e.g. the Sonnenberg mire on the south face of the Bruchberg) (Modified from Jensen 1961). The spring water emerges above impermeable quartzite causing the formation of the mire via groundwater (i.e. soligenous). Although the spring water is highly acidic (pH < 4) and nutrient-poor, it deposits enough mineral nutrients for the upper part of the mire to have a fen character. Further down the slope, the water becomes increasingly nutrient-poor and finally oligotrophic (see also Table 3.5). The Molinia fen turns into an Eriophorum angustifolium fen and finally a Vaccinium uliginosum fen. The following patch, and to some extent the Vaccinium uliginosum fen, can be described as a ‘transition mire’. In the lower part, the mire has long been purely ombrogenous, i.e. a rain-fed bog. The Empetrum bog area occasionally receives groundwater inputs, which promotes more demanding plants, but this does not occur in the typical raised bog areas. The increasing amounts of nutrients (particularly nitrogen and calcium) deposited with the precipitation in recent years have caused the Empetrum community to spread, and the lower part of the mire has increasingly lost its oligotrophic character

bogs are basically flat, but can form domes with centres up to several metres above the edges, depending on the size of the bog. The edges are wetter than the centre, as the rainwater flowing from the dome collects when it comes into contact with the groundwater (see Fig. 3.2), giving these marginal areas a fen-like character (known as lagg-fen, or simply lagg). Upland bogs are also often surrounded by laggs, whether they are found in depressions, on slopes or on flat areas. The majority of a true raised bog is naturally treeless, and only the better-drained sloping edges may support tree growth (see Figs. 3.2 and 3.3). This is particularly the case in cool climates with consistently high rainfall, which promote the formation of large, treeless blankets of bog vegetation. These blanket bogs are relatively poorly developed in Central Europe and only occur in

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montane to subalpine regions, such as the Vosges and Krkonoše Mountains. Typical blanket bogs mainly occur in the oceanic west of Europe, such as in the undulating landscapes of Ireland. At the other extreme in the periodically relatively dry and continental climate of eastern Central Europe, trees can even establish in the centre of raised bogs. The forested sloping edges spread so far that almost the entire bog becomes bog woodland (see Fig. 3.20). In contrast to true raised bogs, flat bogs have neither domed centres nor lagg-­ fens. Marshy heathland in the surrounding area merges with the treeless bog without a clear border between the two, as can be seen e.g. in southern Sweden and Great Britain. Kermi bogs, which have a convex dome and are similar to true raised bogs, are typical for the boreal zone. In contrast to true raised bogs, however, the transition between the lagg and the convex dome is more gradual (Dierßen and Dierßen 2001). The relatively hostile climate for tree growth in the north of Europe supports a further two treeless types of bog, namely aapa and palsa mires. Aapa mires contain rain-fed strings (hummocks) formed by freeze-thaw processes, and groundwater-fed elongated flarks (hollows). Palsa mires occur at the edge of the arctic permafrost, where frozen ice lenses lift mounds of peat several metres above the surrounding area. Only very few mires have a flat surface. Usually there is small-scale variation in the depth of the peat, so that they contain a mixture of water-filled hollows and drier, elevated hummocks (see Fig. 3.5). The causes of this microrelief are probably differences in the average water level, allowing moss and vascular plant species that differ in their tolerance to flooding as well as in their peat-forming abilities, to dominate in patches in close proximity to each other (see Sect. 3.6.2). This microrelief is seen at its clearest in typical domed raised bogs. In larger growing bogs, the centre is always the wettest, as the precipitation can only drain slowly towards the edges. Scattered hollows form here, i.e. water-filled or -saturated, shallow depressions, colonised by bright green Sphagnum and the occasional, usually graminoid, phanerogam. These hollows are rarely more than a few metres wide and several decimetres deep (see Fig. 3.6). The hollows alternate with taller, cushion-shaped hummocks of 0.5–3  m ­diameter, formed from red, brown or yellow Sphagnum species. Each of these ­hummocks is a mini version of a raised bog. Depending on its position in relation to the water table in the mire, it is covered either by tufts of Eriophorum or Cyperaceae, or by dwarf shrubs usually found in heathland (see Fig.  3.7). Some of these ­phanerogams are only able to establish if their roots are not continually standing in water, and at least the upper few centimetres are temporarily aerated. Further ­microstructural elements include lawns and carpets of dense moss. Areas of bog with relatively rapid peat growth have contiguous and usually flat Sphagnum layers with a general absence of hummocks and hollows (often termed ‘growth complexes’). The edges of bogs, in contrast, are often characterised by dwarf shrub- and lichen-rich hummocks but few hollows (‘stagnation complexes’). In northwestern Germany, these are often dominated by Erica communities (Overbeck 1975). Peat accumulation is slowed here mainly by faster run-off of rainwater due to the steeper incline, leading to faster peat decomposition. In places

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Fig. 3.5  A mosaic of vegetation in zones of differing moisture levels in the Esterweger Dose (western Lower Saxony), formerly one of the largest raised bogs in northwestern Germany (Modified from Jahns 1969). Each of the four cross-sections is 6 m long with vertical exaggeration. The bog had already been drained at the edges at the time of the vegetation mapping and did not form a proper dome. The following plant communities can, however, still be found in parts today. The middle of the bog supports Sphagnum pulchrum-Rhynchospora alba hollows with flat Sphagnum magellanicum hummocks. Part of the central area where the water levels have not been affected contain dystrophic pools, the more nutrient-rich edges of which are colonised by ‘fen species’ such as Carex rostrata, Carex limosa, Molinia caerulea, Empetrum nigrum, and more nutrient-­demanding moss species. They correspond to the edges of flarks in kermi bogs (see Fig. 3.10). The dry edge areas drained by surface water run-off never had true hollows, and the frequently dry hummocks provide sufficient light for indicators of periodic drought such as Leucobryum glaucum, Hypnum ericetorum and fruticose lichens. A wide transitional zone has intermediate conditions, but also its own unique character. For example, the hollows here are colonised by Sphagnum tenellum, the edges of which support Sphagnum papillosum, which used to dominate such mires. The numbers refer to the different types of plant communities found in these areas, the following of which are listed here as examples: 1= Sphagnetum cuspidato-obesi; 2= Cuspidato-Scheuchzerietum, Carex-subass.; 3= C.-S., typical subass.; 4= Sphagnum cuspidatum-­ Eriophorum angustifolium community; 5 = Rhynchosporetum sphagnetosum cuspidati, typical var., Sphagnum pulchrum-rich form; 6= R.s.c., Sphagnum tenellum-rich form; 7= R.s.c., variant of Erica tetralix, typical subvar., Sphagnum pulchrum-rich form; 8= as above, Sphagnum tenellum-­ rich form; 9= as above, variant of Erica tetralix, Cladonia-Subvar., Sphagnum tenellum-rich form; 10 = Sphagnetum papillosi, subass. of Rhynchospora alba; 11 = Sphagnetum magellanici, subass. of Rhynchospora alba, typical var. ; 12 = as above, variant of Calluna vulgaris; 13 = as above, typical subass., variant of Calluna vulgaris; 14 = as above, subass. of Aulacomnium palustre, variant of Empetrum nigrum; 15 = as above, variant of Eriophorum vaginatum, typ. subvariant; 16 = as above, subvar. of Cladonia portentosa; 17 = Sphagnetum fusci, facies of Aulacomnium palustre

where the flowing water cuts into the peat, erosion complexes form with wet gullies and their well-drained edges that promote the growth of trees. The desiccated peat is susceptible to wind erosion, allowing small dust bowls to form. The driest parts of an active raised bog are its sloping edges, which fall away quite steeply into the lagg-fen and are occasionally scored by deep channels. Relatively dense forests of birch or pine can establish here, which are floristically quite different from the Sphagnum-rich communities of the rest of the mire complex

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Fig. 3.6  Map of the flowering plants and bryophytes in a 6 × 10  m section of the western Ahlenmoor, south of the Niederelbe (northern Lower Saxony) (Modified from K. Müller 1965) (cf. the cross-sections through the Esterweger Dose, Fig. 3.5)

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Fig. 3.7  Schematic cross-section through a Sphagnum papillosum-rubellum hummock in an Atlantic raised bog in the Ardennes in Belgium. The ice lenses beneath the Sphagnum rubellum hummocks were recorded on the 27th December 1949 after several days of temperatures above freezing. They are much shorter-lived and thinner than ice lenses in subcontinental raised bogs (From Vanden Berghen 1951)

Fig. 3.8  A large dystrophic raised bog pool in the Wildseemoor near Kaltenbronn (Black Forest). The sheltered bank in the foreground supports a quaking bog (Eriophorum angustifolium and Sphagnum cuspidatum) and floating Sphagnum cuspidatum. The opposite bank has been eroded by wave action and only supports a small belt of Carex rostrata. The Pinus mugo bog forest reaches almost to the bank. A spruce swamp forest can be seen in the background

(see Fig.  3.2). Ponds within the centre of the mire (bog eyes or kolks) are often formed as the soft, water-saturated peat gives way. In contrast, pools in bogs formed via terrestrialisation are often the remains of the former lake (see Fig. 3.8). In the relatively cold climates of the eastern low mountains, as well as in Scandinavia and the Baltic, raised bogs often contain large groups of long hollows arranged in parallel rows (see Figs. 3.9 and 3.10). These 1–4 m wide channels are termed flarks, and are separated by 0.5–5  m wide, elevated strings (‘kermi’ in

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Fig. 3.9  Flark complex of a kermi bog near Riga (Latvia), shortly after the snowmelt in spring (cf. Fig 3.10). The ice is melting first on the south-facing bank

Finnish). These flark-string complexes are typical for many of the relatively high domed kermi bogs in southern Scandinavia, central and southern Finland, the Baltic states, Belarus and Poland (see Fig. 3.10). Usually, flarks follow the concentric contours of the domed mire. They are probably formed by the centrifugal movement of the waterlogged and flexible peat, leaving depressions that were then filled with water. The release of oxygen from photosynthesis by the aquatic plants in the flarks promotes the decomposition of peat at their edges, causing them to expand. A frequent phenomenon of active, as well as drained, mires is the seasonal up and down movement of the mire surface by several centimetres per year, such as in the Königsmoor (Lower Saxony) by 1.5–3.0 cm. This is mainly caused by changes in the volume of stored water and known as ‘bog breathing’ (Overbeck 1975).

3.2.3  Macroclimate and Mire Formation In contrast to fens, raised bogs depend on sufficient precipitation and low levels of evaporation. Gignac (1993) states that Sphagnum only grows well in areas where precipitation is higher than evaporation. The sponge-like body of the mire itself, made up mainly of Sphagnum, can therefore only grow in humid-temperate climates. However, excessively large amounts of precipitation can also reduce the growth of Sphagnum. Especially the hummock-forming Sphagnum magellanicum

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Fig. 3.10  Flark complex of a kermi bog in the Neman delta (Lithuania) (Modified from Hueck 1934)

and S. fuscum, which grow highest above the water table in the mire, cannot grow if they are repeatedly flooded for long periods of time. ‘True’ raised bogs and Kermi bogs therefore mainly form in suboceanic and montane climates, i.e. in Southern Sweden, southwestern Finland, the Baltic states, the northwestern plains of Central Europe, the majority of low mountain ranges in Central Europe, the northern foothills of the Alps and the montane to upper montane zone of the outer edges of the Alps. Well-developed raised bogs are found only up to around 500–800 m a.s.l. in the northern Alps (Gams 1962). In the central Alps, raised bogs are also found much higher (between 1000 and 1800 m) due to the greater summer radiation levels (Grünig et al. 1986). Sufficient warmth and length of the growing season are therefore, alongside the amount of precipitation, important factors for the formation of raised bogs. Despite the fact that Sphagnum needs warmth to grow, all regions rich in raised bogs have a relatively cool climate. This is because the high temperatures during the dry periods that often occur in the warmer parts of Europe promote the decomposition of peat, as well as reducing the growth rate of Sphagnum by increasing evapora-

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tion and thereby drying the surface of the mire. As a result, no typical Sphagnum raised bogs are found in the warm continental regions of the east and southeast, as the minimum precipitation required for bog formation is around 620  mm year−1. There are also no raised bogs in the dry summer climate of Southern Europe, even in places where the annual precipitation exceeds 1000 mm. Cool climates therefore primarily promote the formation of raised bogs indirectly, by providing higher and more homogenous air humidity and reduced drought stress for the hummock-­ forming mosses. Nevertheless, the colder and shorter the growing season, the lower the total growth of the mire. As a result, there are no true raised bogs above the timber line in mountainous areas or in the boreal climate of northern Scandinavia. In addition to the low temperatures, also the large quantities of meltwater and the movement of the ground through frequent freezing and thawing reduce mire growth rates.

3.2.4  Microclimate Mires are extreme habitats, not only in terms of hydrology but also from a microclimatic perspective (Schmeidl et al. 1970). This is particularly the case for raised bogs with their thick peat layers. The wet peat remains frozen or very cold until well into the summer (see Fig. 3.7). This is due to the relatively large amount of ice that forms during the winter as well as the high specific heat capacity of water, which slows the warming in the summer. In addition, the low thermal conductivity of the Sphagnum cushions and the peat also plays a role, which when wet is around 5 times and when dry as much as 10 times lower than that of wet sand (Geiger 1961). As a result, the uppermost few centimetres of a Sphagnum hummock can be very warm under the midday sun in spring, but 10 or 20 cm beneath the surface there is still ice. Even in June or July, the temperatures at this depth can still be under 10 °C, or even closer to 0 °C. Layers of peat several metres deep experience very small seasonal changes in temperature, and as cold, stenothermic habitats have very low decomposition rates (Dierßen and Dierßen 2001). Low temperatures and large changes in temperature apparently have only a relatively small effect on the uptake of water by most raised bog phanerogams (Firbas 1931). This is in clear contrast to the temperature-­ sensitivity of the water uptake of many herbaceous species in non-mire habitats or cultivated species, which often have reduced uptake rates already below 10 °C. This is largely due to the increasing viscosity of water with decreasing temperature, as well as the reduced permeability of cell membranes in the roots (Lyons et al. 1979). Cultivated raised bogs are particularly cold habitats, and not suitable for frost-­ sensitive crops. The cold temperatures in mires even affect their surroundings, for example in Aurich and other localities lying between drained raised bogs in ­northwestern Germany, which experience surprisingly cold nights despite their proximity to the coast. Drained raised bog peat has an even lower thermal conductivity than wet peat, due to its high air content. As a result, the lower layers take a particularly long time to warm up in summer, although the uppermost layers where the

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plants germinate are quickly warmed by the sun. Ground frosts frequently occur on clear nights due to the low levels of stored heat in the lower layers of drained mires. The microclimate of the dry surfaces of Sphagnum hummocks is relatively ‘continental’, in that the air above them warms up quickly in the sunlight. Firbas (1931) measured rates of evaporation from treeless raised bogs only 20–30 % lower than those on rocky outcrops and other sun-exposed habitats. Schmeidl (1964) measured a maximum temperature of 43 °C above the surface of wet Sphagnum hummocks in a mire near the Chiemsee (southern Bavaria), 60 °C above the dried but living hummocks, and 77 °C above bare, dry peat. These temperatures were considerably lower during the night. The Sphagnum in the water-filled hollows also warmed to 36 °C during the day. Thus, in contrast to the hummocks, hollows have a more ‘oceanic’ microclimate (Schmidt 1997), that is relatively warm in winter and allows the peat beneath it to thaw almost a month earlier in spring.

3.2.5  Water Regimes and Hydrological Mire Types Rainwater- vs. Groundwater-Feeding  The water regime is central to the formation of mires, and influences not only the composition of the vegetation but also the ratio of peat growth to peat decomposition. From a hydrological point of view, each true raised bog is a microcosm, in fact more so than e.g. a lake. It receives water only from precipitation and nutrients only from the atmosphere, i.e. an ombrogenous formation of peat. According to Kulczinsky (1949), it is mainly the lack of flowing water that is important in the formation of a raised bog. All other types of mire are dependent on inflows from groundwater or surface run-off, meaning their peat deposits are geogenous. The surface of mires fed by groundwater follows that of the water table or the spring water level and is therefore more or less flat (Du Rietz 1954). In contrast, typical raised bogs lie like a dome over a terrestrialised lake or paludified mineral soil (see Fig. 3.29) and form their own water table, independent of the groundwater in the surrounding area. This is also the case where the raised bog is not obviously elevated above its edges (see Sect. 3.2.2). Rainwater falling onto raised bogs flows downhill over the surface from the centre to the sloping edges, whilst surface run-off water from outside the bog flows around its domed centre and remains in the lagg (see Fig. 3.11). Capillary flow from the groundwater into the bog also occurs occasionally during dry periods (Devito et al. 1997). The water regime in fens is different. These are fed from groundwater or surface run-off from the surrounding area (see Fig. 3.11), and therefore contain many plant species that are dependent on groundwater. The difference between rain-fed and groundwater-fed mires is visible in the occurrence of ‘groundwater indicator’ species, particularly in regions with low levels of atmospheric nutrient deposition (Okland et al. 2001, but see Wheeler and Proctor 2000). The influence of the more base-rich groundwater is mainly dependent on the speed of the flow and less on its ion concentration and pH (Vitt et al. 1995), leading Ruttner (1962) to speak of a ‘eutrophying effect of the flow’.

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Fig. 3.11  Schematic illustration of the water balance in a fen and a raised bog with the major water fluxes (From Streefkerk and Casparie (1989) in Dierßen and Dierßen (2001); with permission of Ulmer Verlag, Stuttgart)

Hydrological Mire Types  Because the water regime has a major influence on the species composition and ecology of mires, it makes sense to classify mires according to their water regimes and the resulting developmental processes. For Central Europe, Succow (1988) and Succow and Joosten (2001) have produced a system of hydrologically-defined mire types, which we will follow here (see Fig. 3.12; cf. also Steiner 2005). Paludification mires developed during humid climate phases after the late glacial, as a rising water table led to the saturation of mainly flat areas of mineral soil. In Central Europe, this type of mire is widespread mainly in Pleistocene lowlands (see Fig. 3.13) and are naturally mainly eutrophic or mesotrophic habitats. The peat layer is thin, and formed mainly from reedbeds, alder swamp forests and tall-sedge communities. Today, most of these mires have been drained and turned into managed grassland. Flood mires form in floodplains as well as in coastal areas from the deposition of alluvial clay and silt by overflowing water, which produces heavy, mineral-rich peat due to the hypoxic conditions. The movement of water occurs mainly above the peat body, and peat formation occurred sporadically during floods. Such nutrientrich habitats close to rivers are colonised by tall-sedge communities, reedbeds and alder swamp forests (see Fig. 3.14). Large areas of flood mires were also formed along the North Sea and Baltic Sea coasts during the rising sea levels after the last ice age. Today, they are either on the sea floor or have been replaced by salt marshes or agricultural grasslands.

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Fig. 3.12  Cross-sections of eight hydrological mire types with their characteristic layering of the types of peat and the corresponding zonation of the vegetation. Paludification mires are not shown. These examples are mainly taken from northeastern Germany (From figures in Succow and Jeschke 1990 and Succow and Joosten 2001). Peat and sediment types: (a) CarexSphagnum peat, (b) Carex -non-Sphagnum peat, (c) Carex-Phragmites peat, (d) Phragmites-Carex peat with Sphagnum, (e) tall-sedge peat, (f) Phragmites peat, (g) birch swamp forest peat, (h) pine swamp forest peat, (i) alder swamp forest peat, (j) non-Sphagnum peat, (k) Sphagnum peat, (l) younger Sphagnum peat, (m) older Sphagnum peat, (n) Cladium peat, (o) spring fen peat,

Fig. 3.13  Complex of river flood and percolation mires along a semi-natural river in the Pleistocene lowlands of Kaliningrad Oblast (Russia) close to the Baltic Sea coast

Fig. 3.12  (continued) (p) Juncus ­gerardii (salt marsh) peat, (q) Sphagnum-Scheuchzeria peat, (r) Sphagnum-Eriophorum peat, (s) calcareous peat, (t) tufa from calcareous spring, (u) clayey floodplain sediment, (v) sandy floodplain sediment, (w) lake-bed mud, (x) lake-bed mud rich in algae, (y) coarse-detritus mud with silt, (z) silicate mud, (ä) organic mud, (ö) calcareous mud, (ü) nonSphagnum peat with mud. Vegetation types in the figures (see numbers): Ombrotrophic raised bog (1 – sloping edge, 2 – central plateau, 3 – hummock-hollow complex, 4 – sloping edge with hummock complex and wet heath), Terrestrialisation mire (1 – Molinia-beech forest, 2 – Carici elongatae-Alnetum, 3 – Salix pentandra-­sedge swamp, 4 – moss-sedge swamp, 5 – moss-sedge swamp with Salix repens and Betula humilis, 6 – Scorpidium-Eleocharis swamp, 7 – Scorpidium-Cladium reed, 8  – lake with Potamogeton and Nymphaea, 9  – Phragmitetum with Solanum, 10  – Carici elongatae-Alnetum and adjacent beech forest), Percolation mire (1 – Molinia-Quercus robur forest, 2  – Eriophorum-­Betula pubescens swamp forest, 3  – original peat-forming vegetation, 4  – Eriophorum-Sphagnum lawn, 5  – Sphagnum-Carex-Eriophorum swamp, 6  – Sphagnum-Juncus effusus swamp with Phragmites), River flood mire (1 – Galio-Fagetum, 2 – Sambucus-oak forest, 3 – Symphytum-Iris-­alder carr, 4 – Salix alba-floodplain forest, 5 – Ulmo-Quercetum, 6 – Salix alba-floodplain forest, 7 – Poa palustris-Phalaridetum, 8 – reedbed, 9 – oxbow lake with floating Myriophyllum and Nymphaea, 10 – reedbed, 11 – Caricetum gracilis, 12 – Ulmo-Quercetum, 13 – Sambucus-oak forest, 14 – Salix triandra scrub), Sloping mire (1 – Calamagrostio-Piceetum, 2 – Sphagnum-­Scirpus sylvaticus swamp, 3 – Sphagnum-Carex-Eriophorum swamp, 4 – ‘coloured’ Sphagnum lawn, 5  – Sphagnum-Carex-Eriophorum swamp, 6  – Calamagrostio-Piceetum, 7  – Sphagnum-­Juncus-­Carex nigra swamp, 8 – Sphagnum-Juncus effusus swamp, 9 – CalamagrostioPiceetum), Spring fen (1  – Calamagrostio-Piceetum, 2  – Cardamine-alder carr, 3  – Carex paniculata-swamp alternating with Solanum-reedbed and Cardamine-alder carr, 4 – Cardaminealder carr, 5 – Milium-ash forest, 6 – Carex remota-ash forest, 7 – reedbed with Characeae), Kettle mire (1 – Galio-Fagetum, 2 – Sphagnum-Thelypteris-Salix aurita scrub, 3 – Carex-Eriophorum swamp, 4 – Eriophorum-Sphagnum lawn alternating with pine-Sphagnum swamp, 5 – SphagnumCarex-­Eriophorum swamp, 6 – Sphagnum-Carex-Eriophorum swamp, 7 – Luzulo-Fagetum with Leucobryum)

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Fig. 3.14  A classification of the landscapes of Central Europe according to the dominant hydrological mire types after Succow and Jeschke (1990). Landscapes with mainly coastal flood mires (1), lowland raised bogs (2), paludification mires (3), terrestrialisation and kettle mires (4), percolation mires (5), upland raised bogs and sloping mires (6), terrestrialisation and upland raised bogs (7), paludification and sloping mires (8). Landscapes with less than 1 % mire are shown in white

Sloping mires form in the low and high mountain areas on silicate rock and are fed from surface run-off from higher-lying areas. Some were formed only during the Middle Ages, when forest clearance led to reduced evapotranspiration on the landscape level. Sloping mires generally grow uphill, resulting in the older, lower-­ lying mire areas being less nutrient-rich than the younger parts. Spring fens are small, marshy areas where groundwater emerges on slopes in the mountains or in Pleistocene ground moraine landscapes. Peat formed in spring mires is usually highly decomposed due to the high oxygen content of the spring water, and is also rich in iron and mud, as well as being calcareous in limestone areas. Terrestrialisation mires formed from water bodies in which the primary production led to the formation of gyttja from layers of dead plankton and aquatic macrophytes, which were later covered by peat from mosses, sedge and reed species. These are often formed from the terrestrialisation of quaking mires, which are floating mats of peat colonised by mosses and vascular plants. Terrestrialisation mires can range from oligotrophic to eutrophic.

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137

Kettle mires are small mires in natural depressions without outlets, as e.g. in the glacial dead ice holes of top moraine landscape or in the volcanic ‘maar’ lakes. Usually, the mire formation began at the end of the late glacial with the production of gyttja, followed by the deposition of moss and sedge peat. Percolation mires developed secondarily from terrestrialisation, flood, sloping or spring mires as the result of the increasing supply of groundwater. They are characteristic for the edges of valleys of large lowland rivers, where groundwater flows through the slightly sloping mire (see Fig. 3.13). With increasing distance from the valley edge, ion exchange in the peat body leads to increasingly nutrient-poor and acidic conditions. Probably the largest percolation mire in Central Europe is in the Biebrza valley in northeastern Poland (Succow 1988). Raised bogs are rain-fed (ombrotrophic) mires in humid regions (see Fig. 3.1) that have developed from sloping, kettle or percolation mires through the growth of thick peat layers. These are the ‘true’ bogs. Bogs can also develop in small areas at the foot of scree slopes in mountainous areas, when incoming air is cooled by the rock and deposits condensed water (‘Kondenswassermoore’, Steiner 1992). The main features of these hydrological types of mire are summarised in Table 3.2. Dierßen and Reichelt (1988) studied the water regime of microforms in raised bogs using water level duration curves (see Fig. 3.15). In the hollows, the water was on average 0–20 cm above the surface of the soil, but in the adjacent dense moss layers it was around 10  cm below. The lawn area contained more phanerogams (mainly Cyperaceae) and was on average even drier (−20 cm), as were the hummocks colonised by drought-tolerant mosses and dwarf shrubs (−40 cm). Areas of active peat growth that are rich in Sphagnum will therefore have average water levels of between 15 and 40 cm below the surface. In addition to the average water levels, their fluctuation over time and the associated changes in redox potential also affect the mire vegetation (de Mars and Wassen 1999).

3.2.6  Nutrient Supply and Trophic Mire Types Availability of N, P and Basic Cations  From rich, eutrophic fens to poor oligotrophic or dystrophic raised bogs, there are many chemical gradients with which the nutrient supply of mire plants can be characterised. Most important are nitrogen and phosphorus, but also pH, electrolyte content of the water, base saturation and potassium and calcium supply may be influential (Wassen and Joosten 1996; Okland et al. 2001; Hajek et al. 2006). Most comparative studies on nutrient supply in mires are based mainly on the chemical analysis of peat samples from the upper horizons and the nutrient c­ ontents of the living plants, and rarely on the exchangeable and dissolved nutrients in the acrotelm. The carbon content of the peat may serve as another indirect measure of nutrient availability, which increases from

Reedbeds (partly quaking bogs) c. 15

Main peat-­ forming vegetation

From Succow and Joosten (2001)

% of total mire area

In depressions and valleys

Main occurrence

Flat

Moderate to fast Moderate

Peat growth

Peat decomposition Structure of mire surface

Slow

Water movement

Main water source

Terrestriali-­ sation mires Lake/pond water

c. 5

Reedbeds, swamp forests

Rough, undu-lating surface Along lowland rivers

Fast

Only at surface Slow

Flood mires Periodic flooding

c. 30

c. 1–2

Sphagnum-­ sedge reeds

Swamp forests

Non-­ Sphagnum moss-sedge fens c. 25 S. rubellum > S. balticum > S. tenellum, meaning that S. tenellum dries out quicker than S. fuscum. Particularly dense growth among hummock species also slows the loss of water, and is probably the reason why the hummock mosses S. fuscum, S.rubellum and Leucobryum glaucum have lower evaporation rates than the hollow and lawn species S. recurvum and S. magellanicum (Overbeck and Happach 1957). Indeed, hollow species like S. tenellum bleach much more frequently than hummock species such as S. fuscum, resulting in particularly low growth rates of moss in the hollows during dry years (Rydin et al. 2006).

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Fig. 3.26  Hydrological niches of dominant cryptogams in mires in the Black Forest; water level below- or above-ground (mean, standard deviation, extreme values, right: number of measurements) (From Dierßen and Dierßen 2001; with permission of UImer Verlag, Stuttgart)

However, hummock and hollow mosses do not differ in the effects of drought on growth and photosynthesis, and their recovery after drying out (Clymo and Hayward 1982; Rydin 1993). Green (1968) states that only S. imbricatum is particularly resistant to drought compared to the other species. Dierßen and Dierßen (1984) determined the optimum water level of twelve species of Sphagnum in the Black Forest (see Fig. 3.26). Of the above-mentioned species, only S. cuspidatum and to a lesser extent S. tenellum could grow submerged. All of the true hummock species have to at least have their heads above the water, including S. magellanicum, which has the widest hydrological amplitude of all of the species. There are probably several factors responsible for this variation in hydrological niches. Differences in water conductance and evaporation rates probably determine the zonation of the drier hummocks, while competitive ability is the main factor in the hollows. Nevertheless, some species of hummocks such as S. fuscum are prevented from colonising wetter areas through their lack of flooding tolerance rather than their competitive weakness (Rydin 1993). As in other semi-aquatic habitats, lack of oxygen and potential toxicity of sulphide, manganese or iron (II) ions is typical for the root zone of mires (Crawford 1983). Mire plants, like those of limnic and marine habitats, exhibit various adaptations to these stressors. Molinia is particularly sensitive to anoxia (Boatman and Armstrong 1968), which explains its strong link with groundwater levels. In contrast, species of raised bog hollows (e.g. Eriophorum, Scheuchzeria) are particularly tolerant (Kinzel 1982; Wheeler et al. 1985). Adaptations to anoxia are discussed in more detail in Sects. 4.4.3 and 8.4.2.3. According to the findings of Kotowski et al. (2001), fen species such as Viola palustris and Hydrocotyle vulgaris should primarily be seen as light-demanding herbs that are weak competitors, but which are less damaged by the anoxic conditions in the mires than by shading from more productive terrestrial plants. Thus, as

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is the case in many other habitats, mire vegetation can only be understood in the context of the competitive relationships between its species.

3.4.2  A  daptations to Low Nutrient Levels and the Role of Base Richness Nutrient Limitation in Mires  Oligotrophic and mesotrophic mires are habitats with low nitrogen and/or phosphorus levels. Many mire species therefore have special adaptations to the lack of nutrients. This is particularly the case for species of raised bogs, which only receive nutrient inputs from rainwater and also experience high levels of nitrogen immobilisation by microorganisms. The nutrient supply is also quite poor in calcareous fens, as P is bound by the precipitation of calcite, hydroxylapatite or octocalcium phosphate, and therefore usually largely unavailable (Boyer and Wheeler 1989). Many of the species living here are probably able to mobilise bound phosphate through secretion of external phosphatases and acidification of the rhizosphere (Koerselman and Verhoeven 1995). As a result, growth rates can be improved more by the addition of N than of P. Raised bog vegetation in areas with low levels of N deposition is generally thought to be N limited, whilst those in areas with higher N inputs is presumed to be P limited (Aerts et al. 2001; Limpens et al. 2006). Boeye et al. (1997) found strong P limitation of growth in calcareous fens with low productivity, whereas N was limiting in more productive fens. This is disputed by Koerselman and Verhoeven (1995), who predict mainly P limitation of growth also in bogs with low deposition rates. K is rarely a limiting factor in mires (de Mars 1996), and Ca limitation has not been recorded even in acid bogs. The moss, graminoid, forb and dwarf shrub species of mires possess a variety of different adaptations to the low N and P availability. The thick, evergreen leaves of most raised bog phanerogams resemble xeromorphic plants of dry habitats, such as steppes. This phenomenon (‘peinomorphy’, Greb 1957) is interpreted as an adaptation to low nutrient levels (particularly of P), and perhaps also as a response to low temperatures in the root zone (Kinzel 1982). Evergreen leaves increase the nutrient use efficiency, in that relatively few nutrients are invested in the production of the leaf, and these can be largely recovered before the leaf is shed. Sphagnum species concentrate their nutrients in the uppermost parts of the plant, presumably mainly so that they are not lost in the parts that turn into peat (Malmer 1988). Intensive internal relocation from the dying stems to the young shoots also means that the mosses have a high nutrient use efficiency (Malmer 1993) as well as producing very nutrient-poor peat. Aerts et al. (1999) state that raised bog plants also have a particularly high use efficiency of phosphorus. Mires contain all types of mycorrhizae; ericoid mycorrhizae among the Ericaceae (e.g. Vaccinium and Erica), arbuscular mycorrhizae among several Poaceae and Cyperaceae (e.g. Molinia), orchid mycorrhizae among the mire orchids Liparis, Listera and Orchis, and arbutoid mycorrhiza in Arctostaphylos (Dickinson

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1983). Some species of Cyperaceae and Juncaceae (e.g. Eriophorum vaginatum) apparently also grow well without mycorrhizal symbiosis. The Sphagnum species also appear to have a mutualistic relationship with certain fungal species, although these are not mycorrhizae, but rather from the genus Penicillium (Küster 1993). The free-living zygomycete Mortierella is regularly associated with Sphagnum leaves, and the basidiomycete Omphalina is often found in Sphagnum lawns (Burgeff 1961; Dickinson 1983). Nevertheless, Sphagnum can also be grown under sterile conditions without fungal partners. Nutrient Acquisition in Sphagnum  Sphagnum species can use both ammonium and nitrate, and grow best with a mixture of both. In this respect they are similar to the majority of terrestrial vascular plants (cf. Sect. 4.5.2.6 in Volume I), but differ in that they metabolise nitrate immediately, as its uptake through the cell wall occurs very quickly (Rudolph et al. 1993). NO3− deposition and the induction of nitrate reductase activity were closely linked in S. fuscum, so that its assimilation occurred directly after input (Woodin et  al. 1985). This ability, together with a very large surface area for uptake, produces the unusual capacity of Sphagnum to use (or immobilise) all the ammonium and nitrate received via precipitation (Lee et  al. 1990). Thanks to its water-storing hyalocyte cells, Sphagnum can hold 25 times its dry weight in water. This water absorption probably primarily serves to increase the nutrient uptake of the plant by providing a greater surface area for cation exchange, rather than the storage of water itself (Frahm 2001). Mosses therefore outcompete the vascular plants of raised bogs in terms of nitrogen use (Rydin and Clymo 1989). Sphagnum also has an unusually high capacity for assimilating the available Ca, Mg and K pools; these elements are mainly bound in the hyalocytes due to the very high cation exchange capacity of their cell walls. The high exchange capacity is primarily achieved by the large quantity of negatively charged pectic acid (polygalacturonic acid) in the cell walls, which can make up over 10 % of the cell wall mass (Dainty and Richter 1993), particularly in the hummock mosses (Clymo and Hayward 1982). The cation adsorption leads to the release of hydrogen ions and thus acidification of the peat. Sphagnum can therefore affect its substrate not only through a very efficient exploitation of nitrogen and other nutrient elements, but also through a considerable decrease in pH. The hummock Sphagnum species have a higher cation exchange capacity than the hollow mosses (Vitt et al. 1975), so that raised bog hummocks are often more acidic than hollows (Clymo and Hayward 1982; Wagner 2000). Sphagnum is considered a calcifuge genus, although some species (e.g. Sphagnum subsecundum or S. obtusum) also occur in neutral or basic conditions. Ca2+ is clearly not toxic for it, at least at low concentrations and at low pH. The damage is caused instead by OH− and the general lack of carbon dioxide under alkaline conditions (Clymo 1973; Kinzel 1982). In contrast, the plants of neutral to basic mires may be excluded from acid raised bogs by the lack of HCO3−. The Sphagnum species of acid habitats and those of neutral to basic habitats differ both in their sensitivity to the basic pH and in their acidification of the substrate (Clymo and Hayward 1982; Kooijman and Bakker 1994). Dierßen and Dierßen (1984) recorded the distribution

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Fig. 3.27  pH amplitudes of selected cryptogams from mires in the Black Forest (mean, standard deviation, extreme values, right: number of measurements) (From Dierßen and Dierßen 2001; with permission of UImer Verlag, Stuttgart)

of 14 cryptogam species along a gradient of pH in mires in the Black Forest (see Fig.  3.27). They found a clear preference for weakly acidic to neutral habitats among the mosses Scorpidium revolvens and Campylium stellatum as well as several Sphagnum species (S. obtusum, S. subsecundum). Nutrient Acquisition in Vascular Bog Plants  The vascular plant species of raised bogs are not as efficient as Sphagnum at acquiring nutrients. The chemical composition of the dead Sphagnum litter means that its decomposition, and thus nutrient release via mineralisation, is so slow that the vascular plants are ‘starved’. The inputs of N, P and K via precipitation are also immediately absorbed by the Sphagnum, and Ca, Mg and S is only available to the higher plants from the ions in the water that drains through the layer of moss. Raised bogs thus show a notable differentiation between the nutrient supply of Sphagnum and of vascular plants, whereby the former are supplied mainly by rainwater and the latter by the slow mineralisation of the peat. Species of Cyperaceae with deep roots can also use the low concentrations of N in the water of the catotelm. Many vascular plants of ombrotrophic raised bogs, and particularly members of the Ericaceae, are calcifuge species that probably preferentially take up ammonium. Dwarf shrubs can only colonise peat where their mycorrhizae have a sufficient oxygen supply. The fungal symbiont supplies its host not only with ammonium, but also with low molecular weight organic N compounds, including various amino acids and oligopeptides (Näsholm et  al. 1998). This adaptation gives the plants a competitive advantage over other vascular plants due to the slow N mineralisation of raised bogs. Following Högbom and Olson (1991), mire species differ considerably in their nitrate reductase activity. The dwarf shrubs Vaccinium oxycoc-

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cos and Andromeda polifolia had much lower maximum activities than Eriophorum angustifolium and Saxifraga hirculus. It is unclear whether this is linked to the access of the Ericaceae species (i.e. Vaccinium and Andromeda) to organic N compounds. The greater input of groundwater in fens means that they have a better supply of Ca, K and Mg, which is visible in the higher contents of these minerals in the Sphagnum biomass (Malmer 1993). The more basic the peat, the higher the mineralisation rate and nutrient release, which explains the increasing dominance of Cyperaceae, Juncaceae and other vascular plants with increasing base richness (Damman 1986; Malmer 1993). Given the low nutrient supply in the substrate, it is unsurprising that some vascular plants use additional N and P sources. This is particularly noticeable in the large number of species of carnivorous plants in raised bogs, including many types of Drosera (sundew), Utricularia (bladderwort) and Pinguicula (butterwort). As a typical raised bog hollow species, Drosera rotundifolia can absorb ammonium and nitrate from the peat, but can grow larger and produce more seeds by obtaining additional nutrient from the animals that it catches (Crowder et al. 1990). According to Mejstrik (1976), this species also has a mycorrhizal symbiosis. D. rotundifolia is sensitive to high Ca concentrations and grows best at pH 3, although the other Central European Drosera species are clearly more tolerant to Ca; the growth optimum of D. anglica is at pH 5–7.

3.5  Productivity and Cycling of Water and Nutrients 3.5.1  Productivity Primary production rates in mires can range widely. Peat-forming eutrophic tall-­ sedge marshes and reedbeds are among the most productive habitats in Central Europe, where the above-ground productivity alone can be up to 2500 g m−2 year−1 (Kvet 1971). This is significantly higher than the productivity of the zonal temperate broadleaved forests (see Sect. 4.6.1), due to the unlimited water supply and high nutrient concentrations in these semi-aquatic habitats. At the other end of the scale, raised bog hummock and hollow communities in Central Europe produce around 70–300 g m−2 year−1 (see Table 3.7; further overviews are given in Lindholm and Vasander 1990; Rochefort et  al. 1990 and Gunnarsson 2005). Clymo and Hayward (1982) give typical figures for Sphagnum mires in Great Britain of 800 g m−2 year−1 for hollows, 500 g for lawns and 150 g for hummocks. Gunnarsson (2005) calculated a global average for Sphagnum stands of 370 g m−2 year−1 for hollows, 260 g for lawns and 200 g for hummocks. Moore (1989) also determined the productivity of hollow-inhabiting mosses to be around 100 g m−2 year−1 higher than hummock-inhabiting mosses. However, all these figures apply only to the above-­ ground growth, whereas Vasander (1982) states that the below-ground production accounts for around 40 % of the total production in a Finnish raised bog. With the

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Table 3.7  Above-ground biomass or yield (litter or hay) from various peat-forming, mainly Central European plant communities Vegetation type Caricetum limosae Drepanoclado-Trichophoretum Juncetum squarrosi Molinietum schoenetosum Sphagnum magellanicum lawn Caricetum lasiocarpae typicum Dwarf shrub-Sphagnum-pine bog Erico-Sphagnetum magellanici Caricetum davallianae Fallow Molinietum Juncetum subnodulosi Caricetum gracilis Caricetum elatae typicum Angelico-Cirsietum oleracei Phragmitetum Glycerietum maximae Phragmitetum Typhetum angustifoliae Carici elatae-Alnetum, herb layer

Productivity (g m−2 year−1) 70–90 130–220 100–220 100–250 200 250 180 90–300 100–350 300–500 350–690 550–800 800 900 1614 1500 1700 2592 950

Author Dierßen (1996a) Dierßen (1996a) Klapp (1965) Klapp (1965) Frankl (1996) Warncke-Grüttner (1990) Walter and Breckle (1986) Lütt (1992) Stebler (1897) in Klapp (1965) Warncke-Grüttner (1990) Warncke-Grüttner (1990) Baradziej (1974) Stebler (1898) in Klapp (1965) Klapp (1965) Kvet (1971) Weber (1931) in Klapp (1965) Sieghardt (1973) Kvet (1971) Traczyk (1967)

From a synthesis in Wagner (2000) and other sources Cf. also Table 4.14 and Fig. 4.13

ingrowth core method, Backéus (1990) found that fine roots constituted up to 60 % of the total productivity of vascular plants in bog vegetation dominated by dwarf shrubs. In a Finnish Carex rostrata fen, as much as 88 % of the productivity was below-ground (Saarinen 1996). The light-saturated net photosynthetic rate of Sphagnum species is between 1.3 (S. fuscum) and 3.8 μmol CO2 m−2 s−1 (S. rubellum, S. balticum and S. tenellum). Like other mosses, this is relatively low compared to the photosynthetic rate of vascular plants, and highly dependent on water supply (Rydin and McDonald 1985; Wallén et al. 1988). In Sphagnum, photosynthesis shows rather little variation with temperature but high sensitivity to desiccation (Rydin and Jeglum 2013). A further important factor is excess light, which at 800 μmol m−2 s−1 can cause a long-term reduction in the photosynthetic capacity of S. angustifolium (Murray et al. 1993). Susceptibility to photoinhibition and lack of light adaptation may be linked to the low N contents of Sphagnum leaves. Malmer et al. (1994) recorded Sphagnum growing 10 cm year−1 horizontally in hollows, but only 1–5 cm year−1 in hummocks and in a vertical direction. Overbeck and Happach (1957) found much higher growth rates of over 40  cm year−1

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167

in hollow-­inhabiting species (S. cuspidatum, S. riparium), and up to 12 cm year−1 in hummock species (S. magellanicum, S. rubellum). The cause of this difference in productivity is mainly the better water supply in the hollows, as both hummock- and hollow-inhabiting species had higher photosynthetic and growth rates when transplanted from hummocks into hollows (Johansson &Linder 1980; Wallén et al. 1988, cf. Gunnarsson 2005). Generally, the chemistry of water and peat in the hollows is also more favourable than on the hummocks. The hollows usually have higher pH, microbial nitrogen fixation rates and N content in the peat than the hummocks (Granhall and Selander 1973; Malmer 1988). However, it is also the different physiological properties of the different Sphagnum species that cause these differences, as hollow-inhabiting species have a greater growth potential than those in the hummocks (Hayword & Clymo 1983; Gunnarsson 2005). Sphagnum does not demonstrate any particular annual rhythm in its growth, and can grow all year round under favourable conditions (Kinzel 1982). However, the highest growth rates generally occur in the cool and damp periods in spring and autumn (Wieder 2006). The vascular plants must keep up with the moss growth, which they achieve either by regularly increasing the height over ground of their meristem by one to several centimetres per year (e.g. Drosera or Trichophorum; see Fig. 3.28), or by forming adventitious roots on the shoots that have been overgrown by moss (e.g. Betula nana and the Ericaceae).

Fig. 3.28 (a and b): Growth and embedding of Sphagnum and vascular plants in the surface of mires, semi-schematic (Modified from Grosse-Brauckmann 1962) (a) Eriophorum vaginatum (hare’s tail cottongrass). (b) Vaccinium oxycoccus (cranberry). (c): the profile through a growing raised bog; schematic and at a smaller scale. Modified from Grosse-Brauckmann and Puffe (1967) (see also Grosse-Brauckmann 1986). The growing Sphagnum is green, whilst dying parts turn yellow and lie more or less horizontally. The upper layers are still oxygenated and contain large numbers of aerobic microorganisms that cause a certain amount of humification (producing a brown colour) and mineralisation of the easily decomposable components. At several tens of centimetres depth, the young peat is constantly saturated with water and anoxic and releases hydrogen sulphide

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3.5.2  Peat Accumulation and Decomposition Accumulation  Peat accumulates when rates of plant production are greater than those of microbial decomposition. The net peat accumulation rate of a mire is the difference between peat accumulation in the acrotelm and the decomposition in the catotelm. In growing mires, the surface level of the peat can increase by several millimetres per year. In damp years, mires act as important sinks for carbon, and in dry years they can become sources (Vasander and Kettunen 2006). Average peat growth rates in northeastern Germany are around 0.5 mm year−1 (Couwenberg et al. 2001) but can exceed 10 mm year−1 (Rydin and Jeglum 2013). Clymo et al. (1998) give net peat accumulation rates of 5–17 g C m−2 year−1 for fens and 9–27 g C m−2 year−1 for bogs. Raised bogs in Finland had a mean long-term apparent C accumulation rate of 30–35 g C m−2 year−1 (Turunen 2003). That is only a few percent of the net primary productivity (see Sect. 3.5.1), according to estimates by Päivänen and Vasander (1994) around 2–16 %. According to these figures, fens have higher net primary production rates than raised bogs, but their decomposition rates are even higher (Turunen and Tolonen 1996), which results in lower peat accumulation rates in fens. In Central Europe, peat accumulation is mainly linked to the presence of Sphagnum. However, not all Sphagnum species are able to form bogs, and it would be wrong to assume that all forests, heaths or grasslands containing Sphagnum are incipient bogs. Sphagnum squarrosum, S. palustre and S. girgensohnii, for example, have a similar structure to the bog Sphagnum species, but require higher nutrient levels. On the other hand, bog-forming Sphagnum species are not restricted to raised bogs. Many occur in fens, where peat decomposition is faster so that they do not form the typical domed shape of raised bogs, although only a few Sphagnum species occur at pH values above 5.5. Decomposition  Peat decomposition is mainly carried out by fungi and bacteria, with almost no involvement of animals (Collins et  al. 1978). Decomposition is slowed by lack of oxygen, low temperatures, low pH and low nutrient concentrations. Isotalo (1951) states that the optimum water content for the mineralisation of peat is 60 %. Generally, the speed of decomposition decreases with decreasing nutrient levels and increasing acidity. A further decisive factor is the availability of energy-rich carbon compounds (sugars, proteins etc.) that can be used by the microorganisms, whilst the presence of tannins, resins, waxes and lignin slow decomposition. Verhoeven et  al. (1990) recorded decomposition rates of cellulose in minerotrophic mires 2–5 times higher than those in ombrotrophic mires. Acidic oligotrophic raised bogs therefore have a higher peat accumulation rate than basic mesotrophic or eutrophic mires, although their primary productivity is usually much lower. The peat remains undecomposed for long periods, particularly in the anoxic catotelm of raised bogs. 80 % of the carbon originally present in the catotelm of a mire in southern Sweden was released over a 150-year period (Malmer and Holm 1984). Certain fractions of the peat can remain for over 10,000 years in the catotelm (Göttlich 1990).

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The decomposition rate is closely linked to the proportion of Sphagnum in the dead plant matter. Vascular plant litter is decomposed much faster than that of Sphagnum, which is partly to do with the relatively low N and P contents of the dead moss, as higher nutrient levels accelerate the decomposition (Lee et al. 1993; Aerts et al. 2001). More important, however, is its content of inhibitory substances, such as polyuronic acid, polysaccharides such as sphagnan (Hajek et al. 2011) and high concentrations of polyphenols, including ‘Sphagnum acid’ (p-hydroxy-β-­ carboxymethyl-cinnamic acid; Johnson and Damman 1993; Verhoeven and Liefveld 1997) and other phenolics with antibacterial effects such as p-hydroxyacetophenone (Mellegard et al. 2009). These are almost impossible for phenol-oxidising fungi to break down under anaerobic conditions (Thormann 2006). Particularly pectic acid is important as an ion exchanger in the cell walls of Sphagnum; it binds nutrients in the peat and thus hinders the mineralising activity of the microorganisms (Painter 1991). Fertilisation experiments have clearly shown that the low nitrogen content of peat is not the most important factor in its decomposition. Mineralisation also occurs at very low C:N ratios of 80–100, as long as enough P is present (Damman 1988). The variation in chemical composition of Sphagnum species causes species-­ specific differences in decomposition rates. According to Rochefort et  al. (1990) and Johnson and Damman (1991), this increases along the sequence S. fuscum < S. magellanicum, S. balticum, S. recurvum < S. angustifolium, S. cuspidatum, S. lindbergii, S. fimbriatum. Generally, hollow-inhabiting species are decomposed faster than lawn species, and these faster than hummock species. As a result, peat decomposition is faster in hollows despite their greater water content, as the species contain fewer inhibitory substances.

3.5.3  Water and Nutrient Cycling Water Fluxes  Fens and raised bogs differ considerably in their evapotranspiration (ET) rates. Under unlimited water supply, both fens and raised bogs have ET rates around or slightly above the potential evapotranspiration rate, i.e. similar to a wet grassland (Schmeidl et al. 1970; Mundel 1982; Ingram 1983). The evapotranspiration drops rapidly only when the water supply is very low, and particularly in continental regions it can fall far below the potential ET. Raised bogs dominated by Sphagnum and with few deeper rooting phanerogams can reduce their evapotranspiration suddenly if summer drought causes the water table to fall below 25  cm beneath the surface and the hummocks to dry out. The evapotranspiration experiments of Overbeck and Happach (1957) show that the Sphagnum species differ both in their maximum transpiration and in the degree of reduction of transpiration after sinking of the groundwater. With a good water supply (water table 2 cm below the surface), Sphagnum fallax transpires much more than S. magellanicum and S. rubellum, whilst there was little difference between the species with water at 12 cm below the surface.

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In contrast, fens reduce their ET more gradually with falling water tables, as the fine-pored capillary system in the acrotelm reaches higher than in raised bogs. In addition, the deep-rooted Cyperaceae also maintain the transpiration rates in dry periods. The potential evapotranspiration is maintained until the water table is around 50 cm below the surface (Edom in Succow and Joosten 2001). Particularly high ET rates, which can even exceed the potential evapotranspiration, are achieved by tall-sedge communities, reedbeds and some bog woodlands (Mundel 1982). This is especially the case when isolated mires receive additional energy for evaporation through warm air advection from their surroundings. Lysimeter measurements by Behrendt and Hölzel (1995) show that tall-sedge communities and reedbeds on mire soils in Central Europe can have evapotranspiration rates far in excess of 1000 mm year−1, which is much higher than the precipitation rates (cf. Sect. 4.6.2). These mires therefore do not contribute to groundwater recharge. Neuhäusl (1975) studied the small-scale variability in evapotranspiration in Czech mires. He showed that the transpiration of Sphagnum in raised bogs is relatively low, and the total evapotranspiration, which usually remained below 2 mm d−1, was mainly influenced by the relatively high evaporation in the hollows. Fens and densely forested mires had a higher ET (up to >4 mm d−1), as the transpiration rates of the phanerogams dominating there were high. Ingram (1983) gives maximum evapotranspiration rates for Central European mires of up to 6 mm d−1 (High Fen, Belgium). Eggelsmann (1981, 1990) studied the ET in semi-natural and managed fens and raised bogs in northwestern Germany. According to his figures, the summer evapotranspiration decreases in the sequence alder swamp forest > birch swamp forest > tall-sedge community > Sphagnum raised bog > small-sedge community > meadow or pasture > crop field. This suggests that northwest German mires with high water tables have a 10–15 % higher evapotranspiration than neighbouring vegetation on mineral soil. Increasing forest cover increases the ET and lowers the water level in the mire, as shown by Frankl (1996) in his long-term study of Bavarian mires. Water running off a raised bog follows gullies towards the sloping edges. Relatively little seeps into the catotelm due to the high density of the peat (around 50  mm year−1, Schouwenaars 1994). The hydraulic conductivity of decomposed Sphagnum peat is particularly low, while that of decomposed Carex peat and lignoid (woody) peat is considerably higher (Päivänen 1982). This explains why Eggelsmann (1973) measured inputs into the groundwater of less than 30 mm year−1 in growing raised bogs in northwestern Germany. This is much less than in neighbouring mineral soils, or in fens (30–60 mm year−1) or swamp forests with peat (> 60 mm year−1). A layer of peat clay impermeable to water beneath the peat can further reduce the seepage (Edom in Succow and Joosten 2001; Siegel and Glaser 2006). Due to the high water saturation of the peat, run-off peaks after heavy rainfall can often be higher from raised bogs and fens than from neighbouring mineral soils, as very little additional water can be stored in the peat. On the other hand, summer drought can also lead to a total loss of run-off from raised bogs, but not from fens.

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Nutrient Fluxes  True mires (i.e. not eutrophic tall-sedge, reedbed or swamp forest communities) are usually (very) nutrient poor. However, some fens with high throughput rates of water have a greater nutrient availability. For example, fens surrounded by fertilised pastures receive an N input of around 20 kg N ha−1 year−1 from the groundwater and 1–11 kg N from the surface run-off (Koerselman and Verhoeven 1995). In contrast, biological nitrogen fixation only plays a small role in most mires. Waughman and Bellamy (1980) found this to be around 0.7 kg N ha−1 year−1 in southern German raised bogs, and 5–20 (in exceptional cases up to 60) kg N in fens. Only bog myrtle and alder swamp forests with their symbiotic actinomycetes have a significant additional nitrogen source (Myrica gale: 24–34, Alnus glutinosa: 56–360 kg N ha−1 year−1, Boring et al. 1988; Koerselman and Verhoeven 1992). In raised bogs, atmospheric deposition is the most important external source of nitrogen. Pre-industrial deposition rates in Central Europe would have been around 5 kg N ha−1 year−1, whilst current rates are often 3–8 times higher (see Sect. 3.7.2 in Volume I). Direct absorption of gaseous ammonia (NH3) by the wet vegetation may play a role in mires; Mattson and Koutler-Andersson (1955) estimate this to be over 10 kg N ha−1 year−1. Whatever the source, growing raised bogs are very effective nitrogen sinks, which can bind up to 75 % of the input in their biomass. The net N mineralisation rate is usually low in intact mires due to the high bacterial immobilisation and low N content, and according to Koerselman and Verhoeven (1992) is around 20 kg N ha−1 year−1. Lütke Twenhöven (1992) recorded only 2–10 kg N ha−1 year−1 in a raised bog in Schleswig Holstein. Measurements using incubation in the field gave yearly mineralisation rates between 0 and 40 kg N ha−1 for colline and alpine calcareous small-sedge communities, and 0–2 kg ha−1 for acid small-sedge communities (see Table 3.8). Contrary to expectations, the N mineralisation is apparently sometimes higher in raised bogs than in groundwater-fed fens (Waughman 1980; Koerselman et al. 1993; Bridgeham et al. 1998). Much higher nitrogen mineralisation takes place in drained mire meadows (up to 360 kg N ha−1 year−1) and in forests on fen peat (up to 570 kg N), as they contain large amounts of easily decomposable C compounds (Wild and Pfadenhauer 1997; Succow and Joosten 2001). The mineral N is then usually released as nitrate. N leaching is negligible in intact raised bogs, but can reach high rates in drained bogs (150–300 kg N ha−1 year−1). In contrast to N, leaching of K, Mg and Ca can play a role in both raised bogs and fens, and can lead lower peat layers becoming increasingly nutrient-poor (Malmer 1993). There is next to no denitrification in nitrate-­ poor raised bogs, but it can reach higher rates particularly in drained fens with high NO3− concentrations (up to 110–130 kg N ha−1 year−1, Urban and Eisenreich 1988; Eschner 1989). Phosphorus is often found at higher concentrations in the water in acid mires than in basic ones, in which P is immobilised via precipitation. Oxidation-reduction processes cause P to enter solution as Fe(II) phosphate, particularly at the border between the acrotelm and the catotelm. This contradicts the presumed P limitation of growth of some raised bog species. Waughman (1980) and Verhoeven et  al. (1988) therefore suppose that in the case of P, it is not the availability, but the plant

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Table 3.8  Annual net N mineralisation rate in small-sedge communities and spring mires in northern Switzerland Plant community 1. Small-sedge acid fens  (a) Black sedge fen (Caricetum canescenti-nigrae)  (b) Northern deergrass fen (Tomenthypno-­ Trichophoretum), montane 2. Calcareous fens  (a) Davall’s sedge fen (Caricetum davallianae), near Zurich    as above, in western Switzerland, montane   as above, subalpine (Caricetum ferrugineo-davallianae)  (b) Brown bog rush fen (Schoenetum ferruginei)

Net N mineralisation ratea (kg N ha−1 year−1)

Degree of nitrificationb

0–2

I

0–1

II

0–5

II

0–5 2

II II

0–40

IV–V

From Yerli (1970) in Ellenberg (1977); cf. Fig. 8.50 The higher values refer to stands that have not been mown for a long time, but where the mown material used to be collected and used as livestock bedding; b I = 0–20 % nitrate, V = 80–100 % nitrate a

uptake that is limited. This would also explain the weak relationship between P concentration in the peat or mire water and the species composition of the vegetation. P availability can be higher in fens with low iron and high sulphate concentrations in the water, as SO42− precipitates after reduction as iron sulphide, binding the Fe(II) that had bound phosphate as Fe(III) (Caraco et al. 1989; de Mars 1996).

3.6  Vegetation Dynamics 3.6.1  Quaternary Mire Development All Central European mires developed after the last ice age, so are at most 12,000 years old. Palynological studies of soil profiles in mires show a number of different phases of intensive mire formation, caused by changes in climate, landscape hydrology and vegetation. Figure 3.29 illustrates the development of a raised bog in northwestern Germany since the end of the last ice age.

Fig. 3.29  (continued) forests of the Allerød interstadial (around 10,000–9000 BC), 3 = Preboreal (until around 8000 BC). 4–6 = Holocene climatic optimum: 4 = Boreal stage (around 6800–5500 BC) with rich terrestrialisation vegetation and dense forest; 5 = Atlantic stage (5500–2500 BC), greater warmth and higher rainfall lead to the formation of raised bogs; 6 = Subboreal (around 2500–600 BC), older black peat rich in cottongrass and heather, beginning of the spread of beech. 7–8 = Subatlantic: 7 = formation of white peat and convex dome shape; 8 = sudden end to natural growth through drainage and peat cutting since the seventeenth century

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Fig. 3.29  The development of a raised bog in the northwest German lowlands, semi-schematic. From a coloured illustration by Overbeck. The profiles are exaggerated and the trees are not to scale. 1–3 = Late glacial: 1= the forestless Older Dryas (until around 10,000 BC), 2= birch-pine

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The hydrological mire types presented in Sect. 3.2.5 differ in their average ages and therefore the phases in the Holocene during which they began to develop. Couwenberg et  al. (2001) analysed over 150 mire profiles from northeastern Germany to determine their development patterns. They found that paludification mires were the oldest at around 12,000 years BP (14C years, which are only roughly equivalent to calendar years), followed by terrestrialisation mires at around 11,500 years BP. Kettle mires and spring mires were first found from around 10,000 14C years BP. The other types of mire are younger, as they are secondary mire formations or developed during the somewhat warmer and more humid climate of the Atlantic period (around 7500–4500 BP). The first percolation mires appeared approximately 9000 14C years BP, and often developed from terrestrialisation mires (see Fig.  3.30). True raised bogs were only found 7500 BP.  This emphasises the importance of sufficient warmth and high rainfall for the development of Sphagnum raised bogs, which was the case during the somewhat warmer Atlantic period, when it was around 1 °C warmer in summer than today (see Fig.  3.5 in Volume I). The climate in northwestern Germany became more oceanic around 6000 BC as the sea level reached roughly its current level, thereby promoting the formation of mires (Bittmann 2004). Before this, only fen and swamp forest peat and gyttja had formed (see Fig. 3.29). Since the Middle Ages, drainage has reduced peat accumulation in many kettle and percolation mires and raised bogs. Pop (1964) states that, at least in Europe, Sphagnum raised bogs first formed during the post-glacial period, as there are no peat remains from the interglacials, the Tertiary or even older periods that clearly belong to this oligotrophic bog type. During the Atlantic period, raised bogs formed at higher elevations and much further north than is the case today. These subalpine and alpine or subpolar mires are no longer growing, and are continuously degraded by frost, snow pressure, water and wind (see Sect. 5.3.10.1), so should be considered as fossil or subfossil formations. Some mountain sloping mires only formed 2000–3000 years ago and thus are even younger than the raised bogs (see Fig. 3.30). The formation of paludification and sloping mires was particularly promoted by increased run-off over the last 1000 years due to human forest clearance. However, this is not the case for the sloping mires of mountains with higher rainfall like the Harz. According to Beug et  al. (1999), mire formation here began at least 10,000 14C years ago (around 11,600 BP) at the end of the Younger Dryas, and has continued uninterrupted until today. However, new mires only formed in the Harz Mountains until the end of the Subboreal (around 2800 BP). Periods of greater mire growth (e.g. in the late Subboreal) occurred during colder periods. The youngest mire formations are the salt marsh peatlands on the Baltic coast, which mostly formed less than 1000 14C years ago with rising sea levels. Historical studies of mires show that some mire formations are rather short-­ lived. In northeastern Germany, most paludification and terrestrialisation mires persist for less than 2000 14C years, and 60 % for only up to 1000 years (Couwenberg et  al. 2001). Some of these mires then transform into raised bogs or percolation mires.

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Fig. 3.30  Development and sequence of deposition in the different hydrological mire types, the distribution of which is shown in Fig. 3.14. (Modified from Succow (1988). Further information in the text for Fig.  3.14). (1) Terrestrialisation mires develop during vegetation succession from reedbed to swamp forest. (2) Paludification mires form due to the increasing moisture content of the soil, usually due to a rise in groundwater levels; in river valleys, these are in some cases flood mires (7). (3) Percolation mires can develop from 1, 2 or 7 if there is a constant flow of water through them, often sloping away from a river on both sides. (4) Kettle mires form in depressions without outflows (e.g. in young drift moraine landscapes) from 1 or 2; if the water supply is mainly from the rain water then they can transform into 5. (5) Lowland raised bogs usually form from 1 if the climate is humid enough, but also from 2 via the growth of Sphagnum. (6) Upland raised bogs develop in the same way from 1 or 7. (7) Sloping mires form on sloping mineral soils due to a constant flow of surface water. (8) Coastal flood mires are a special case; they develop in a similar way to the paludification mires with frequent flooding, e.g. through the trapping of water after a high tide. After the dyking of the North Sea, they either changed their character or were lost. Several large mires transformed into 5. (9) Spring fens develop in the groundwater emerging under pressure from almost level or sloping mineral soils. (10). The 10 symbols refer to Fig. 3.14. Types 1, 2, 8 and 9 are primary, Types 3–6 (and to some extent 7) are secondary. Type 9 only occurs in small areas but can be found in almost all landscapes. Almost all hydrological mire types are no longer in their natural states in Central Europe today

In a cool, temperate and humid climate, any increase in soil water can create raised bogs by paludification (see Fig. 3.1), provided that the water is still and base-­ poor. Even calcareous fens and fast-growing forests on nutrient-rich mineral soils are eventually overgrown by raised bogs under these conditions (see Fig. 3.29). In

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drier climates such as in Brandenburg and Poland, raised bogs are formed by terrestrialisation, but only from acid and humus-rich (dystrophic) still water bodies. The formation of raised bogs can therefore be seen as a mainly a­ utogenic process of succession, whereby the water regime is fundamentally changed. For the majority of large raised bogs, there is no longer any visible evidence of how they formed, and only palynological studies of peat cores allow the development to be reconstructed.

3.6.2  Recent Developmental Processes Hollows and hummocks, or in their extended form flarks and strings, are the result of the dynamic processes of mire growth, which are stronger in oceanic climates than in warmer and summer-drier landscapes (Lindsay et al. 1985). This microrelief is probably primarily the result of small-scale differences in water levels within the mire, whereby a mosaic of peat-forming mire plants with different flood- and drought-tolerance develops (Dierßen and Dierßen 2001). Sinking water levels promote more drought-tolerant hummock-inhabiting mosses such as Sphagnum fuscum and S. rubellum, thereby supporting the growth of the hummock until the increasingly dry conditions force it to stop growing. At this point, the increased oxygen levels in the peat start accelerating its decomposition. In contrast, increasing water levels accelerate the expansion of hollows. These support plankton growth in the summer, which release oxygen into the water of the hollow, thereby increasing the peat decomposition rates at the edge of the hollows (Foster et  al. 1988; Svensson 1988). When there are periods of alternating high and low water levels, sloping areas develop transverse string and flark structures, which are characteristic for the widespread aapa mires in Scandinavian mountain areas. Once initiated, this mosaic structure is then intensified by the different production and decomposition rates in hummocks and hollows. The Sphagnum species of hollows are usually more productive than those in the hummocks (see Sect. 3.5.1), but this is more than compensated for by faster decomposition of Sphagnum litter in the hollows (Johnson and Damman 1993). The hummock and hollow complex is therefore mainly caused by the different decomposition rates of the Sphagnum species in these two sub-habitats, and not through the greater production of hummock mosses (cf. Sect. 3.5.2). The previously widespread view that mire hummocks and hollows not only alternate in space but also in time has been shown to be false in almost all studies (Walker and Walker 1961; Jensen 1961; Eurola 1962; Casparie 1969; Overbeck 1975; Svensson 1988). A hummock can remain so for centuries, as can a hollow, so that a mosaic of both forms grows upwards either evenly or unevenly (Rydin et al. 2006; see Fig.  3.31). Rydin and Barber (2001) showed with 14C dating that Sphagnum clones persisted in an English bog in the same square decimetre on average over 400 years, and in one case >1600 years. Nevertheless, cyclical succession can occur in certain places and over certain periods, with repeated changes from hummocks to

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Fig. 3.31  Stages of raised bog development in northwestern Central Europe illustrated using the example of a well-analysed 50 m section of bog near Emmen (The Netherlands). The growth rates vary and do not follow a cyclical alternation between hollows and hummocks (Modified from Casparie 1969). To illustrate the temporal sequence, the same profile is shown at intervals of 640– 730 years (intermediate stages were omitted). The oldest, strongly decomposed Sphagnum peat began to be covered by lighter S. cuspidatum hollow peat around 1550 BC.  Other profiles are dominated from bottom to top by strongly decomposed heather or cottongrass peat, until the area became much wetter around 600 BC.  This led to widespread and rapid covering with weakly decomposed and relatively light-coloured Sphagnum peat (so-called white peat). The layers formed after 100 AD are not shown as they had been disturbed. There are no permanent, deep, water-filled hollows such as are characteristic for kermi bogs (cf. Figs. 3.8, 3.9 and 3.10)

hollows at the same spot, e.g. in the Varrassuo raised bog in southern Finland during part of the Atlantic period (Tolonen 1966). Cyclical succession of hummocks and hollows is the exception rather than the rule in mires, and once formed, these structures are generally very stable (Dierßen 1982). On a shorter time scale, certain weather conditions can lead to surprisingly large inter-annual population fluctua-

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tions in Sphagnum carpets, as was observed e.g. in southern Sweden (Gunnarson and Flodin 2007). Cross-sections of raised bogs and flat bogs often display large areas alternating between hollow-like ‘regeneration phases’, i.e. periods of rapid moss growth and low decomposition rates, and ‘maturation phases’, in which the mire surface remains relatively stable and the peat decomposition is faster (see Fig. 3.31). Casparie (1969) considers these long-term dynamics to be due to local changes in the surface run-­ off, i.e. temporal fluctuations in the water regime. The changes in the degree of humification do not, therefore, necessarily have to be caused by the climate (see also Overbeck 1961, 1975). A cyclical alternation between a wet peat growth phase with few trees and a drying phase less hostile to trees has also been postulated for the bog woodlands of eastern Central Europe (Kulczynski 1949). In eastern German kettle mires, Müller-­ Stoll and Gruhl (1959); Landgraf and Notni (2003) and Timmermann (2003) found evidence of irregular changes in the vegetation caused by natural changes in the water level. Pine and birch colonised the quaking bogs of the Eriophoro-Sphagnetum recurvi after a period of drying out. The trees were later killed by windthrow, rewetting, or the fact that they became too heavy to be supported by the floating quaking bog.

3.6.3  Primary Succession in Growing Raised Bogs Raised bogs can form from mineral soils or fens if the water supply increases or evapotranspiration/run-off decreases. High soil moisture promotes Sphagnum over vascular plants. The transition from a groundwater-fed fen to a rainwater-fed bog can occur in less than 30 years (van Diggelen et al. 1996). This is usually caused by a change in climate to more humid conditions, but it can also be initiated by a dry period during which Eriophorum dominates and deposits strongly humified peat, causing water to accumulate and thereby promoting the growth of Sphagnum (Hughes and Barber 2004). Increased Sphagnum growth then triggers autogenic succession towards a deep raised bog with oligotraphent (nutrient-poor) vegetation, which itself leads to a progressive reduction in nutrient availability (‘meiotrophication’, Sjörs 1983) in the peat. The accumulating Sphagnum litter not only decreases the decomposition rate, but also causes acidification. A pioneer of autogenic acidification is Sphagnum subsecundum (Kuhry et  al. 1993). The slow decomposition causes peat accumulation, and its organic matter stores a large proportion of the nutrients present, whilst bacteria immobilise most of the available nitrogen (Limpens et al. 2006). The growing peat layer also increasingly isolates the vegetation from the mineral soil, leading to the stage where it becomes purely rain-­ fed. Sphagnum is well adapted to these conditions, but few vascular plants are (Zobel 1988). The chemical composition of its cell walls and the slow decomposition rate allows Sphagnum species to produce conditions where it faces relatively little competition, allowing its own mass development. As is the case in

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many other successional sequences, this is directional succession with a clearly defined end stage, yet this can be achieved via a number of different succession routes (Sjörs 1980).

3.6.4  Secondary Succession After Mire Drainage The majority of bog woodlands in eastern Central Europe now contain so little Sphagnum that it is surprising that such a thick peat layer could develop at all. One potential cause of this is the recorded decrease in summer precipitation over the last 100 years by up to 50 mm in this region (Schönwiese et al. 2003; see Fig. 3.63 in Volume I), which may have led to increased drying of the peat and a decreased Sphagnum growth. However, a more important cause is likely to be human influence, whereby many modern bog woodlands were drained to promote the growth of the trees at the cost of the Sphagnum (Steffen 1931; Lötschert 1964; Overbeck 1975). Bog woodlands and true raised bogs often occur alongside one another in eastern Central Europe, i.e. under the same climate conditions, both in coastal and in more continental areas. Many raised bogs in western Central Europe and the foothills of the Alps, which had been completely treeless until a few decades ago, now support vegetation similar to the eastern bog woodlands. This is doubtless the result of drainage, which has stopped the growth of nearly all lowland raised bogs. Succession now also occurs in many fens that had been used as agricultural land until a few decades ago, but were abandoned for economic reasons. Eutrophic to polytrophic reed and tall-sedge communities form that finally develop into woody vegetation. Jensen and Schrautzer (1999) analysed the succession on an abandoned fen in Schleswig-Holstein. Succow and Joosten (2001) describe the succession in northeastern German mires depending on nutrient levels and human influence.

3.7  Human Influence 3.7.1  Exploitation by Peat Cutting, Drainage and Cultivation Humans have strongly influenced mire distribution and habitat conditions since the widespread forest clearances of the Middle Ages. At first, this was mainly an unintentional facilitation of peat growth by increasing surface run-off, improving the conditions for mire formation (see Sect. 3.6.1), but later mires were cultivated, drained and even completely eliminated. Peat cutting for burning is known to have occurred in Central Europe since the Bronze Age, whilst other uses such as the mining of bog iron ore or of salt in coastal areas have been practiced since the Iron Age and more intensively since the Middle Ages. Peat cutting accelerated in Germany around 1750, as the demand for wood grew with the industrial expansion and peat

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became more valuable for fuel. In the following 200 years, the large peat deposits in raised bogs and flood mires in the southern part of the North Sea coastal region, the northeastern German and Polish lowlands, and the foothills of the Alps, were all harvested, at first by hand and then by machines from the twentieth century onwards. Today, industrial peat cutting is mainly carried out to supply horticulture and agriculture. More intensive agricultural use of mires after drainage began on a large scale only towards the end of the nineteenth century. This was at first mostly damp mire meadows that were cut once or twice a year, and later transformed by fertilisation and ploughing into intensive grasslands (Radlmair et  al. 1999; Dierßen and Dierßen 2001; Succow and Joosten 2001). Most mire grasslands in Germany are now grazed. As a result of their widespread exploitation, Central Europe has now lost around 95 % of its original intact mire area (see Table 3.9). The Netherlands and Lower Saxony were once particularly rich in mires, with huge areas of raised bogs, but these have now been reduced to small fragments by cultivation and/or drainage. The Stapeler Moor in East Frisia once covered an area of 69 km2, but by 2007 only 2 % of it remained uncut. The number of typical mire vascular plants had decreased here by 44 % since 1936, and the number of lichens by 83 % since 1968 (Huntke 2008). The total number of higher plants and bryophytes remained stable, however, as increasing numbers of non-mire species established in the area. To find a typical patch of this once widespread habitat today, it is necessary to travel up into the mountains such as the Harz, the High Fen, the Black Forest and the Krkonoše Mountains, or remote areas of the foothills of the Alps. However, the barren expanses of lowland raised bogs and their silent, lonely beauty can only truly be imagined having experienced the large mires in the subboreal conifer regions in

Table 3.9  Former and current mire areas in Central Europe (areas with peat ≥ 30 cm thick and organic contents of ≥ 30 %)

Country Netherlands Belgium Germany Switzerland Austria Czech Rep. Slovakia Poland Central Europea

Land surface area (× 1000 km2) 42 31 357 41 84 79 49 313 ca. 1000

Former Former proportion mire area of mires in (× 1000 km2) land area (%) 15 36 1 3.3 15 4.2 2 4.9 3 3.6 0.3 0.4 0.1 0.2 13 4.2 ca. 50 ca. 5

From Succow and Joosten (2001) a Incl. Luxembourg

Recent area of growing mires (× 1000 km2) 0.15 0.01 0.15 0.2 0.3 0.015 0.005 1.95 ca. 2.8

Remaining mire area (% of former area) 1 1 1 10 10 5 5 15 ca. 5

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Estonia, Finland or Sweden. Such untouched expanses of mire were still present during the lifetime of Grisebach (1846). He writes ‘at the border between Hanover and Holland, crossing the pathless mire from Bourtange between Hesepertwist and Ruetenbrock, I visited a spot where, like on the high seas, the level horizon formed a circle around me, and no tree, no shrub, no hut and no object higher than a child could been distinguished in the apparently endless wasteland.’ The commercial exploitation of coastal raised bogs began in the Netherlands as early as the sixteenth century, mainly at first to provide peat for fuel for towns and industry. A canal was built for the peat barges, acting at the same time as the main channel receiving the water from the drainage ditches and representing the central waterway of the community who farmed the areas where peat had been cut. Such pioneer farming communities (‘Fehnkolonien’) also colonised peatland in northwestern Germany (see Ellenberg 1990). Mires far from larger settlements were generally safe from peat cutting, as the transport costs were too high. Low-intensity swidden agriculture developed here, which mainly consisted of burning the surface of partially drained mires and sowing buckwheat or other crops adapted to low nutrient conditions in the ashes. This form of exploitation destroyed many large raised bogs, as their height above the surrounding area allowed them to be easily drained. The repeated burning prevented forest from developing, so that the mires remained as treeless expanses. Peat cutting was only practised close to larger villages for their own use (see Fig. 3.32). However, in the nineteenth and twentieth centuries, these unproductive cultivation types were replaced by the planned German raised bog cultivation, which still determines the character of the majority of northwest German mires. Small settlements were established on raised bogs that had been drained but not cut for peat, and the farmers cultivated the peat using lime and mineral fertiliser to produce relatively high crop yields. Lime is no longer used, as it accelerates the decomposition of the peat and leads to a poor soil structure. The houses of the old mire colonies along the roads are now surrounded by tall birch, oak and pine, so that the former raised bogs are now only visible in the undulating surface caused by the irregular sagging of the peat. Many cultivated mires no longer contain any remains of the former mire vegetation. Only some areas where peat has been cut develop into hollow-like stages rich in Eriophorum vaginatum and some Sphagnum magellanicum hollows. The old surface of the mire develops heathland vegetation several years after drainage, so becomes covered in hummock-like vegetation or even with a pure Calluna heath. Fire promotes the spread of Molinia caerulea in these wet heaths, particularly in places with phosphorus limitation (Olde Venterink 2000; see Fig.  3.33). This is because Molinia can achieve much higher growth rates than other mire phanerogams using the high nutrient concentration in the ash (van Vuuren 1992). Drainage and subsequent intensive agricultural use of former mires fundamentally changes the character of this ecosystem. It causes a significant decrease not only in the diversity of the typical mire plants and animals (Succow 1988; Bootsma 2000), but the mire also loses most of the functions that it once performed in the landscape (see Sect. 3.7.3). If mires are exploited, subsidence and shrinking, i.e.

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Fig. 3.32  Cut peat in a raised bog in Upper Bavaria. The upper layer of white peat is thrown back into the empty pit, whilst the more decomposed black peat that is used for fuel is cut into square sods. These are laid out to dry (right) and then stacked (back right). The black peat is almost impermeable to water, so that it takes days for the pit to fill with water. A drained bog with Molinia and birch community is visible in the background

irreversible compaction of the organic matter caused by the loss of water, and oxidative decomposition of peat leads to a decrease in the depth of peat and large carbon losses to the atmosphere. This shrinkage is much greater in easily mineralisable fens than in raised bogs. Succow and Joosten (2001) quote an average loss rate of 5–10 mm year−1 for grassland use, and 12–20 mm year−1 for arable use in German mires. The loss of peat is faster in dry and summer-warm climates. The degradation of peat results not only in fundamental changes to the hydrological and chemical properties of the substrate, but also influences the water regime (see Fig. 3.33). If the mire has sunk by several tens of centimetres or even metres, then it will need draining again to remain useable. On the other hand, many former mires, particularly in eastern Germany, now need irrigation in summer as the water storage capacity of the degraded peat is now much lower (Succow and Joosten 2001). Particularly large losses have been suffered not only by plant communities of raised bogs, but also by those of calcareous fens outside of the Alps. These produce low hay yields that can only be used for animal bedding, so have mostly been drained and transformed into high-yielding hay or silage meadows.

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Fig. 3.33  Semi-schematic diagram of increasing levels of drainage in a northwestern German raised bog (Modified from Ellenberg 1954a). Even a minor reduction in the groundwater level within the mire is enough to reduce the growth of Sphagnum and promote the growth of dwarf shrubs and other hummock plants (left). If the water level sinks further, then agricultural use becomes possible. If the mire is left undisturbed, then Calluna and other heathland species spread and trees can establish. Molinia caerulea can dominate if the soil is wet enough. The peat shrinks and subsides depending on the degree of drainage

3.7.2  Eutrophication Koerselman and Verhoeven (1995) distinguish external and internal eutrophication of mires. External eutrophication refers to nutrient inputs from outside the system, whilst internal eutrophication describes the nutrient circulation within the system (e.g. of P) through accelerated mineralisation or increased desorption from mineral surfaces. Ombrotrophic bogs in particular are highly sensitive to external atmospheric deposition of nutrients and pollutants. The critical load for raised bogs is considered to be an N deposition rate of up to 5–10 kg N ha−1 year−1 (Tomassen et al. 2004; Huntke 2008). If the nitrogen deposition increases, then the N content of the Sphagnum biomass increases (Lee et al. 1993; Malmer 1993). This causes the decomposition rate of the dead moss to increase, accelerating the N turnover and promoting the growth particularly of vascular plants, e.g. Vaccinium oxycoccos (J. Tüxen 1983), Dryopteris carthusiana, Molinia caerulea and Deschampsia flexuosa (Aaby 1994; Baumann 2009), Trichophorum cespitosum (Baumann 2009) or the Rhynchospora species (Heijmans et  al. 2002). N deposition is also the likely cause of recent invasions of spruce or downy birch in raised bogs of the Harz Mountains (Baumann 2009) and the Netherlands (Tomassen et al. 2004). Malmer and Wallén (2004) assume that increased N inputs can reduce peat growth as the ombrotrophic Sphagnum species are weakened.

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In the long term, atmospheric N deposition causes profound changes to the species composition and nutrient turnover of Central European mires, similar to other Central European ecosystems. Whilst low N inputs initially increase the growth of Sphagnum, larger amounts reduce its growth (Limpens et  al. 2006) and make it easier for it to be outcompeted by vascular plants and nitrophytic mosses. In Poland and England, Lee et al. (1993) found that N deposition increased the competitive ability of the minerotrophic Sphagnum recurvum, whilst in northwest German and Dutch mires, it is increasingly S. fallax, S. squarrosum or Polytrichum commune that replaces S. magellanicum and S. rubellum (J. Tüxen 1983; Lütke Twenhöven 1992; Lütt 1992; Kooijman and Bakker 1995; Bootsma 2000; Dierßen and Dierßen 2001). This is caused partly by a shift in the competitive relationships, as different moss species profit to different degrees from ammonium and nitrate (Paulissen et al. 2004). Partly, however, it is the toxic effect of the deposited ammonium, to which S. magellanicum is particularly sensitive (Rudolph and Voigt 1986; Limpens et al. 2004). According to a permanent plot study by Hajkova et al. (2011) in montane raised bogs in the Sudeten Mountains (Czech Republic), 14–17 years of high nitrogen (up to 30 kg N ha−1 year−1) and sulphate deposition did not lead to significant change in the species composition. Over longer times spans (25–60 years), however, significant increases in the frequency of Sphagnum recurvum agg., Deschampsia flexuosa, Trientalis europaea, Eriophorum angustifolium and E. vaginatum at the expense of Carex limosa, Scheuchzeria palustris and S. capillifolium were recorded. The authors suggested that deposition-driven vegetation change is slow in natural bogs as long as the water regime is not altered.

3.7.3  A  ccumulation of Pollutants and Exchange of Trace Gases with the Atmosphere Raised bogs in regions of England strongly affected by sulphur dioxide now only support minerotrophic Sphagnum species such as S. recurvum. Lee et  al. (1993) state that this is because Sphagnum species dependent on groundwater are more tolerant of SO2 than ombrotrophic species. This tolerance seems to depend on relatively high iron contents in the cell walls of the mosses. Acid deposition and the oxidation of sulphide to sulphate during mire drainage promoted acid-tolerant plants such as Eriophorum angustifolium, Ranunculus flammula and Hydrocotyle vulgaris in base-rich mesotrophic mires in the Netherlands in place of acid-sensitive species such as Scorpidium scorpidioides (de Mars 1996; Bootsma 2000). Deposition also led to a steep decline in populations of sensitive Cladonia species in Danish mires (Aaby 1994). The direct uptake of nutrients over the plant surfaces increases the sensitivity of bryophytes to heavy metals that are readily accumulated. Sphagnum is a particularly good indicator of heavy metal contamination. Wandtner (1981) used analyses of Sphagnum from western German raised bogs to estimate the regional deposition

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of cadmium, copper, lead and manganese. Historical changes in emissions can be measured in the peat profile. Profiles from mires in the Harz Mountains showed evidence of ore smelting in the Middle Ages, which were linked with the release of extremely high concentrations of heavy metals into the atmosphere through dust particles (Görres et al. 1995). Growing mires are sinks not only for carbon, but also for nitrogen and other nutrients, which are bound in the peat and thus reduce the eutrophication of the surrounding water bodies. Drained mires largely lose this important immobilising function and usually turn into sources. Drainage accelerates the mineralisation and thus the release of nitrogen, so that cultivated grasslands on former mires often have high N concentrations in their seepage water, especially when they have been additionally fertilised (Behrendt et al. 1994). In the Rote Moor in the Upper Harz, which has been partially drained and affected by acid deposition, Müller and Bauche (1998) found that the peat acted as a source for all nutrients and pollutants that they measured (except N), i.e. it had already lost its sink function. The accelerated mineralisation following drainage also enhances the release of greenhouse gases, notably carbon dioxide and nitrous oxide. In contrast to growing Central European fens, which have an average CO2 fixation of around 170 kg CO2-C ha−1 year−1 (Armentano and Menges 1986), drained fens emit several times this amount in the course of peat degradation. Mundel (1976) recorded the release of 2800–6700 kg CO2-C ha−1 year−1, depending on the groundwater level. The release of nitrous oxide from a northeast German fen also increased considerably after the lowering of the water table (Augustin et al. 1996). N2O emissions of 16–20 kg N ha−1 year−1 were measured from a grazed acid fen (Franken et al. 1992; cf. also Sect. 9.6.2 in Volume I). The emission of methane is fairly low in raised bogs with slow peat decomposition in the catotelm. This is also the case for drained mires where oxygen reaches the peat (< 4 kg CH4-C ha−1 year−1, Augustin in Succow and Joosten 2001), while high emissions (up to 521 kg CH4-C ha−1 year−1) were measured in the first years after rewetting fens in the course of restoration.

3.7.4  Conservation and Restoration of Mires Drained mires cannot simply be restored by increasing the water levels again (Eigner and Schmatzler 1991; Schwaar 1991; de Mars 1996; Rochefort and Lode 2006). The evidence collected from raised bog restoration from northwestern and northeastern Germany, the Netherlands and the foothills of the Alps shows the considerable difficulties involved in this undertaking (Joosten 1992, 1993, Succow 1998; Joosten and Clarke 2002). Rewetting of large areas of raised bog can prevent further losses of peat, but the restoration of the typical peat-forming mire vegetation is usually not possible within a time frame of a few decades (Dierßen 1996a). The water regime has often changed so drastically in mires where peat cutting has taken place that the blockage of the drainage ditches alone is not sufficient to permanently rehydrate them. The upper layers of peat that have oxidised must in some cases be

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removed to prevent the colonisation of eutrophic reed species (de Mars 1996). A further problem is the provision of nutrient-poor surface water to prevent further eutrophication. This can, for example, be achieved by storing winter precipitation in collecting ponds before allowing it to drain into the mire in summer. Restoration is usually only successful if buffer zones are created around the mire where no fertiliser is applied, both in the surrounding agricultural land and in the region supplying the groundwater to the mire (Lütt 1992). The buffer zones should be at least 80 m wide in lowland raised bogs (Eggelsmann 1982) and at least 200–500 m wide in mountain areas (Kaule 1974; Grünig et al. 1986). The recolonisation of bare peat by mire species also proved surprisingly difficult. Carex rostrata and Eriophorum species could be made to establish after sowing seeds and providing an initial P or PK fertilisation (Sliva et al. 2000). Mires where peat has been cut by hand are easier to restore, at least in small areas (Bertram 1988; Pfadenhauer 1989; Poschlod 1990 and Lütt 1992). In the relatively small cuts made by farmers in the 1930s to 1950s (see Fig. 3.34), the water supply remained sufficient for the regrowth of the mire. This is because the black peat, which maintains the water level in the mire, was not cut down to the water-­permeable sand layer beneath. In addition, regeneration was improved by the presence of the uppermost white peat (which was thrown back into the water-filled hole as it had little value as fuel), as these apparently dead Sphagnum remains, which had been in the ground sometimes for over 30 years, sprouted again and continued to grow. Even fragments of growing Sphagnum that fell into the hole where the peat was cut often continued to grow. The variation in microrelief in the permanently water-filled pits promoted the growth of both hollow- and hummock-inhabiting mosses (see Fig. 3.34), so that important species such as S. cuspidatum, S. magellanicum and S. fallax remained present in the system. The climate in these small, sometimes steep-­ sided pits is apparently also favourable for the regeneration of Sphagnum communities, as the dark peat walls are heated by the sun and also protect the vegetation from the wind. Widespread agricultural intensification means that calcareous fens are found today mainly as small fragments. Habitat fragmentation has reduced the vitality and genotypic diversity of the remaining populations of many (but not all) typical species, such as Carex davalliana and Tofieldia calyculata, which could lead to the extinction of small stands (Hooftman and Diemer 2003). The removal of nutrients and rewetting of previously drained and intensively used fens is also problematic. Increasing the water level reduces the nitrogen availability, due to slower mineralisation and greater denitrification, but also increases the phosphorus availability due to the reducing conditions (Olde Venterink et  al. 2002). Removal of topsoil can reduce the amount of nitrogen in the soil and simultaneously produce a small-scale mosaic of habitat types (Sliva et  al. 2000). The removal of P from the system is best achieved by mowing and removal of cut grass (Koerselman and Verhoeven 1995; de Mars 1996). Succow (1998) and Succow and Joosten (2001) provide numerous practical examples of mire restoration from northeastern Germany, as does BAFU (2009) from Swiss mires.

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Fig. 3.34  A schematic cross-section through a cutover bog from which peat has been cut by hand, top: during continuing work in stages, bottom: after cutting has ceased (Modified from Pfadenhauer 1989). Partitioning walls are left so that peat can be cut from a dry pit the next year (the groundwater within the mire seeps only slowly, see Fig. 3.32). The upper white peat has little value as a fuel and is disposed of in the empty pits, and peat containing tree roots is also rejected as it cannot be cut into regular peat sods. The cut sods are dried by first laying them out individually and then stacking them loosely. The extracted pit provides wet to relatively dry conditions for further Sphagnum growth, i.e. favourable conditions for the development of a range of raised bog communities

Chapter 4

Vegetation of Freshwater Habitats

This chapter will give a comprehensive overview of the vegetation ecology of lakes and ponds, as well as springs, streams and rivers. The ecology of these systems is the focus of the field of limnology, one of the oldest branches of ecology. Here, we will concentrate on the aquatic and semi-aquatic macrophytes and their ecology, including flowering plants, ferns, bryophytes and stoneworts (Characeae) as macroscopic primary produces in still and flowing water (see Fig.  4.1). More detailed works on the ecology of freshwater habitats can be found in the limnological literature (e.g. Allan 1995; Brehm and Meijering 1996; Brönmark and Hansson 2005; Giller and Malmquist 1998; Lampert and Sommer 1999; Schwoerbel 1999 and Wetzel 2001).

4.1  Freshwater Macrophytes and Their Origins The majority of aquatic primary producers are only 0.01–1 mm in size. In Central Europe, this includes several hundred species of unicellular or multicellular ­planktonic or periphytic (i.e. drifting or attached to substrates) freshwater algae in the families of the Chrysophyceae, Bacillariophyceae (diatoms), Chlorophyceae, Euglenophyceae, Pyrrhophyceae (dinoflagellates), Xantophyceae and Cyanophyceae (cyanobacteria). In the relatively well-studied Lake Neusiedl, for example, over 600 species of algae and 60 species of cyanobacteria have been recorded (Löffler 1979). Further important primary producers include species of picoplankton (0.2–2 μm), especially the photoautotrophic cyanobacteria, the contribution of which to the primary production of lakes has only relatively recently been recognised (Stockner 1991). The diversity of aquatic macrophytes is, in contrast, much smaller. In Central Europe, there are around 200 to 300 species of flowering plants and ferns that are found in aquatic or swamp habitats, depending on how broadly the term is defined, and fewer than 100 bryophyte and 30 Characeae species. Comparing the number of © Springer International Publishing Switzerland 2017 C. Leuschner, H. Ellenberg, Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats, DOI 10.1007/978-3-319-43048-5_4

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Fig. 4.1 The terrestrialisation of a eutrophic lake in the Pleistocene lowlands of the Müritz region (northeast Germany). The floating-­ leaved plant dominating the surface of the lake is Nuphar lutea. In the reed belt, Typha angustifolia indicates eutrophic conditions. Phragmites australis is visible on the left of the photograph. Butomus umbellatus is flowering in pink. Alder swamp forest is the terminal stage of the terrestrialisation process

species in acidic and neutral to basic waters, it is clear that the majority of algal groups are more species-rich in neutral to alkaline water; this is particularly the case for diatoms and the Characeae. In the region of Bremen, for example, Behre (1956) found three times the number of algal species in lakes with water of pH 7–9 compared to those with pH 4–5 (see Fig. 4.2). The dominance of basiphytic over acidophytic plant species thus applies to both aquatic and terrestrial habitats in Central Europe (cf. Ewald 2003). Whilst only relatively few species of flowering plants and ferns have colonised freshwater habitats, these stem from a large number of families. The terrestrial origins of numerous freshwater macrophytes can be detected in relictual features such as the possession of stomata and a cuticle, which are of little functional use in submerged leaves.

4.2  Environmental Conditions and Habitat Classification Due to the unique physical and chemical characteristics of water, aquatic habitats present certain challenges for the higher plants that colonise them. They are briefly summerised in the following.

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191

Fig. 4.2  In the lakes around Bremen, the number of species decreases with the pH of the water in most groups of algae. Particularly the number of diatoms drops consistently from the alkaline Zwischenahner Meer to the highly acidic Bullenseen (Modified from Behre 1956)

4.2.1  Physical Characteristics Water is densest at 4 °C, i.e. above freezing point. This density anomaly means that water can freeze at the surface, protecting lower water layers as well as the substrate at the bottom from freezing. Higher plants can therefore produce shoots in spring from buds in the sediment, even if the ice at the surface has largely destroyed any biomass remaining from the previous summer. The strong polar nature of water molecules means that flowing water experiences internal friction (viscosity). This friction has to be overcome in order for the water to move. As the viscosity of water at 0 °C is around twice as high as at 25 °C (see Table 4.1), planktonic organisms sink twice as fast in warm water than in cold. The force of moving water against static plants in rivers is similarly increased at higher temperature. The high viscosity of cold water reduces the water uptake and movement within the plant when temperatures are low, as it does in transpiring terrestrial plants during the cold season (cf. Sect. 4.6.1 in Volume I). These properties largely explain why cold waters have a much lower primary productivity than warm water bodies. Flowing water removes particles of sediment and stone from the river bed and transports these downstream, where they are deposited in areas with lower flow

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Table 4.1  Some biologically important characteristics of water Property Density

Specific heat Viscosity Diffusivity of CO2 Light extinction Red light Blue light

Temperature (°C) 0 4 25 15 5 25 25

Value 0.999 1.000 0.997 4186 1.561 0.890 1.7∙10−9

Unit kg l−1 kg l−1 kg l−1 kJ kg−1 K−1 Pa s ∙10−3 Pa s ∙10−3 m2 s−1

65 0.5

% m−1 % m−1

velocities. Whether particles are eroded, transported or deposited depends on the speed of the water and the size of the particle (see Fig. 4.3), as well as their form. This determines the type of sediment and its nutrient content. The increasing hydrostatic pressure with greater water depth influences the depth distribution of aquatic plants, as it hinders the elongation of leaves and shoots, as well as the activity of the photosynthetic apparatus and the gas transport within the plant (Gessner 1961; Golubic 1963; Wetzel 2001). The majority of freshwater macrophytes grow in water less than 10 m deep, where the water pressure is less than 0.1  MPa. Areas deeper than this (until around 40 m) are dominated by the Characeae and other macroalgae as well as mosses, which do not have vascular systems. The light entering the water is selectively reflected and absorbed. According to the Lambert-Beer law, the energy flux density decreases exponentially with increasing water depth, so that in pure water at 1 m depth, only around 75 % of the light at the surface is available. Red light is the most strongly absorbed, so that the proportion of blue light increases with depth in pure water. In natural water bodies, the absorption and scattering of the various wave lengths is highly dependent on the dissolved substances in the water (humic substances and other coloured dissolved organic matter) and suspended particles. The penetration depth of light and its spectral composition can therefore differ considerably between lakes (see Fig.  4.4). Oligotrophic waters with relatively little suspended sediment thus appear blue, plankton-rich waters are green and mire waters with high humic substance contents are yellow-brown. Shade-adapted aquatic plants typically have a positive carbon balance, i.e. a photosynthesis:respiration ratio above 1, if they receive at least 1 % of the light above the surface. The depth at which this level occurs (compensation depth) varies between water bodies depending on their productivity and turbidity, and can be several centimetres (e.g. in eutrophic fishponds) to over 20 m in some oligotrophic lakes. Shading by trees on the banks can also considerably reduce the light intensity in lakes and rivers.

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Fig. 4.3  A Hjulström diagram of the transport ability of flowing water as a function of grain size and flow velocity (From Hjulström in Schwoerbel 1999)

Fig. 4.4  Spectral transmission through a layer of water 1 m thick (A to L = various lakes in southern Sweden) (From Aberg and Rodhe (1942, in Schwoerbel 1999))

Water has a very high specific heat capacity (4186 kJ kg−1 K−1 at 15 °C) combined with very low thermal conductivity. Large lakes, like seas, thus have a moderating influence on the climate of the surrounding area. Due to the low thermal conductivity, the summer warming of the water in deep lakes is limited to the surface layer, the epilimnion. Below this is a layer with a steep drop in temperature with depth (the thermocline), which effectively prevents the exchange of warmth between the warm surface water and cold lower layer (hypolimnion), causing stratification (see Fig. 4.5). Depending on the depth and the amount of heat absorbed during the summer, the following lake circulation types can be distinguished: (1) dimictic lakes with summer stratification and mixing in spring and autumn when the epilimnion and hypolimnion have approximately the same temperature (this applies to most deeper lakes),

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Fig. 4.5  Annual circulation and stratification in a dimictic, fully mixing lake (From Welch in Schwoerbel 1999)

(2) warm monomictic lakes that almost never freeze over and only mix during the winter months when the surface water is sufficiently cool (e.g. Lake Constance and Lake Plön), and (3) cold monomictic high mountain lakes that are covered by ice for most of the year and mix thoroughly during the summer (Hutchinson and Löffler 1956). This grouping applies to larger lakes, whilst shallow lakes and pools produce temporary diurnal stratification and can freeze down to the very bottom in winter. Trees growing on the banks can reduce the wind speed and thus the circulation of the water. The water movement in streams and rivers means that they do not form stable temperature layers. The annual variation in water temperature, as well as the average summer temperature, generally increases with increasing distance from the source. The water temperature of springwater is roughly the same as the average annual temperature of the surrounding area; therefore springs usually do not freeze in winter in Central Europe.

4.2.2  Chemical Characteristics In contrast to seawater, which is dominated by highly soluble sodium and chloride ions, lake and river water mainly contains Ca(HCO3)2 with small amounts of SO42−, SiO2, Mg2+, Na+ and other ions (see Table 4.2). The chemical composition of Central European freshwater bodies differs depending on the degree of biological activity, the bedrock and the precipitation patterns (see e.g. Wiegleb 1978, 1984 and Remy 1993a). Carbon dioxide is highly soluble in water, and almost 200 times more so than oxygen (see Table 4.3). Some of the CO2 reacts with the water to form H2CO3 (carbonic acid), which then dissociates to H+, HCO3− (hydrogen carbonate) and CO32− (carbonate ion). The pH of water determines the relative concentrations of dissolved CO2, HCO3− and CO32−. At pH values below 5, there is almost exclusively CO2, between pH 5 and 7.5 a mixture of CO2 and HCO3−, and between pH 7.5 and 9.5 almost exclusively HCO3−. Carbonate ions dominate in very alkaline water at pH >

4.2  Environmental Conditions and Habitat Classification Table 4.2  Comparison of the chemical composition of seawater and freshwater (in % of the total ion content)

195

CO32−, HCO3− SO42− Cl− NO3− Ca2+ Mg2+ Na+ K+ (Fe, Al)2O3 SiO2 Sr2+, B(OH4)−, Br−

Seawater 0.41 7.68 55.04 − 1.15 3.69 30.62 1.10 − − 0.31

Freshwatera 35.15 12.14 5.68 0.90 20.39 3.41 5.79 2.12 2.75 11.67 −

After Schwoerbel (1999) River water

a

Table 4.3  Relative concentrations of oxygen and carbon dioxide in pure water under normal conditions at 380 ppm CO2, and solubility of both gases at 20 °C (in mg l−1)

O2 CO2

Atmospheric concentration (vol.%) 20.99 0.038

0 °C 14.5 1.16

10 °C 11.1 0.81

20 °C 8.9 0.59

30 °C 7.2 0.44

Solubility at 20 °C 69.5 1690

10 (see Fig.  4.6). The total dissolved inorganic carbon (DIC), i.e. the content of H2CO3, HCO3− and CO32−, increases exponentially as water becomes more alkaline (see Fig. 4.6). All aquatic plants can use CO2 as a source of carbon, and some can also use the anion HCO3−, but there is as yet no evidence of plants using CO32− (see Sect. 4.4.1). The vertical distribution of CO2 and HCO3− in water bodies is mainly determined by the activity of the aquatic organisms. In the euphotic zone (i.e. where light levels are high enough for photosynthesis to occur), high rates of photosynthesis lead to a gradual reduction in DIC, while deeper zones, particularly in productive lakes, become enriched with CO2 from the respiration of heterotrophs (see Fig. 4.7). If the solubility of carbon dioxide sinks due to an increase in water temperature, or the balance of inorganic carbon shifts towards CO32− due to the production of hydroxyl ions with photosynthetic HCO3− assimilation, then this can lead to a chemical or biogenic calcification of the water. Such carbonate precipitation can be seen e.g. in the build-up of travertine in limestone springs, calcium carbonate deposits on intensively assimilating macrophytes and the formation of fine calcium carbonate crystals in the water through the activity of phytoplankton. The oxygen content of the water controls the respiration rates of the aquatic plants and of the heterotrophic aquatic organisms. The solubility of oxygen sinks with rising temperature (see Table  4.3), but at the same time the O2 demand for respiration increases. High rates of photosynthesis in summer increase the O2

196

4  Vegetation of Freshwater Habitats

Fig. 4.6  Concentration of dissolved inorganic carbon compounds (DIC) in natural water bodies as a function of pH (top: in moles, bottom: relative) (From Wetzel 2001; with permission of Academic Press)

c­ ontent of the water close to the surface, sometime leading to supersaturation. In contrast, decomposition in the hypolimnion and particularly in the sediment leads to the consumption of much of the oxygen, often leading to hypoxia. The oxygen content of a lake is thus mainly dependent on its productivity and the depth of its water. Deep oligotrophic lakes can be saturated with oxygen throughout the water column, even in summer (orthograde oxygen curve), whilst intensive use of oxygen in eutrophic lakes leads to low oxygen levels in the hypolimnion (clinograde curve, see Fig. 4.7). In running water, the oxygen consumption in summer depends on its load of organic matter, whereby the slow-moving, sediment-rich lower reaches use up more O2 than the turbulent upper reaches. Phosphorus is often a more limiting factor for growth than nitrogen in freshwater bodies. This is because phosphate, in contrast to nitrate and nitrite, forms poorly soluble compounds with Fe3+, Al3+ and Ca2+ and is thus naturally less mobile. Aquatic plants use orthophosphate (PO43−, and in acid water also HPO4− and H2PO42−), which makes up less than 5 % of the total phosphorus content of the water and, together with other labile plant-available P compounds, is referred to as ‘soluble reactive phosphorus’ (SRP). Unpolluted water bodies have SRP concentrations

4.2  Environmental Conditions and Habitat Classification

197

Fig. 4.7  Patterns of vertical variation in radiation (I), temperature (T), dissolved inorganic carbon (DIC), pH, dissolved oxygen (O2), NH4+, NO3− and soluble and total phosphorus (Ps and PT) in oligotrophic and eutrophic lakes (From data in Wetzel 2001; with permission of Academic Press)

of around 10–50 μg P l−1. Often more than 70 % of the total phosphorus content is locked up in the biomass of the aquatic organisms or bound within the detritus, and the rest is present as poorly accessible organic compounds or in colloidal form (Wetzel 2001). In eutrophic lakes with hypoxic lower water layers, the content of available phosphorus close to the lake bed is much higher, as the reducing conditions in the sediment lead to the release of P that was previously bound to iron (see Sect. 4.6.3). In contrast, nutrient-poor and/or calcareous lakes only have very low levels of available phosphorus (SRP) and total phosphorus (see Fig. 4.7). Generally, the shallower the lake, the higher the summer P concentration in its epilimnion. Under natural conditions, rivers and streams also have higher P inputs from the surrounding land and thus higher P contents than still water. Dissolved nitrogen is present in water as NH4+, NO2− or NO3− (dissolved inorganic N, DIN) or as organic nitrogen compounds (dissolved organic N, DON, e.g. as amino acids or polypeptides). Nitrogen limitation occurs mainly in ponds and lakes with high phosphorus input, and mainly affects the tall species of the reedbeds (e.g. Phragmites and Typha). Biogenic turnover leads to strong temporal fluctuations and spatial differences in the ammonium and nitrate contents of the water. DON mainly derives from animal excrement or from the decomposition of detritus, and often makes up over half of the dissolved nitrogen content of the water. Acidic water bodies, similar to acidic terrestrial soils, mainly contain NH4+, but naturally in low concentrations. Neutral to basic waters contain mostly NO3−. Nutrient-poor lakes generally have concentrations of DIN and DON of 200–400 μg N l−1 each (Vollenweider 1968, in Wetzel 2001), whereby water bodies with acid bedrock types usually have lower nitrate contents than those on limestone or clay

198

4  Vegetation of Freshwater Habitats

substrates (Bücking 1975). Neither the nitrate nor the ammonium content changes much with depth in nutrient-poor lakes (see Fig. 4.7). In nutrient-rich lakes, DIN and DON are both around 500–1500 μg N l−1. The ammonium contents increase in these lakes with increasing depth, as the lack of oxygen in deeper water leads to the release of NH4+ from the sediment. In contrast, nitrate is usually largely consumed in the surface waters in the summer due to the high rates of photosynthesis, and in the oxygen-poor deeper water by denitrifying and nitrate assimilating bacteria. Ammonium usually only occurs in high concentrations if the lake has been polluted with sewage. If enough oxygen is available, then nitrification leads to a rapid increase in the nitrate content of the water (see Sect. 4.7.3). The vast majority of Central European lakes and rivers have a pH of between 6 and 9, as they are buffered by the carbonate/bicarbonate system. Dystrophic bog lakes are naturally highly acidic (pH 3.0–4.5), as the high cation exchange in the Sphagnum moss layer leads to a strong acidification of the water (see Sect. 3.4.2). In contrast, some sodium carbonate lakes in the continental Pannonian region have pH values of over 9, as do eutrophic pools if they support high rates of photosynthesis. The calcium, magnesium and sodium contents of water are largely dependent on the geology of the catchment, whilst potassium concentrations are always low, unless the water body has been affected by fertiliser inputs (Bücking 1975).

4.2.3  Ecological Classification of Freshwater Systems 4.2.3.1  N  utrient Contents of Lakes and Ponds and Zonation of Shore Vegetation As plant species differ in their abilities to colonise open and deep water, they usually form bands along the lake shores. Their composition and width is mainly determined by the nutrient content of the water, and thus the intensity of primary production, as well as the morphology of the lake basin. Four basic nutrient levels can be distinguished here: oligo-, meso-, eu- and hypertrophic. These levels can then be subdivided based on their pH and level of humic substances into ‘acidic’ (non-calcareous), ‘neutral to alkaline’ (calcareous), or ‘dystrophic’. The six lake types in Table 4.4 differ in their nutrient content and primary productivity, as well as in their electrolyte content and sediment type. The following macrophyte vegetation is characteristic for these lake types: Oligotrophic non-calcareous (acidic) lakes and ponds typically contain submerged stands of Littorella uniflora, Lobelia dortmanna and other species of the Isoëto-Littorelletea (see Sect. 4.3.2.1) as well as sparse stands of Nitellion communities. The latter can grow in clear water on the humus-poor lake bed under 10 m or more of water. Communities of floating-leaved plants and reeds, in contrast, suffer from lack of nutrients and are poorly developed (see Fig. 4.8a).

7–8

>8

Eutrophic

Hypertrophic

0.5–2

1–7

3–20

> 1500

500–1500

300–500

< 200

150–300

Nin a (μg l−1) 150–500

> 100

35– 100

10–35

5–10

< 10

PTb (μg l−1) < 10

450–1200

300–500

< 200

< 100

< 300

Electrolytic conductivity (μS cm−1) < 100

35 to > 100

12–35

4–12

4.0)

5

4 5 5 4

o 4 5 5 2 4 3 4 1 1 2 3 2 5 4 4

5 2 1 1 1 1

e

2.4

2 1 5 1 1 1 4 2 5 5 4 4 2 5

m

4

1 3 3 1 3 3 1 5 4 2 2

e

3.3

Potamogetonetea 1 2 3 4 0.7 1.5

1 5 2 5 4

e

3.6

5

2 5 2

2

h

6

4 4

2 5

4.6 m

5

5 5

3

e

1.8

4

e

(3.0)

Lemnetea 7 8 9 1.5

2

2

h

10

3 3

– – 4.6 e

11 –

3 2

– – – e

12 –

– T 5 6 6 4 6 6 6 6 6 4 6 6 6 5 5 6

R 6 7 6 6 5 8 5 4 7 6 7 7 5 7 7 7

N 3 6 6 6 4 6 3 – 8 4 6 5 6 5 5 7

(continued)

poor in P poor in P? poor in P poor in P poor in P poor in P

poor in P

Ecological characteristics of the species

Table 4.7  Rooted and free-floating aquatic plant communities in eastern Lower Saxony and two water fern communities of southern Central Europe. Based on tables in Weber-Oldecop (1969) and Müller and Görs (1960)a

4.3 Vegetation 211

Vegetation class: Serial no.: Callitriche platycarpa fo. natans L Spirodela polyrhiza L Lemna minor P Potamogeton friesii Po Ceratophyllum demersum L Lemna gibba Po Myriophyllum spicatum Po Potamogeton pectinatus Po P. panormitanus H Utricularia vulgaris N Nymphaea alba N Polygonum amphibium fo. natans Po Potamogeton lucens Glyceria fluitans Gl Sium latifolium Ph Oenanthe aquatica Ph Butomus umbellatus Ph Alisma plantago-­aquatica Ph L M Ricciocarpos natans Montia fontana Nasturtium officinale Gl P Zannichellia palustris L Salvinia natans L Azolla filiculoides

Table 4.7 (continued) Potamogetonetea 1 2 3 4 5 3 3 2 5 5 5 5 5 5 5 5 2 2 1 5 4 3 3 3 2 1 1 1 1 1 1 5 2 1 3 2

5 2 1 2 1

5 2 4 2

6 3 5 5 5 5 5 3 2 3

5 4 5 5 5

2

3

2 2 2

5

Lemnetea 7 8 9 2 5 5 5 5 5 5 5

2 1

5

10 2 5 5 4

4

2 5

1 1

2 1

3 2

1

12

11 6 6 5 6 7 6 6 × 5 6 6 6 6 × 6 6 6 5 6 4 × 6 8 8

7 × × 7 8 8 9 8 7 5 7 6 6 × 7 7 × × 6 5 7 8 7 ×

7 6 6 6 8 8 7 8 ┴ 4 5 4 7 7 7 6 7 8 – 4 7 8 7 8

rich in P? rich in P

Poor in P

rich in P

rich in P?

rich in P

rich in P? rich in P rich in P

212 4  Vegetation of Freshwater Habitats

3: 4: 5: 6: Nr. 7 bis 10: 7: 8: 9: 10: Nr. 11 u 12: 11: 12:

1: 2: Nr. 3 to 6:

Communities (No. 1–10 in eastern Lower Saxony) Nr. 1 and 2:

mT

mR

mN

Water lily communities: Myriophyllo-Nupharetum (alliance Nymphaeion), mainly rooted in the sediment: Potamogeton obtusifolius-Nuphar community, nutrient-poor, 5.6 6.4 5.3 typical Myriophyllo-Nupharetum, relatively nutrient-rich 5.9 6.9 5.7 Floating plant communities: Hydrocharition morsus-ranae, partly rooted in the sediment, and the eutraphent Ranunculion aquatilis: Hottonietum of fairly nutrient-rich shallow lakes, sometimes shaded 5.8 6.4 6.1 Hydrocharitetum, 6.1 6.6 6.1 Ranunculetum peltati, 5.4 7.3 6.5 Ceratophylletum demersi of highly polluted water bodies 6.0 7.5 6.7 Duckweed and other communities of free-floating species: Lemnion Riccietum fluitantis, 5.8 6.0 5.6 Lemno-Spirodeletum, 5.9 7.2 6.3 ‚Lemnetum minoris’, 6.0 7.0 6.4 Lemnetum gibbae 5.8 7.4 6.6 Floating fern communities (in Baden-Württemberg, usually thermophilic), assignable to Lemnion: Spirodelo-Salvinietum, 6.4 7.0 6.0 Lemno-Azolletum 6.8 6.7 6.5 (continued)

Ecological characterisation:

L Species of the Lemnetea (duckweed communities), H Species of the Hydrocharition morsus-­ranae (Potamogetonetalia), P Species of the pondweed communities (Potamogetonion, Po = Potamogetonetalia), N Species of the Nymphaeion (Potamogetonetalia) (Water lily communities), R Species of the Ranunculion aquatilis (Potamogetonetalia), Ph Phragmitetea etc. Gl Glycerio-Sparganion b Neutralising capacity of the water (alkalinity) determined by titration of 100 cm3 water with n/10 HCl. Average amplitude as graphically depicted by WeberOldecop (1969) for each community Indicator values: T Temperature, R Reaction, N Nitrogen after Ellenberg et al. (1992)

a

4.3 Vegetation 213

Table 4.7 (continued) Still water bodies are usually warmer (averaged across the year) than the soils in the surrounding land. As a result, plants can live here that require higher temperatures than the zonal terrestrial vegetation. Even in eastern Lower Saxony, the species of this habitat have average temperature indicator values >5. The Hydrocharitetum (no. 4) is relatively thermophilic. The warmer parts of southwestern Germany support floating fern communities, some of which have a submediterranean character (no. 11 and 12) Increasing alkalinity of the water is reflected by increasing reaction values for the plant communities. They have a somewhat closer correlation with the average nitrogen values that, with caveats, can be used to express the trophic status of the water. The concentration of plant-available phosphorus, which is necessary to calculate the latter, was not measured by Weber-Oldecop. It is, however, relatively certain that communities 1, 3 and 7 colonise relatively clean water, whilst 6 and 10 (and probably 12) indicate highly polluted conditions. Species that only survive in water with relatively low phosphorus levels (based on WeberOldecop 1969, Wiegleb pers. comm. and own observations) are marked on the right hand side of the table with ‘poor in P’, whilst the indicators of eutrophic conditions are marked with ‘rich in P’

214 4  Vegetation of Freshwater Habitats

4.3 Vegetation

215

constancy, as the vegetation surveys were selected based on their presence. These are the only species that reach coverage values of more than 50 %, so are often the dominant species. They include three character species of the alliance, Nymphaea alba, Potamogeton natans and Polygonum amphibium. The most common and also the most species-rich community of the Nymphaeion albae in Central Europe is the Myriophyllo-Nupharetum (Table 4.7: no. 2, see also Fig.  4.1). It contains both Nuphar lutea (yellow pond-lily) and Nymphaea alba (white waterlily), whereby the latter is more common in shallow water. Nymphaea alba cannot grow below depths of 3 or 4 m, as unlike Nuphar, it cannot form submerged leaves. For the same reason, the Nuphar lutea dominates in lakes with strong fluctuations in water level as well as in slow-flowing water, as long as the nutrient content in the water is high enough (Roweck 1988). Alongside one or even both of these Nymphaeaceae, communities of floating-leaved plants almost always contain Potamogeton natans (floating pondweed), Myriophyllum verticillatum (whorled water milfoil) and Ceratophyllum demersum. In warmer and usually quite shallow lakes and oxbows, e.g. in the Upper Rhine valley, Trapa natans (water caltrop) occurs. During the Holocene climatic optimum its range stretched much further north than today (Walter and Straka 1970). Its hard seeds with their hooked sides used to be used as a source of food and were crushed, similar to hazelnuts, with special wooden batons (Apinis 1940). Whilst the floating-leaved plant communities of eutrophic and relatively warm still waters are characterised by the presence of Ceratophyllum demersum, nutrient-­ poor waters are dominated by Potamogeton natans or P. obtusifolius with Myriophyllum alternifolium (see no. 1 in Table 4.7). Cooler and more acidic bog lakes e.g. in the Prealps contain Nupharetum pumilae communities, thought to be a glacial relict (Sebald et  al. 1990–1992, Roweck and Reinöhl 1986). In montane, cooler climates, the species richness of the Nymphaeion communities decreases, especially as the water is usually relatively nutrient-poor. Oligo– to mesotrophic shallow ponds in the southeast of Central Europe contain the Nymphaeetum candidae with Nymphaea candida; this community is locally more common than assemblages dominated by Nymphaea alba (Nympaeetum albae). Bladderwort Communities  In oligotrophic to mesotrophic waters are found the low-productive assemblages of the class Utricularietea which are composed mainly of free-swimming species only rarely anchored in the sediment. The Utricularia species have adapted to the low nutrient conditions by being facultative carnivores. According to Pietsch (1977) and Hofmann (2001), bladderwort species can grow over a wide range of pH and conductivity, from extremely acidic waters poor in electrolytes (U. minor can be found as low as pH 3.5) to moderately base-­ rich, mesotrophic pools, in which U. vulgaris and australis occur. Today, the majority of these communities are seriously threatened by eutrophication.

216

4  Vegetation of Freshwater Habitats

Fig. 4.11  The duckweed Lemna gibba forms a dense carpet on the surface of an eutrophic pond (Aeroe, southern Denmark)

4.3.2.3  Duckweed and Other Free-Floating Communities The sheltered surfaces of oxbow lakes, pools and inlets provide a habitat for floating plant communities that would not be able to form on moving water. Pott (1995) and Rennwald (2000) divide the class and order of duckweed communities (Lemnetea, Lemnetalia) into three alliances (see Chap. 14: no. 1.1.1): 1. the Riccio-Lemnion trisculcae of submerged pleustophyte (i.e. non-rooted) communities in weakly eutrophic waters; 2. the Lemnion gibbae (true duckweed communities) on the surface of highly eutrophic to hypertrophic water, which also contains the thermophilic floating fern communities; 3. the Hydrocharition (multi-layered, complex Statiotes aloides and Utricularia communities). Each of these alliances has multiple associations that are each characterised by their dominant pleustophytic species, and are all otherwise species-poor (cf. also Tüxen 1979b). They are often found mixed with floating-leaved plant and reed communities. The Lemnetalia communities are fairly insensitive to water chemistry, and the distribution of the individual species of Lemnaceae is mainly dependent on the climate (Landolt 1984). Strasburger and Homann (1982) state that all Lemnetalia communities can occur at pH values of 6.0–6.5. However, the higher the nutrient content of the water, the better these floating communities grow (Klose 1963; see Fig. 4.11). All species of duckweed grow well at moderate to high P and N concentrations, and are therefore reliable indicators of eutrophic to hypertrophic conditions

4.3 Vegetation

217

Fig. 4.12  Layers of floating plant communities, particularly the Hydrochari-Stratiotetum (Modified from Weber-­ Oldekop (1969) (the perspective of the surface of the water is exaggerated))

(Pott and Wittig 1985; Pott 1995). Lüönd (1983) arranged the duckweed communities according to decreasing nutrient levels thus: Lemnetum gibbae > Spirodeletum polyrhizae > Lemnetum trisulcae > Ricciocarpetum natantis > Riccietum fluitantis. Lemna minor has a broad ecological amplitude and occurs in most of these communities. All free-floating plant species are to some extent thermophilic and grow preferentially on water bodies that remain free of ice for most of the year. The Lemnetea communities are thus richest in atlantic and submediterranean Europe, and become poorer towards Central Europe. Only Lemna minor reaches into the mountains up to 1000 m a.s.l. In the mild Upper Rhine Plain and other warm basins, the duckweed species are joined by tropical fern species, such as Salvinia natans and the rarer Azolla filiculoides (water fern, see e.g. Philippi 1969); recently, these have also migrated into northern Germany (Kesel and Gödeke 1996). These ferns, as well as duckweed and Riccia spp. (floating liverworts) possess many derived characteristics, so cannot be considered phylogenetically primitive. In nutrient-rich, shallow lowland rivers, ponds and lakes with high organic matter contents in the sediment, layers of large-leaved floating Hydrocharis morsusranae (frogbit) and Stratiotes aloides (water pineapple) form, which crowd out the duckweed (Klosowski 1994; see Fig. 4.12). The Stratiotetum aloidis often mixes so thoroughly with the Hydrocharietum morsus-ranae, the Myriophyllo-Nupharetum or other communities of floating-leaved plants rooted in the substrate, that they are difficult to separate. If the water body has a lot of boat traffic, then the leaves of rooted plants can be damaged or ripped off. The almost free-floating Stratiotes aloides, in contrast, simply move apart and then back together after the boat has passed. Eventually, this will become the dominant species, as used to be the case e.g. in the canals of the Havelland near Berlin. Hydrocharis morsus-ranae and Stratiotes aloides can also occur in submerged, rooted forms. Their connection to their roots breaks, however, if the water level rises too high, e.g. through the flooding of mead-

218

4  Vegetation of Freshwater Habitats

ows where they are present in the ditches. Handke (1993) observed that populations of Stratiotes were destroyed by the flooding of the Weser marshes. 4.3.2.4  Still Freshwater Reedbeds Reeds that grow high above the water dominate over the hydrophytes that remain below or close to the surface as soon as they can form denser stands and block out the light. This is the case e.g. for Phragmites australis (common reed) at average water depths up to 1.2–2 m. Its white, tubular runners stretch through the mud into even deeper water, and in sheltered lakes it can grow at water depths of up to 2.7 m (Rodewald-Rudescu 1974). However, these stems are usually quite weak and only reach their full strength if the water level drops. The limits of reedbeds are almost always physiological, and not caused by competition. Phragmites is the most competitive of the Central European aquatic macrophytes, playing a role somewhat similar to that of beech on land. It also has wide ecological amplitudes, ranging from basic to acidic and oligotrophic to eutrophic water bodies, as well as from deep water to occasionally exposed soils (Krausch 1965b). Pietsch (1965) even found Phragmites, Typha latifolia (bulrush) and other reed species in a pool with pH values between 2.9 and 3.0 in an open-cast mine where brown coal had been extracted. The communities formed by Phragmites are similarly diverse, and in addition to the above mentioned studies have been described by a number of authors (e.g. Balatova-­ Tulackova 1963; Hilbig 1971b; Weisser 1970). Nevertheless, common reed communities are now usually considered as one association (Phragmitetum australis). The habitat conditions and the appearance of reed communities (order Phragmitetalia australis) are largely similar across much of Europe, from Spain (Bellot Rodriguez 1964) to Finland (Eurola 1968) and the Danube delta (Krausch 1965a; Horvat et al. 1974), where the largest contiguous stands are found. Even the stands in the uplands of southwestern Japan had parallels to the European communities (Horikawa et al. 1959). In still lakes, Schoenoplectus lacustris (common club-rush) forms pioneer reed vegetation before Phragmites begins to dominate (see Fig. 4.8: eutrophic), as its green stems continue to assimilate when covered by water during flooding (see Sect. 4.4.4). Their stalks can reach depths of 1.5  m below the surface, and their rhizomes can even reach below 5 m in places. Wherever Phragmites grows well, it forms a dense stand up to 3.5 m above the water and can reduce the light intensity at the water level to less than 1 % of incident light (Meyer 1957). As a result, only low densities of the order and alliance character species listed in Table 4.8 can survive alongside it. The only exceptions to this are the two bulrush (Typha) species that sometimes dominate in nutrient-rich, muddy and still inlets or shallow lakes that dry out in years with low rainfall. Typha species are more successful than Phragmites at regenerating by seed (see Sect. 4.5.2) and can therefore quickly form dense stands on bare mud (Coops et al. 1996). Phragmites is by far the most important reed species in Central Europe, at least in still water bodies. Strong eutrophication also promotes Glyceria maxima (reed

4.3 Vegetation

219

Table 4.8  An overview of the reedbeds and tall sedge communities and their character species. Based on Tüxen, Preising, Oberdorfer and particularly Pott (1992). A = also a character species of an association that is named after this usually dominant species Class and order Phragmitetea and Phragmitetalia (+)   Alisma plantago-aquatica   Carex pseudocyperus Iris pseudacorus   Comarum palustre Lycopus europaeus   Eleocharis palustris Phragmites australis   Equisetum fluviatile Rumex hydrolapathum   Galium palustre Scutellaria galericulata A   Glyceria maxima Sium latifolium   Still water reedbeds Reedbeds on streams and rivers   Phragmition (+) Nasturtio-­Glycerietalia A   Acorus calamus A   Butomus umbellatus   Eupatorium cannabinum A   Hippuris vulgaris   Lysimachia vulgaris   Lythrum salicaria   Mentha aquatica   Myosotis palustris A   Oenanthe aquatica   Ranunculus lingua   Sagittaria sagittifolia   Schoenoplectus lacustris   Sparganium emersum A   Sp. erectum subsp. microcarpum A   Sp. e. ssp. neglectum   Typha angustifolia   T. latifolia   Reedbeds in brackish water   bolboschoenion   Bolboschoenus maritimus A   B. m. var. maritimus   Eleocharis uniglumis (?)   Schoenoplectus americanus A   Sch. tabernaemontani   Sch. triqueter

     Reedbeds on streams      Glycerio-Sparganion A  Catabrosa aquatica A  Glyceria fluitans A  G. plicata A  Leersia oryzoides A  Nasturtium microphyllum A  N. officinale      Sium erectum       Veronica beccabunga      V. catenata      Reedbeds on rivers      Phalaridion A  Phalaris arundinacea A  Rorippa sylvestris

Tall-sedge communities Magnocaricion elatae (+) A  Calamagrostis canescens A  Calla palustris A  Carex acutiformis A  C. appropinquata A  C. aquatilis A  C. buekii A  C. cespitosa A  C. disticha A  C. elata A  A  A  A  A  A 

C. gracilis C. oenensis C. paniulata C. riparia C. rostrata C. vesicaria

A  C. vulpina A  Cicuta virosa      Lysimachia thyrsiflora      L. vulgaris

A  Menyanthes trifoliata A  Peucedanum palustre      Scutellaria galericulata Intermediate between reedbeds and tall-sedge communities: A  Cladium mariscus

220

4  Vegetation of Freshwater Habitats

Fig. 4.13  A common zonation of plant communities in a nutrient-rich lake in southern Germany. From Ellenberg (1952a). The water lily, reedbed and Carex elata communities are still semi-­ natural, whilst the small-sedge and Molinia communities are maintained by regular mowing in the place of alder swamp and alder-ash forest or moist oak-hornbeam forest. The numbers refer to the mass of hay cut in autumn in Mg dry mass per hectare. The decrease in yield in the unfertilised litter meadows, particularly that of the small-sedge swamp, is mainly due to low N mineralisation and denitrification caused by the lack of oxygen in the permanently wet soil. Today, these meadows have mostly been abandoned and the forest is beginning to develop again (see Fig. 4.3 in Vol. I)

mannagrass), although this does not reach the same size as Phragmites (Lang 1967a). The Glycerietum maximae usually also form monospecific stands, and further resemble the Phragmitetum in that they reach their greatest biomass in late summer (Westlake 1966; cf. Fig. 4.27). Although Sparganium erectum (bur reed) has one of the highest levels of productivity in the temperate zone (Dykyjova and Ondok 1973), it does not form dense stands and rarely grows above 1 m in height. It thus plays a less important role in reedbeds and forms the transition to the tall-­ sedge communities, which we will discuss in the next section. 4.3.2.5  Tall-Sedge Communities (Magnocaricion) If the water becomes shallower so that the lake increasingly dries out, then the beds of common reed will become smaller and sparser, leaving space for tall-sedge communities (see Figs 4.8d and 4.13). These sedge communities share so many species with the reedbeds that they are still placed in the same order and class (Phragmitetalia, Phragmito-magnocaricetea), but are assigned their own alliance (Magnocaricion; see Table 4.8). Alongside the Carex species, they are characterised by a sub-species of Galium palustre adapted to flooding and Poa palustris. The species composition of the Magnocaricion depends on the level and duration of flooding and the nutrient content of the water. Generally, they are dominated by one sedge species that is also a character species. The comparative study by

4.3 Vegetation

221

Fig. 4.14  In early winter, the column-like tussocks of Carex paniculata stand like thick palm trees on the mud of a dried out pond near Heiligenstadt (Thuringia, central Germany)

Busch (2000) showed that the different amounts of aerenchyma in the tall-sedge species probably determine their preference for a certain water depth. Their photosynthetic capacity and water use, in contrast, show no clear relationship with the habitat. The greatest changes in water level are tolerated by the Caricetum elatae, which still cover large areas of lake edges in the foothills of the Alps and in Southeastern Europe, and used to be mown for stall litter. In Hungary, it also grows well in areas where the water levels fluctuate little and remain for long periods at or somewhat above the bases of the leaves of the sedge (Kovacs 1968b). In autumn when the lakes have dried out, the columns of Carex elata (tussock sedge) stand up to 1.2 tall above the lake bed (see Fig. 4.13) consisting mostly of their roots and leaf sheaths. These clumps become denser moving further away from the water, until only a few small gaps remain. Without the repeated cutting for litter, these areas would have long since been colonised by Alnus glutinosa and Salix cinerea and would have quickly turned into an alder swamp (see Ellenberg and Klötzli 1967). In the forest swamps of the middle Rhine region described by Ullmann et al. (1983), the transition to dry land also begins with the Caricetum elatae and leads to the Carici elongatae-­Alnetum. In northeastern Central Europe, the Caricetum elatae is rare and is usually replaced by the Caricetum gracilis. The species that lends its name to the association (now Carex acuta) does not form hummocks but smooth swards that, tousled by the wind, resemble animal fur when viewed from afar (see Fig. 4.15). Only when trampled by livestock does Carex acuta form tussocks. Usually these communities are flooded in spring and are less wet in summer than the Caricetum elatae, but this pattern can also be reversed (cf. Balatova-Tulackova 1957 and Kovacs 1968b).

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Fig. 4.15  Large expanses of the Caricetum gracilis, which consists almost solely of Carex acuta and a few stalks of Phragmites, in the Billetal east of Hamburg

The periodically flooded shoreline of acidic oligotrophic to moderately dystrophic lakes supports only sparse and low-growing sedge communities, which are usually dominated by Carex rostrata (bottle sedge). The Caricetum rostratae is even found in small pools and marshy edges of raised bogs, where the low nutrient level means that reedbeds cannot form (see Fig. 3.5). Apart from the three communities of tall-sedges mentioned above, numerous other communities have been described from Central Europe, which are also dominated by Carex species of at least 30–50 cm or even over 1 m in height (see for example Chytrý 2011). This is usually a single species that excludes most other herbaceous species from the stand. Tall-­ sedge communities are thus almost always species-poor. Since they are no longer mown in autumn for livestock bedding (cf. Sect. 8.3.3.2.1), they are often invaded by the rhizomes of neighbouring reedbeds. They have also been colonised by black alder or willow scrub of the alliance Salicion cinereae (see Sect. 9.3.2.1  in Volume I), showing that tall-sedge communities resemble swamp forests in their water regime. It is only the wetter Caricetum elatae and Caricetum rostratae communities that are completely free of woody vegetation. Various authors have classified the Magnocaricion alliance into around 10 to 25 associations (Rennwald 2000, for example, lists 13), leading to several attempts to arrange these into suballiances. However, none of these attempts have been particularly successful, as the communities react differently to different habitat factors. Instead of providing a further systematic classification here, we will group the most important species of tall-sedges according to their ecology (see Table 4.9). Each of the species listed in the table is a character species of an association – often the only

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Table 4.9  Ecological characterisation of the major species of tall sedge in Central Europe based on their indicator values (L, T, C, M, R, N) Ecological group Species 1 Carex rostrata C. lasiocarpa 2a C. elata 2b C. paniculata C. riparia C. pseudocyparus C. randalpina 2c C. acutiformis 3 C. acuta C. cespitosa C. vesicaria C. diandra 4a C. appropinquata C. aquatilis C. disticha 4b C. vulpina C. buekii

L 9 9 8 7 7 7 6 7 7 6 7 8 8 9 8 9 8

T × 4 × × 6 6 6 × 5 6 4 6 5 6 6 6 6

C × × 2 3 3 3 4 3 7 7 × 7 5 5 × 5 6

M 10 9= 10~ 9 9= 9= 9 9~ 9~ 9= 9= 9= 9= 9= 9= 8? 8?

R 3 4 × 6 7 6 7 7 6 7 6 6 9 7 8 × 8

N 3 3 5 5 4 5 6 5 4 4 5 3 4 4 5 5 5

From Ellenberg et al. (1992) ‘×’ indifferent, ‘=’ frequently flooded, ‘~’ periodically wet

one and also lending its name to the association. The species can be divided into four groups according to their habitat types: The central group of the Magnocaricion is the Carex elata group (# 2  in Table 4.9). Carex elata (2a), as we have seen, occupies a somewhat unique position as the species best adapted to changes in water level. However, it is otherwise quite similar to the species of 2b in terms of their oceanic range and their moderate demand for nitrogen and basic cations. This is also the case for Carex acutiformis (2c), which is unusual in that it can form distinct stands in various different Magnocaricion communities (even in the Caricetum elatae) and even penetrates into swamp forests. Recently, it has even been recorded as a dominant species in abandoned wet meadows, as it is clearly better at dispersing and adapting to new habitats than other tall-sedge species. All species in subgroup 2b (see Carex paniculata as an example; Fig. 4.14) can also be found in the semi-shade of alder, ash and other floodplain tree species, and are thus assigned low light indicator values. The Carex acuta group (3) behaves in some respects like subgroup 2b, but is mostly found in more continental climates and on more nutrient-rich soils. The Carex appropinquata group (4) indicates more calcareous soils, and the water regime of 4a is largely the same as group 3. Carex vulpina and C. buekii (4b) appear to tolerate temporary drought better than the other tall-sedge species. C. buekii is often found along streams and rivers in eastern Central Europe. The Carex rostrata group (1) is the opposite of group 4, as it is the best adapted to acid and nutrient-poor conditions. Like Carex elata, C. rostrata is more of an aquatic plant

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than a species of wet soil. Carex lasiocarpa, in contrast, behaves more like a small-­ sedge species, which is discussed in Sects. 3.3.4 and 3.3.5. Many authors also consider this species as characteristic for its alliance (Caricion lasiocarpae, see Chap. 14: no. 1.7.1.2) in the order Scheuchzerietalia. Carex diandra (in group 3) also occurs mainly in small-sedge communities, but those on calcareous soils (Caricion davallianae, see Sect. 3.3.5). According to Marti (1994a), the Magnocaricion species differ considerably in the nutrient content of their biomass, but relatively little in that of their soil. In his study, Carex elata was rich in NH4+ dissolved in the cell sap, whilst the other species generally contained more NO3−. Carex riparia was characterised by relatively large concentrations of P. It is still not entirely clear how these differences evolved and what their ecological significance is. A comparison of the photosynthetic capacity, water use and anatomic characteristics of many of the Central Europe sedge species of damp and dry habitats was conducted by Busch (2000). The Cladietum marisci is unusual in that it grows on calcareous but nutrient-poor lake sediments, but does not tolerate large fluctuations in water levels. Görs (1975) concluded that the typical subassociation can be placed within the Phragmition, whilst the remaining subunits are more related to the Magnocaricion. The Cladietum reached its largest distribution in the early post-glacial period, as the lakes of Central Europe were mostly calcareous and oligotrophic (Hafstein 1965). The Sphagnum palustre type of this community can be seen as a ‘living fossil’ (Görs 1975). On the one hand, Cladium mariscus was found to be sensitive to both high water levels and low oxygen levels (Convay 1937). It is thus much more sensitive than Phragmites, and such a weak competitor that it is slowly disappearing across its global range. On the other hand, Balatova-Tulackova (1991) described the Cladietum marisci as being highly adaptable to changing environmental conditions. Cladium mariscus is particularly resistant to fire and overgrazing, especially as it can regenerate easily from its long underground runners. It has even largely outcompeted Phragmites on calcareous sediments in the Rhône valley since the practice of reed cutting was abandoned, and forms almost impenetrable thickets in places (Pantou and Griel 1983).

4.3.3  Streams and Rivers 4.3.3.1  Reedbeds in the North Sea Tidal Estuaries The tidal areas of the Central European rivers that flow into the North Sea experience rather unique conditions. The reedbeds in these estuaries colonise tidal mudflats on the river banks which stretch from the oligohaline to the mesohaline zone and finally give way to the euhaline communities of the lower salt marshes on the North Sea coast. The vegetation of the largest freshwater tidal area in Europe, the Biesbosch in the Brabant region of the Netherlands, was described by Zonneveld

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Fig. 4.16  Zonation of the sand bank vegetation in the freshwater tidal zone along the Elbe downstream of Hamburg. Semi-schematic and not to scale. Modified from Kötter (1961). At high tide, so much organic material is deposited that nitrophilic algae (Vaucheria), annual ruderal plants (e.g. Bidens species) and moisture-dependent tall reed beds thrive here. In contrast, apart from a few drift line deposits, the area above average high tide levels is nutrient poor, and dominated by communities of Festuca, Thymus or Carex arenaria on dry soil

(1960). In Germany, Kötter (1961) and Meyer (1957) studied the plant communities in the lower reaches of the Elbe. The power of the rising North Sea tide is so great that it reverses the flow of the river water twice a day in the river estuaries. However, salty seawater only reaches the lower part of the river mouth, as the ebb tide sucks it back out after 6 h with the additional pressure from the river water. In the Elbe, for example, the water remains fresh upstream of Glückstadt (around 30 km from the sea), but still undergoes daily fluctuations in water level that, even further inland in Hamburg, are around 3 m in height (see Fig.  4.16). The freshwater mudflats with large tidal amplitude in the lower reaches of Elbe and Weser are a unique habitat for plants. Northwesterly storms and spring tides, or offshore winds and neap tides can increase the amplitude of these fluctuations to up to 8 m. Phragmites can survive under these conditions, and even thrives in estuaries better than almost anywhere else, as the sediments here are rich in nutrients and provide some of the highest rates of mineral nitrogen supply in Central Europe (Meyer 1957). Reeds can only colonise areas that are less than 0.5 m below the average high tide level, i.e. are not covered by water for several hours a day. These sediments even support earthworms and other aerobic animals, as the water does not cover it for long enough to expel all the air from the porous mud. Before and whilst the reeds start growing in spring, other lower-growing plants use the fertile sediments to their advantage. Caltha palustris (kingcup) can be found here growing at heights of up to 1 m (see Fig. 4.17), far greater than in any other habitat, whilst Ranunculus ficaria (lesser celandine) covers the soil surface like in a

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Fig. 4.17 Large Caltha palustris (kingcup) plants in the freshwater tidal reed beds on the Schweinesand island in the Elbe near Hamburg-Blankenese. Phalaris arundinacea is visible at the front of the picture, and Typha angustifolia at the back

floodplain forest. Just like the forest trees, the reeds finally push these early-­ flowering plants back into a dormant state through shading. These estuarine tidal reedbeds differ in many respects from the reedbeds of lakes and ponds. Particularly noticeable is that two endemic species have developed in the reedbeds and inland neighbouring communities of the Elbe estuary: Deschampsia wibeliana and Oenanthe conioides (von Weihe and Reese 1968; Below et al. 1996). The only community to occupy deeper water in the estuaries than the common reedbeds is that of the Schoenoplecti triquetri-Bolboschoenetum maritimi (Fig. 4.18). It is often falsely referred to as community of brackish water, but can in fact also be found in freshwater (Kötter 1961). However, it is reliant on the tidal movements of the water level, without which it would be replaced by Phragmites or floating-leaved plants. Like Phragmites, Bolboschoenus maritimus is also found across the globe, and not only in reedbeds. It is also found in freshwater lakes, rivers, oxbow lakes and ditches, and even on damp arable fields across Central Europe (Hilbig 1994) where five associations have been distinguished (Hroudová et  al. 2009). A special form of Bolboschoenus maritimus is salt-tolerant and can grow in brackish water at the coast and in saline inland waters, whether it has fluctuations in water level or not. It is more resistant to salt content than Phragmites (Borhidi 1970), although Phragmites can also be found in saline water up to 15 ‰ (Rodewald-­

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Fig. 4.18  A Bolboschoenetum maritimi community in the freshwater tidal zone near Haseldorf (Elbe estuary). Foreground centre: Schoenoplectus tabernaemontani (blue green colour) with Nasturtium officinale and other tall herbs; background: Scirpus maritimus (light green) on the higher banks. A shelduck (Tadorna tadorna) family feeds on the tidal flats

Rudescu 1974). This is a salt-tolerant ecotype that, even after several years of cultivation by Dykyjova (1971) on the same soil as the normal erect type, still had a characteristic low-growing habitus. Schoenoplectus lacustris and Typha latifolia have much lower salt tolerance (up to 2.5 ‰, Luther 1951). Other moderately tall sedges can form patches alongside Bolboschoenus maritimus, including Schoenoplectus tabernaemontani (softstem bulrush). However, even the Bolboschoenetum maritimi cannot survive further than 1.2 m below the mean high tide level. Water lilies and other floating-leaved plants, as well as submerged pondweeds and the Characeae, are unable to cope with the mechanical stress of daily changes in flow direction and water level. Further inland, the reed belt in the freshwater tidal zone is usually followed by Phalaris arundinacea or drift line communities instead of a tall-sedge community, and then finally a willow floodplain forest. These communities are discussed along with the floodplain vegetation in Sect. 9.3.1.3.1 in Volume I. The brackish water reedbeds along the Baltic Sea coast (which doesn’t have tides) and the adjacent Bodden lagoons are discussed in Sect. 1.3.2.2 in this volume.

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4.3.3.2  Aquatic and Riparian Vegetation of Rivers and Streams Flowing water is an extreme and relatively hostile environment (Thienemann 1925), in which only a few plant and animal species thrive. The ‘lentic’ habitats of rivers and streams, i.e. in the quieter inlets and slip-off slopes (inside banks of meanders), contain many more species than the ‘lotic’ habitats with stronger currents. However, the few species that can survive in the fast flowing water largely escape competition (Haslam 1978), and often dominate in species-poor communities. Species-rich river communities, in contrast, can be seen as indicators of human disturbance (Philippi 1973b). Quieter parts support vegetation similar to that in eutrophic lakes, whilst areas with stronger currents have almost no species, or mainly aquatic bryophytes (Roll 1938; see Figs. 4.19, 4.20, 4.21 and 4.23). The plant life form types of lakes or ponds are thus also found alongside or within flowing water bodies (Soltau 1993; cf. Sect. 4.4.5). However, they are usually present in rivers and streams as different species, forms, varieties or subspecies adapted to flowing water. Many species with round or quite broad leaves in still water develop long, thin and grass-like leaves in moving water, such as Sagittaria sagittifolia fo. vallisneriifolia (arrowhead). For example, the following algae, bryophytes and higher plants (or their river form) are characteristic for fast flowing rivers in Schleswig-Holstein: Higher plants: Butomus umbellatus fo. submersus Glyceria fluitans fo. submersa Potamogeton nodosus

Fig. 4.19 A Ranunculus aquatilis community in the outlet stream of a pond near Clausthal-­ Zellerfeld in the Harz Mountains on acid bedrock

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Fig. 4.20  Submerged stands of Callitriche palustris agg. dominate the slightly polluted rhitral water of a small rivulet in the Pleistocene lowlands of northern Germany (Lüneburg Heath), while Berula erecta is an abundant helophyte in the riparian zone

Fig. 4.21  The distribution of plant species in a mountain river with a stony bed (River Ourthe in the Ardennes, Belgium) caused by flow patterns (From Vanden Berghen 1953). R Ranunculus aquatilis, M Myriophyllum (low flow rates!), L Lemanea (alga), F Fontinalis, B Brachythecium rivulare, C Cinclidotus fontinaloides (F, B and C are bryophytes)

Ranunculus trichophyllos Sium erectum fo. submersum Algae (A) and bryopyhtes (B): A Cladophora glomerata A Hildenbrandia rivularis A Hydrurus sp. B Hygroamblystegium irriguum B Rhynchostegium rusciforme

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Whilst slower moving water (between 13 and 17 cm s−1) supports: Elodea canadensis Glyceria maxima fo. submersa Lysimachia nummularia fo. submersa Phalaris arundinacea fo. submersa Sparganium erectum fo. submersum Potamogeton crispus, P. lucens, P. nitens, P. perfoliatus. Overviews of the plant communities of moving and still water bodies can be found in e.g. Kopecky (1966) for southeastern Central Europe and Weber-Oldecop (1969) for eastern Lower Saxony. Weber-Oldecop (1977) provides a phytosociological typology of river vegetation for the whole of Lower Saxony. This has since been modified by Wiegleb (1979, 1981a, b) and extended by Pott (1990). Preising et al. (1990a) also discuss the plant communities of the rivers and streams of Lower Saxony, and Runge (1981b) those of the Ems. Regional overviews of the river vegetation can be found e.g. in Moor (1958) for Switzerland, Roll (1938) for eastern Holstein, Niemann (1965) for Thuringia and Chytrý (2011) for the Czech Republic. Remy (1993a) characterised the rivers of northwestern Germany according to the chemical properties of their water (see Fig. 4.22). Pott and Remy (2000) list characteristic plant communities for the main river types of Central Europe according to elevation (high mountain, low mountain, lowland) and bedrock. The majority of authors distinguish between vegetation types of the upper (‘rhitral’) and lower (‘potamal’) reaches. The turbulent upper reaches, which are often shaded by trees, are frequently dominated by aquatic mosses. These are usually more tolerant of currents and shading than the phanerogams, as well as being dependent on dissolved CO2 as a source of carbon, which is in good supply here due to the movement of the water. Areas of water with slightly less turbulence are dominated, and characterised, by higher plants of the genera Callitriche and Ranunculus (see Fig. 4.20). It is only in the lower reaches with their quieter inlets and nutrient– and hydrogen carbonate-rich water that the characteristic communities of the Potamogetonetalia and Lemnetalia develop (Wiegleb 1981a). The riparian zone is characterised by reedbeds that differ in both their ecology and their species composition to still water reedbeds (see Sect. 4.3.2.4). This is because flooding and the resulting hypoxia in the root zone occurs less frequently on the banks of larger streams and rivers than on lakeshores, the alluvial sediments are coarser-grained and contain less humus, and the riverbanks are affected by flowing water (Kopecky 1966; see Table 4.10). These conditions are less favourable for Phragmites and the tall-sedge species, so that Phalaris arundinacea (reed canarygrass) can dominate (Phalaridetum arundinaceae). At least from an ecological point of view, Kopecky (1966, 1967b) was therefore right to assign the river reedbed communities to their own alliance (Phalaridion arundinaceae), separating them from the Phragmition and Magnocaricion (cf. also Chytrý 2011). It is only in the lower reaches that the differences between the two reedbed types become smaller, as the habitat conditions become increasingly similar (cf. Libbert 1931/1932 and Roll 1939). On the banks of Alpine rivers, Phalaris arundinacea is

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Fig. 4.22  Environmental conditions in communities of flowing water in northwestern Germany; modified from Remy (1993a). High oxygen contents are characteristic for river and stream ecosystems due to the movement of the water, which strongly influences the plant and animal communities that form there. In northwestern Germany, the oxygen saturation is on average 94–98 % and rarely below 60 %. There is even often O2 supersaturation. All of the communities here have a wide amplitude concerning temperature (not shown here). The Myriophyllo-Nupharetum can only develop in slow-flowing or still waters, and grows in warmer habitats (annual average temperature 12.5 °C). The typical communities for fast-flowing upper reaches (particularly the Juncus bulbosus-­Isolepis fluitans community, the Callitricho hammulati-Myriophylletum alterniflori, the Callitricho-­Ranunculetum penicillati and the submerged Sietum erecti) have lower average temperatures (10.5–11 °C). The river and stream macrophyte ecosystems differ most in the conductivity of the water, which is closely related to the carbonate hardness (and the similar water hardness). The soft-water species have the lowest nutrient supply, particularly of phosphate and nitrate. The variation in nitrate and ammonium contents is the widest due to the effects of varying atmospheric deposition and wastewater pollution. The pH is also influenced by a range of factors

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232

Table 4.10  Differences in habitat conditions between still water reedbeds (Glyceria maxima and Phragmites australis communities) and flowing water reedbeds (Phalaris arundinacea communities) Position rel. to water level Flooding regime Grain size in topsoil Aeration and water regime Influence of current Humus content

Still water reedbeds from slightly below mean to high water level long-lasting inundation, dry period few weeks silt to clay enduring anoxia, drought periods rare absent high (22–48 %)

Reedbeds on rivers and streams from low water to mean water level inundation few weeks and abrupt, long dry period sand to gravel anoxia not important, periodic droughts important low (5–15 %)

From measurements in Kopecky (1966) in northern Czech Republic

often replaced by dense stands of Calamagrostis pseudophragmites (Calamagrostietum pseudophragmitis; cf. Mucina et al. 1993) that, like Phalaris, mainly spreads via runners. C. Schmidt (1993) described the bryophyte communities of the river vegetation of the Westphalian uplands (western Germany), assigning them to the class Platyhypnidio-Fontinalitetea antipyreticae, in which he distinguished two orders: 1. The Brachythecietalia plumulosi in mainly mineral-poor rivers on silicate bedrocks, with the single alliance Racomitrion acicularis containing four acidophytic associations, which can be found from the lowlands into the subalpine belt; 2. The Leptodictyetalia riparii in neutral to basic lowland to submontane rivers, containing three alliances: 2.1. The Platyhypnidion rusciformis with one association, 2.2. The Cinclidotion fontinaloidis with three associations in basic streams that often dry out in the summer, 2.3. The Brachythecion rivularis with a single association (Brachythecio rivularis-Hygrohypnetum luridi) that is also the most frequent and widespread river moss association in Westphalia. Stony stream and river banks, as well as boulders that are temporarily exposed above the water also support communities of lichens and mosses. Figure 4.23 shows their vertical distribution in the upper littoral zone (cf. also Wahrenburg et al. 1991).

4.3.4  Communities of Springs The springs forming the origins of streams and rivers are characterised by a diverse group of spring plant communities. They provide the most constant conditions of any Central European habitat, as they maintain almost the same temperature

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Fig. 4.23  Zonation of aquatic bryophyte communities on the banks of the Moselle (western Germany) with a natural cliff face (right) and a stone embankment (left), semi-schematic. Modified from Hübschmann (1967). The habitat conditions between the summer flood water level and maximum flood level are poor for bryophytes. In contrast, newly deposited substrates below the summer flood level are rapidly colonised

in both summer and winter, which is either the same or somewhat higher than the average annual air temperature in the region. The chemical composition of the water also varies relatively little, apart from in karst springs, which are generally unique among spring ecosystems. As the spring water is relatively cool during the summer and warm during the winter, and never freezes, it supports a mixture of cold-stenothermic, alpine-arctic or subalpine-boreal and frost-sensitive or even submediterranean species. Depending on the type of water discharge, the following spring types can be differentiated (based on Thienemann 1922): 1. Rheocrene, in which the water flows directly out of horizontal or slanting layers of substrate and into the valley. Its vegetation is very similar to that of streams. 2. Limnocrene, in which the water emerges from underground aquifers into a pool. Particularly the large karst springs have similar habitat conditions to the oligotrophic to mesotrophic calcareous lakes and pools. 3. Helocrene, in which the water seeps through the soil and forms a marsh. This form is probably the most interesting in terms of its vegetation, which forms strips or concentric rings around the wettest point. At its edge, it becomes similar to swamp and bog vegetation (see Sect. 3.3). More than in still water bodies, the lime content of springs is a decisive factor in the species composition of their vegetation. The extreme groups in terms of lime content are:

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Table 4.11  Overview of the plant communities of springs in the lowland and upland areas of Central Europe Class Montio-Cardaminetea Cardamine amara Saxifraga stellaris Stellaria alsine Poor in lime Not shaded: Order Montio-Cardaminetalia Order Cardamino-Cratoneuretalia Epilobium alsinifolium B Bryum schleicheri B B. pseudotriquetrum Saxifraga aizoides B Scapania undulata B S. uliginosa Alliance Cardamino-Montion Epilobium obscurum Montia fontana subsp. fontana Slightly shaded: Cardamine amara ass. Mostly shaded: Order Cardamino-Chrysosplenietalia Alliance Caricion remotaea Circaea intermedia Lysimachia nemorum Cardamine flexuosa Carex remota (weak indicator)

B Brachythecium rivulare B Rhizomnium pseudopunctatum Rich in lime

B Philonotis fontana B Ph. seriata B Pohlia wahlenbergii etc. Alliance Cratoneurion commutati Arabis soyeri B Cratoneuron commutatum B Cratoneuron filicinum B Philonotis calcarea B Cinclidotus fontinaloides

Chrysosplenium oppositifolium C. alternifolium Stellaria nemorum Plagiomnium undulatum (?)

From data in Hinterlang (1992a, b), Pott (1995), Dierßen (1996a), Rennwald (2000) and others. Further alliances have been identified by Zechmeister (1993) in the alpine belt a not considered as separate alliance in Rennwald (2000), due to weak characterisation

(a) Soft water springs, which are chemically similar to acid-oligotrophic lakes (see the left side of Table 4.11, particularly ‘not shaded’). (b) Hard water springs, in which hydrogen carbonate-assimilating plants form tufa. For helocrene and some rheocrene springs, the hard and soft water types belong to the alliances of silicate (Cardamino-Montion) and calcareous (Cratoneurion commutati) spring communities (see Table 4.11). In addition, light intensity also plays an important role. Both of the spring alliances are (with a few exceptions) only found in open conditions and can occur far above the alpine timber line. The order Cardamino-Chrysosplenietalia described by Hinterlang (1992a, b), the communities of which occur in close neighbourhood to forest communities, tolerates and even needs the shade of the trees to free it from the competition of marsh

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235

species. Other authors assign the basiphytic forest swamps to the alliance Caricion remotae, characterised particularly by Chrysosplenium alternifolium, Carex remota and Cardamine flexuosa. All the communities listed in Table 4.11 belong to the class Montio-Cardaminetea, and are given along with the character species of the higher units. An overview of the classification of spring vegetation in Europe can be found in Zechmeister and Mucina (1994), whilst further regional overviews are given e.g. by Beierkuhnlein and Gollan (1999); Chytrý (2011); Doerpinghaus (2003); Hinterlang (1992a, b); Huml et  al. (1979); Maas (1959); Philippi (1975); Schläfli (1979); Sebald (1975); Zechmeister (1993) and Wey (1988). The ecology of Central European springs, particularly with regard to their chemical properties and hydrology, is discussed by e.g. Schlüter (1970); Warnke (1980) and Beierkuhnlein and Gollan (1999 and further references within). Whilst the spring communities in lowland and montane Central Europe are small in scale, they make up a large proportion of the vegetation above the timber line in the Alps. We will therefore leave the discussion of these communities until the chapters on alpine vegetation, especially as they often mix with alpine grasslands and scree communities (see Sects. 5.3.10.3 and 5.3.10.4). Small-sedge and other communities often linked to spring vegetation are maintained mostly by human activity. If they are not occasionally cut they become overgrown with Phragmites or willow scrub and swamp forest trees. They also share many floristic and ecological characteristics with fen vegetation (Schläfli 1979), which is why these communities are discussed in more detail in the chapters on mires (3.3) and meadows (8.3).

4.4  Adaptations to the Environment 4.4.1  Photosynthesis in Aquatic Plants As carbon dioxide diffuses around 10,000 times slower in water than in air (diffusivity in water = 1.7 10−9, and in air = 1.5 10−5 m2 s−1), intensive photosynthesis can lead to a lack of dissolved inorganic carbon (DIC) in the vicinity of the leaves (Vadstrup and Madsen 1995; van den Berg et al. 2002). Some aquatic macrophytes therefore experience an afternoon depression of photosynthesis. In order to avoid this problem, water plants have developed numerous adaptations to optimise carbon uptake. All aquatic and marsh plants can take up CO2, whilst only some can also take up HCO3−. In order to maintain electrical neutrality in their cells, the uptake of large quantities of HCO3− must be balanced either by its co-transport with H+, Ca2+ or other cations, or compensated for by the elimination of equivalent amounts of OH−. HCO3− uptake and OH− elimination are often spatially separated in submerged plants on the upper and lower surfaces of the leaves. Some species also produce the

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enzyme carbonic anhydrase to catalyse the extracellular conversion of HCO3− into CO2, which is then taken up. The Characeae mainly use HCO3− by dehydrating it extracellularly into CO2 using a proton pump, but in alkaline conditions HCO3−can also be taken up directly (Ray et al. 2003). Intensive photosynthesis can lead to the formation of a carbonate crust or rings around the leaves, as e.g. found on the Characeae (‘stoneworts’). The most effective mechanism to avoid C limitation of photosynthesis is the active transport (i.e. using ATPases) of CO2 or HCO3− through the cell membranes into the chloroplasts, known from cyanobacteria, many algae species and some aquatic phanerogams (e.g. Ruppia cirrhosa) (Madsen and Baattrup-Pedersen 1995; Beer and Rehnberg 1997; Hellblom and Axelsson 2003; Raven 2003). Depending on the relative proportions of HCO3−/ CO2 uptake, Ruttner (1962) distinguished three basic types of C acquisition in aquatic plants: (a) Mainly HCO3− uptake (with a ratio of c. 25:1; including many green algae, around 80 % of submerged macrophytes in neutral to alkaline waters; ‘Scenedesmus type’); (b) Mainly CO2 uptake with some HCO3 (Elodea canadensis and around 70 % of submerged dicots in weakly acidic water, e.g. Myriophyllum spicatum; ‘Elodea type’); (c) Exclusively or mainly CO2 uptake (all aquatic bryophytes, Nymphaea alba, Nuphar lutea, Myriophyllum verticillatum, Callitriche hamulata, Potamogeton polygonifolius, many species of Isoëtes in acidic water; ‘Fontinalis type’). The presence of CO2 causes a decrease in the uptake of HCO3− in the species of groups (a) and (b), as CO2 uptake allows faster growth (Maberly and Spence 1983). The different abilities to use HCO3− is an important factor in the relative frequencies of macrophyte species in weakly acidic to alkaline waters (e.g. Wiegleb 1978; Smits et al. 1988; Madsen and Sand-Jensen 1994). Other means of improving carbon assimilation in aquatic plants include the formation of aerial leaves by emergent macrophytes, or the uptake of CO2 from the carbon dioxide-rich sediment via the roots (e.g. Isoëtes, Lobelia, Littorella, Juncus bulbosus; Brix 1990; Boston et al. 1987; Nielsen and Sand-Jensen 1997). This adaptation is particularly striking in Lobelia dortmanna (water lobelia), which has a thick cuticle but no stomata, and thus absorbs almost all its CO2 via its roots, whether the plant is emergent or not (Pedersen and Sand-Jensen 1992). Some water plants, such as Potamogeton pectinatus, species of Eleocharis and Elodea canadensis, use elements of the C4 photosynthesis pathway. This includes the binding of HCO3− through phosphoenolpyruvate carboxylase and the production of malate or aspartate, but not the formation of the characteristic kranz anatomy of the leaves (Keeley 1990). Some species show CAM-like photosynthesis with nocturnal fixation of CO2 in organic acids, at a time when relatively large amounts of CO2 are available in the water and the sediments (e.g. Isoëtes, Sagittaria, Littorella). This strategy is particularly effective at exploiting low CO2 levels and re-fixing CO2 released by respiration, allowing them to reach a low CO2 compensation point of

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2–12 μM CO2 (compared to 60–110 μM in other aquatic plants; Hough and Wetzel 1977; Sand-Jensen 1987). Whilst emergent aquatic plants and floating-leaved plants have typical characteristics for plants of open light-exposed habitats, such as high light saturation points and high optimum temperature for photosynthesis (25–37 °C; Sale and Orr 1987; Spencer and Bowes 1990), submerged macrophytes behave more like terrestrial shade plants. Their large and thin leaves have relatively low respiration rates and high chlorophyll concentrations, so that some species can still have positive net photosynthesis at only 1–3 % of the light levels above the water (or 20 μmol photons m−2 s−1) (Gessner 1955; Sand-Jensen and Madsen 1991; Spence and Chrystal 1970). In contrast to phytoplankton, neither emergent nor submerged macrophytes show clear photoinhibition. However, the assimilation rate of submerged macrophytes is often considerably reduced by epiphytic algae and microorganisms, particularly in eutrophic water bodies (Sand-Jensen 1990). Some submerged plants have a positive C balance close to freezing point, such as Elodea canadensis and Lemna trisulca, which can assimilate down to around 2 °C below a layer of ice (Boylen and Sheldon 1976; Pip 1989).

4.4.2  Nutrient Uptake by Aquatic Plants The majority of aquatic macrophytes have roots, and many also have rhizomes. They provide not only anchorage but are also an important means of nutrient uptake (Agami and Waisel 1986; Wetzel 2001). Nutrient uptake from the oxygen-poor substrate is facilitated by the fact that the rhizomes are often largely insensitive to hypoxia (Braendle and Crawford 1987). Root hairs are also formed in anoxic substrates, and many species form symbioses with arbuscular mycorrhizae, which profit from the oxygenation of the rhizophere by O2 released from the roots (Stenlund and Charvat 1994). Thus, not only emergent but also many submerged rooted aquatic plants take up their nutrients mainly from the sediment, where the concentrations are much higher than in the water. Aquatic plants therefore usually grow better on silty, nutrient-rich sediments than on sand (Chambers and Kalff 1985; Wilson and Keddy 1985; Lenssen et al. 1999), as the N and P concentrations of sediments increase with the clay and humus content. Accordingly, Lang (1967a) found communities of oligotrophic aquatic plants mainly on gravel-sand substrates on the shore of Lake Constance, and ­communities of eutrophic plants on silt-clay sediments (see Fig.  4.24). Lenssen et  al. (1999) found that the leaf litter of reeds hinders the growth of some helophytes. This is probably due to the phytotoxic effects of the acetic and butyric acids produced in the anaerobic decomposition of the lignin-rich Phragmites litter (Armstrong et al. 1996a, b). Floating macrophytes such as Elodea canadensis and Ceratophyllum demersum absorb their nutrients over the whole surface of the plant, even if they are rooted. As

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Fig. 4.24  Soil type and nutrient content of the sediment for nine communities in the terrestrialisation zone of Lake Constance (from data in Lang 1967b). Left: Relative importance (in %) of five grain size classes (clay, silt, sand, fine and coarse gravel) in the sediments, middle: total nitrogen content, right: total phosphorus content in % of soil dry mass. In the N and P graphs, the right end of the black bar shows the lowest and the right end of the striped bar the highest measured concentrations from two to six experimental plots. The majority of soils consist mainly of sand or silt with low clay contents. The communities of the oligotrophic shore (2 and 4) grow on gravelly soil that is extremely poor in N and P. All other soils are richer in smaller-grained sediments and nutrients. The Phragmitetum (7) can be found on both nutrient-poor and nutrient-rich soils. The best soil conditions are indicated by communities of the eutrophic vegetation zones (12 and 11). The soils of the Najadetum marinae (8) are particularly nutrient-rich. The C content (not shown here) varies similarly to the N content, and the Ca content is high in all soils

their access to the nutrient-rich sediment is limited, they only grow well in water bodies with high nutrient concentrations in the water, and have to compete for nutrients with the often dense covering of epiphytic algae on their leaves (Pelton et al. 1998). Many submerged flowering plants preferentially use ammonium rather than nitrate as a source of nitrogen (Schwoerbel and Tillmanns 1972; Gabrielson et al. 1984; Madsen and Baattrup-Pedersen 1995), including helophytes rooted in anoxic substrates such as Phragmites (Dokulil et al. 2001). In contrast, other aquatic plants, including Potamogeton lucens, are sensitive to high concentrations of NH4+ in the water (Glänzer et al. 1977; Wiegleb 1978) and disappear in hypertrophic conditions.

4.4.3  Survival in Hypoxic Sediments The position of a plant species in the bank vegetation is largely dependent on the flooding-tolerance of its seedlings (Voesenek et al. 1993) and adult plants (Crawford 1996; Blom and Voesenek 1996). Duration and frequency of flooding depend on the

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microtopography at the water’s edge, which therefore determines the vegetation mosaic on the banks of lakes and rivers (Prach 1992). Almost all hydrophytes and helophytes have an aerenchyma, with large air spaces that can fill up to 70 % of the tissue volume and ensure exchange of CO2 and O2 between the leaves and the roots (Armstrong et al. 1994). The formation of this lacuna system is promoted by ethylene, which is produced by the plant in larger amounts under anoxic conditions (Moog and Janiesch 1990; Jackson and Armstrong 1999; cf. Sect. 9.4.1 in Volume I). Oxygen released by photosynthesis or taken up by emergent leaves is transported to the roots or rhizomes by thermo-osmosis and pressure flow (Ricard et al. 2006), while carbon dioxide from root respiration or photorespiration in the leaves is re-fixed by photosynthesis. Some aquatic plants thus recycle a large proportion of the CO2 and O2 within their tissues. Considerable amounts of oxygen (up to 0.9 μmol O2 m−2 s−1) may be released by the roots into the surrounding anoxic sediment, where they oxidise a layer of up to 0.5 mm thick and can lead to the precipitation of Fe(OH)3 in the form of iron plaques around the roots (e.g. in Juncus bulbosus and Potamogeton pectinatus) (Caffrey and Kemp 1991). This transport of O2 into the sediment improves the nutrient supply, as it promotes aerobic mineralisation and improves the conditions for mycorrhizae. Even if the roots receive insufficient oxygen, aquatic plants can survive for long periods without apparent damage by forming adventitious roots that avoid the oxygen-­poor sediments (Barklay and Crawford 1982; Ricard et al. 2006). They are also able to avoid the toxic products of glycolysis (such as ethanol) by excreting them or producing non-toxic metabolites such as malate and shikimate (Bertani et al. 1980; McKee et al. 1989; Crawford and Braendle 1996). Toxic ammonium and sulphide are detoxified in the plant to alanine and thiols (Weber and Braendle 1996).

4.4.4  Adaptations to Flowing Water and Wave Action Flowing Water  Aquatic plants are exposed to more drag and shear than plants of other habitats, both from currents and from wave action. The studies of Wiegleb (1984) and Steffen et al. (2014) showed that the water flow is a more important factor determining the composition of river communities than the chemical properties of the water. In streams and small rivers, Pott and Remy (2000) measured average speeds in the current thread of between 5 and 60 cm s−1 in lowland and between 40 and 110  cm s−1 in upland areas, whilst large lowland rivers and high mountain streams with greater gradients have even higher velocities. These average speeds can increase by up to 50–100 % during flooding. Directly above the river bed, the internal friction of the water causes a laminar boundary layer to form with much slower flow than the rest of the stream. The depth of this layer depends on the free stream velocity, the distance downstream from the start of the boundary layer, and the viscosity of the water, and in small streams can be just a few millimetres thick. Aquatic mosses and epilithic or epiphytic algae exploit this quieter zone and form dense layers on the surface of the river bed.

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Table 4.12  Field observations of tolerance to flowing water in selected aquatic plant species. Tolerance to flowing water High

Average Low Non-existent

Species Ranunculus fluitans, R. trichophyllos, R. penicillatus, R. peltatus, Callitriche hamulata, Myriophyllum alterniflorum, M. spicatum, Potamogeton pectinatus, Fontinalis antipyretica Callitriche platycarpa, C. stagnalis, Potamogeton perfoliatus, P. crispus, Groenlandia densa, Zannichellia palustris, Nasturtium officinale Nuphar lutea, Sparganium emersum, Sagittaria sagittifolia, Ranunculus circinatus, Callitriche obtusangula, Elodea canadensis, Butomus umbellatus Nymphaea alba, Lemna sp., Spirodela polyrhiza

Excerpt from Pott and Remy (2000), with additional data from Ackenheil (1944) and Sirjola (1969)

Higher plants only tolerate flows of up to c. 110–120 cm s−1 (Gessner 1955; Jorga and Weise 1977), and above this only bryophytes and algae can colonise. Table 4.12 lists various aquatic plants that have been classified from ‘highly tolerant’ to ‘intolerant’ based on field observations of their sensitivity to flowing water. Macrophytes minimise the dynamic pressure and the frictional forces on their biomass within the water by adopting long and thin hydrodynamic shapes, such as seen in Potamogeton pectinatus. They can also strengthen their sclenenchyma fibres and form additional layers of epidermal cells to increase their resistance to pressure without losing the elasticity of their tissues (Haslam 1987; Brewer and Parker 1990). In this way, the stream form of the aquatic moss Fontinalis antipyretica (with 5 layers of epidermal cells) has a tensile strength of 535 g mm−2, whilst its still-water form (1 layer) breaks at only 351 g mm−2 (Fuchsig 1926). Wave Action  Mechanical forces also play an important role in the littoral zone of still water bodies. The zonation of the reedbed is determined not only by the water level and the type of lake substrate, but also to a large extent by the intensity of wave action (Spence 1982). In shallow lakes such as Lake Neusiedl, the shoreline vegetation is furthermore influenced by the degree of coverage with silty sediments. This factor is dependent on the movement of the water and leads to a reduction in light levels (Schiemer 1979). The importance of wave action for the vitality of reed plants is demonstrated by the experiments of Coops et al. (1996) in lakes in the Netherlands. They found that Schoenoplectus lacustris, Phragmites australis, Bolboschoenus maritimus and Phalaris arundinacea colonise increasingly shallow littoral zones (S. lacustris: 69 ±19 cm water depth, P. australis: 45 ±20 cm, B. maritimus: 36 ±8 cm, P. arundinacea: 25 ±8 cm). Growth experiments without wave action showed that Schoenoplectus lacustris, which grows in quiet inlets at greater depths than any of its competitors, does not suffer any reductions in growth rate up to depths of 80 cm. Phragmites australis and Bolboschoenus maritimus are more sensitive and did not grow well at 80 cm, whilst particularly Phalaris arundinacea was weakened at depths of only 30 cm. However, if wave action was present then Phragmites australis and Bolboschoenus maritimus outcompeted Schoenoplectus lacustris at depths of 80 cm, which follows the pattern of their natural distribution.

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The high sensitivity of Schoenoplectus lacustris to wave action is explained by Coops and van der Velde (1996) by its lower tolerance to bending compared to P. australis (0.15 compared to 0.38 N m−2) and lower elasticity modulus (0.25 compared to 4.93 GPa). Phragmites thus has a higher productivity as well as greater stability in shallow water than Schoenoplectus lacustris. However, Phragmites is damaged by strong wave action and ice drift, as e.g. in the Bodden lagoons of northeastern Germany or Lake Constance in the south, and in the Baltic it is replaced by Bolboschoenus maritimus and Schoenoplectus tabernaemontani (see Sect. 1.3.2.2). Submerged macrophytes are constrained by wave action, in that they shorten their growing season to avoid stormy periods and thus suffer a loss of productivity (as is the case for e.g. Potamogeton pectinatus; van Wijk 1988). On the other hand, the movement of the water by waves or currents reduces the growth of epiphytes and removes the film of sediments that forms on the plant. In some places, exposed sediments thus support greater macrophyte biomass than still areas, especially as the oxygen supply to the sediment is better in turbulent areas (Strand and Weisner 1996).

4.4.5  L  ife Forms and Morphological Adaptations of Aquatic Plants Life Form Classification  Based in part on Hejny (1960); Mäkirinta (1978); Wiegleb (1991); Pott and Remy (2000) and Wetzel (2001), we can distinguish the following habitat-related life forms of aquatic macrophytes: I. Free-floating (pleustophytes): 1. submerged, free-floating plants that at most expose their generative organs above the surface of the water (e.g. species of Utricularia and Elodea) and use dissolved CO2 or HCO3−; 2. plants floating on the surface of the water, the leaves of which are in contact with the air and can (at least occasionally) absorb gaseous CO2 (e.g. Stratiotes, Salvinia, Hydrocharis and the majority of Lemna species). II. Rooted in the substrate (true hydrophytes): 1. plants that assimilate only under water and use dissolved CO2 or HCO3−, (a) only benthic (on the lake or river bed) and forming low growing underwater stands (e.g. Isoëtes and species of the Characeae), (b) reaching from the sediment into open water (e.g. Zannichellia, Myriophyllum and many species of Potamogeton); 2. plants assimilating both above and below the surface of the water. Their leaves either float on the surface of the water and take up gaseous CO2 or remain below the surface, and can tolerate alternation between the two states (e.g. species of Nuphar and Nymphaea, Fig. 4.1).

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III. Riparian vegetation (littoral helophytes, or amphiphytes sensu Pott and Remy 2000). These plants live in the changing conditions of the littoral and are usually heterophyllous, i.e. they have both emerged and submerged leaves, and can survive periods of being submerged in still water. The following types can be distinguished: 1. plants that can also assimilate under water using dissolved CO2 or HCO3−, and their submerged leaves have similar photosynthetic capacities to true hydrophytes (e.g. Hippuris, Hottonia, Alisma, Sagittaria, Littorella, Polygonum amphibium, Glyceria fluitans); 2. plants assimilating mainly above water, and their leaves have a much reduced photosynthetic capacity under water and usually die (including many reed species dependent on still water such as Eleocharis palustris, Veronica beccabunga, Myosotis palustris; these are equivalent to the amphiphytes of Pott and Remy 2000). IV. Marsh plants (true helophytes); these colonisers of wet, often waterlogged soils do not depend on temporary standing water, but can survive several weeks of flooding. They are land plants, i.e. do not show heterophylly, and include: 1. plants capable of assimilating under water, but with a lower photosynthetic capacity (e.g. Agrostis stolonifera, Rorippa amphibia, Juncus bulbosus, species of Schoenoplectus, Nasturtium officinale); 2. plants that only assimilate above water (the majority of species of wet sedge communities and drought-tolerant reed species such as Phragmites australis, Phalaris arundinacea, Carex elata, C. acuta, C. paniculata, Typha species, Sparganium erectum, Glyceria plicata). As shown early on by Glück (1936) through observations of disturbed habitats and growth experiments, the majority of species are quite tolerant of different habitats or water conditions and could be assigned to several of the above-listed groups. However, competitive pressure means that they are usually in nature restricted to only one type of life form. Foliar Adaptations  The submerged or floating leaves of aquatic plants have a number of characteristic adaptations absent in terrestrial plants (cf. Sculthorpe 1967). The aerenchyma enables gas exchange and provides the plant with buoyancy, whilst vascular and supporting tissues are greatly reduced or even absent. The poor availability of inorganic carbon is compensated for by an increase in leaf surface area through divided leaves, and the placement of chloroplasts in the epidermis to reduce the length of the diffusion path. The hydropotes on the undersides of floating leaves e.g. in Nymphaea and Nuphar species serve to improve nutrient uptake from the water. These are 0.05 μm wide pores in the epidermis surrounded by a ring of cells capable of active element uptake (Lüttge 1964). Heterophylly, i.e. high plasticity in leaf morphology within a single individual, improves the chances of survival under fluctuating water levels. A change from simple to dissected leaves can be caused by the age of the plant, the depth of the

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water, the current, the temperature, the CO2 content of the water, the light levels or other environmental stimuli (Bristow 1969; Johnson 1967; Bodkin et al. 1980; Pott and Remy 2000).

4.5  Population Biology and Community Ecology 4.5.1  Phenology Whilst emergent helophytes are mostly perennial, submerged and floating aquatic plants include both annual and perennial evergreen species or ecotypes. Within the genus Potamogeton, for example P. friesii, P. pusillus and P. compressus are usually annual, whilst P. crispus, P. pectinatus and P. natans are mostly perennial (Pott and Remy 2000). Challenging environmental conditions such as the complete freezing of shallow lakes can, however, also induce the latter species to live as annuals and survive with overwintering buds (turions) or seeds (van Wijk 1988). The perennial Stratiotes aloides overwinters by allowing its rosettes to sink to the bottom in autumn and re-emerge from the mud in spring (Casper and Krausch 1980). In the case of Nuphar lutea, only its lower submerged leaves overwinter, whilst its floating leaves die in autumn (Pott and Remy 2000). This annual rhythm of aquatic plants is driven by day length and water temperature (see Fig. 4.25).

4.5.2  Life Cycles of Aquatic Plants Vegetative Reproduction  The population growth and dispersal of aquatic macrophytes occurs mainly through vegetative or clonal reproduction, whilst sexual reproduction is of lesser importance. Many species reproduce by the formation of clones, i.e. the active fragmentation of above-ground organs (as e.g. in Hydrocharis morsus-ranae) or the breaking up of rhizomes (e.g. Nuphar lutea, Potamogeton pectinatus, Phragmites australis or Typha species). Moving water also causes the passive fragmentation of above-ground organs, e.g. in many species of Characeae (van Wijk 1989; Wiegleb and Brux 1991; Wiegleb et al. 1991a). A characteristic form of vegetative reproduction is the production of overwintering buds (turions), i.e. dormant buds formed from modified shoot apices, or of bulbils, i.e. leafless axial or rhizomal organs (e.g. in Lemna, Spirodela, Utricularia and Myriophyllum species, Stratiotes aloides or Potamogeton pectinatus). This adaptation allows the plant to survive periods of hostile conditions, such as extreme cold or lack of oxygen. The formation of turions is induced by temperatures between around 10 and 15 °C in autumn; they form a new plant once temperatures rise again in spring. Wind pollination of flowers extending above the surface is the main form of pollination in aquatic plants. Water pollination occurs in less than 5 % of aquatic

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Fig. 4.25  Phenological development and life forms of higher aquatic plants in Lake Constance in relation to water temperature (Modified from Lang 1967b). Zannichellia repens is a subspecies of Z. palustris

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Table 4.13  Seeds from various plant communities eaten and dispersed by ducks. Modified from observations by Olney (1963–1967) cited in Gillham (1970)a Hydrophytes   Ceratophyllum demersum   Hippuris vulgaris   Myriophyllum species   Potamogeton, many species   Ranunculus aquatilis   Ruppia species Plants of reedbeds   Bolboschoenus maritimus   Glyceria fluitans   Phalaris arundinacea   Phragmites australis   Schoenoplectus lacustris   Sparganium species Halophytes   Armeria maritima   Salicornia species   Suaeda maritima

Plants of marshes   Alopecurus geniculatus   Cirsium palustre   Eleocharis palustris   Galium palustre

Ruderal and crop plants   Atriplex patula   Bromus sterilis   Chenopodium album   Galium aparine

  Juncus inflexus and other species   Polygonum amphibium

  Hordeum distichon (barley)

  P. hydropiper, P. lapathifolium etc.   Ranunculus repens   Rumex conglomeratus   Taraxacum palustre Other plants of grasslands   Carex hirta and other species   Holcus lanatus   Lolium multiflorum   Medicago lupulina   Phleum pratense   Poa trivialis and other species

  Triticum aestivum (wheat) etc.

Woody plants   Alnus glutinosa   Betula species   Crataegus monogyna   Quercus species   Rosa species   Rubus species   Sambucus nigra etc.

In addition to this list, K. Müller (1973) noted numerous species frequently found in areas often used by ducks in bogs in northwestern Germany, including: Bidens cernuus, B. tripartitus, Holcus lanatus, Juncus bufonius, Lemna minor, Lycopus europaeus, Myosotis palustris, Peplis portula, Poa annua, Polygonum persicaria, Ranunculus flammula, R. hederaceus, Rumex acetosa, R. acetosella, Urtica dioica

a

flowering plants (Cox 1993), for example in Callitriche hamulata and Zannichellia palustris. In contrast, water plays a central role in the dispersal of propagules in streams and rivers (Cellot et  al. 1998; Johansson et  al. 1996) as well as in lakes (Smits et  al. 1989). Air sacs attached to the seeds (e.g. in Ranunculus and Myriophyllum species) increase their buoyancy, and they can float for several days to months in the water (Sculthorpe 1967). However, the most common and long-­ distance means of transport for seeds or vegetative propagules of aquatic plants are water and wetland birds such as ducks (van der Pijl 1969; Landolt 1984; Krahulec and Leps 1993; see Table 4.13) as well as fish and muskrats. Birds are a large part of the reason why the plant communities of similar types of wetlands across Europe, and even the world, are more uniform than any other type of vegetation. They also contribute to the rapid colonisation of newly created habitats. Anemochory (wind dispersal) only plays a role for a few marsh plants (e.g. Phragmites and Typha) as well as in the colonisation of pioneer habitats on riverbanks and shorelines (Bonn and Poschlod 1998).

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Although some hydrophytes produce large numbers of seeds each year, the germination rate is usually low (Rogers and Breen 1980; Brux et al. 1987; Kautsky 1987). This was shown by the experiments of Wiegleb et al. (1991a) for Potamogeton polygonifolius, P. natans, P. lucens and P. alpinus. Helophytes germinate best in damp soils of the occasionally flooded littoral zone (Bonn and Poschlod 1998). The dormancy of the seeds is broken particularly by changes in temperature (Schütz 2000), as well as changes in water levels (Stockey and Hunt 1992). Flooding reduces the germination rate, and the seedlings of marsh plants tolerate, depending on the species, up to 8 weeks of being covered by water (Lenssen et al. 1998). Reproduction of Phragmites  Previous claims that highly productive reed does not reproduce well sexually (e.g. Bittmann 1953) are, according to Ekstam and Forseby (1999), incorrect. Phragmites produces large numbers of seeds (Hürlimann 1951) that are blown by the wind over similar distances to the long-tasselled seeds of Typha. The non-dormant seeds need a wide variation in temperature to germinate, which prevents germination in autumn and winter, as well as under water, but do not necessarily need high temperatures (Ekstam and Forseby 1999). However, attack by fungal species (Clavipes microcephala; Luther 1950) can prevent the seed ripening. The seeds remain on the dead spikes of the mother plant until the end of April. As they can float for several days on the surface of the water and need full light levels, constant moisture and high levels of oxygen to germinate (Hürlimann 1951), they are most likely to colonise drift lines on bare riverbanks or lake shores (Weisner et al. 1993). The resulting plants then move further towards the water via stolons. Wide belts of reedbed of several thousand m2 have developed from a single successful germination (Neuhaus et al. 1993). These findings emphasise the importance of vegetative propagation of water and wetland plants. The vegetative growth of reeds only begins in spring at temperatures of around 8–10 °C, so that the young shoots only appear well after the surrounding forests and grasslands have begun to grow. In experimental tanks without standing water, Dykyjova et al. (1971) found that Phragmites was killed in January at −8.7 °C, and in March at −0.8 °C (Typha latifolia and Schoenoplectus lacustris showed similar reactions). Although Phragmites and other reed species need a certain period of winter dormancy in order to develop normally (Dykyjova et al. 1972), its need for warmth means that it does not occur further than northern Fennoscandia or above the montane belt. In the subtropics, in contrast, it grows well and can be found across the world. Aquatic Neophytes  The aquatic flora of Central Europe includes many non-­ indigenous plants, and 22 species are now considered naturalised (www.aquatischeneophyten.de; cf. Pott and Remy 2000). Based on abundance, the most important of these are the species of Elodea (waterweed), and E. canadensis also serves as a character species in Central European macrophyte communities. These pleustophytes originate from North America and can spread rapidly via vegetative reproduction. The ‘Canadian’ waterweed first spread to the British Isles in 1834, and soon conquered the rest of Europe within an alarmingly short space of time, despite the fact that only female individuals of this dioecious member of the Hydrocharitaceae were originally transported via plant fragments. Since around 1880, E. canadensis

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Fig. 4.26  Top: A comparison of duckweed forms: Wolffia arrhiza (rootless), Lemna species (a single root) and Spirodela polyrhiza (many roots) (Modified from Wolek 1974). Bottom: Lemna gibba grows less well than Spirodela in pure culture, but in mixed cultures, L. gibba can still outcompete Spirodela. The large air sacs on its leaves mean that it floats higher on the surface, so in dense populations it can push itself on top of Spirodela (Modified from Clatworthy (1960) in Harper (1961))

has spread less aggressively in Central Europe and has even become rare in places, although the causes of this change in behaviour have still not been fully explained (see Mabberly 1990). Today, increasing attention is being paid to E. nutallii, which also arrived via Great Britain as a female plant in 1966. It has invaded similarly rapidly to E. canadensis e.g. around Hanover (Lower Saxony; Seehaus 1992), and is already being used as a character species for pondweed communities.

4.5.3  Interspecific Competition Between Aquatic Plant Species Pleustophytes and Hydrophytes  The small size and rapid growth of duckweed species has lent itself to numerous experiments on competitive behaviour among aquatic plants. According to Harper (1961), Spirodela polyrhiza shows higher productivity per unit area of water surface than Lemna gibba. Nevertheless, the latter outcompetes the former in mixed cultures, as it pushes itself slightly above the surface of the water and thus shades out the low-floating Spirodela (see Fig. 4.26). Clatworthy and Harper (1962) additionally included Lemna minor and Salvinia natans in their experiments. In pure culture, Lemna minor had the fastest growth in biomass of all species, explaining its high frequency and relatively large range. However, in mixed cultures it is outcompeted by all of the other three, and only

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Fig. 4.27 Seasonal changes in the total biomass (above– and below-ground: thick line) and the leaf area index (thin line) of Phragmites australis (solid line) and Typha latifolia (dashed line) in eastern Czech Republic (Modified from Kvet et al. 1969)

survives in the gaps of the duckweed community. Salvinia dominates the community if the temperature is high enough, as its stems and leaves create dense and impenetrable groups of individuals. The opposite extreme is the thermophilic subtropical rootless duckweed (Wolffia arrhiza, Fig. 4.26). Wolek (1974) compared this to the other duckweed species, whereby it was the weakest partner in all species combinations, which may explain why it is so rarely present in duckweed communities. Wolek (1974) states that the outcome of competition is mainly dependent on the morphological and physiological characteristics of the duckweed species, and allelopathy only plays a very small role. In the competition between submerged plants, it appears to be inorganic carbon that is the most limiting resource. For example, species of the Characeae are the strongest competitors for HCO3− (van den Berg et al. 2002), whilst species of Isoëtes use CO2 from the sediment and can be disadvantaged by eutrophication and increased DIC contents. Some submerged angiosperms can quickly grow to the surface in spring and monopolise the light supply (Barko and Smart 1981). The effect of mild or severe winters on which species dominates in the following summer, depends mainly on their overwintering form (Idestam-Almquist 1998). Helophytes  Interspecific competition is also extremely important in the productive reedbed zone (Keddy 2000). In mixed reed stands, Glyceria maxima can outcompete the much higher-growing Phragmites if the sediment dries out in spring, as Glyceria starts growing much earlier. River banks or lake shores that temporarily dry out can also be dominated by species of Typha, as it reaches its maximum biomass earlier in summer than Phragmites (see Fig. 4.27). However, Phragmites is a much stronger competitor if water levels are high (Buttery and Lambert 1965; Buttery et al. 1965). Allelopathy may also play a role in the interactions between Phragmites and Typha (Szczepanska 1971; cf. also Gopal and Goel 1993). Typha latifolia, with its large, strongly shading leaves, only outcompetes the thinner-leaved T. angustifolia when water levels are low. In deep, nutrient-poor water, T. angustifolia is superior as its smaller leaves are more economical in terms of the carbon return of leaf mass production (Grace and Wetzel 1981, 1998). However, this situation is reversed in hypertrophic lakes (Weisner 1993). In the competition between Phragmites and Schoenoplectus lacustris, it is their sensitivity to wave action that is decisive (see Sect. 4.4.4).

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4.6  Productivity and Cycling of Water and Nutrients 4.6.1  Productivity Aquatic and wetland plant communities are among the most productive vegetation types in Central Europe. This is, however, only the case for the emergent reed and tall-sedge stands. The submerged macrophytes have a much lower biomass production and are relatively unproductive in oligotrophic and mesotrophic water bodies (Sculthorpe 1967). Phragmition and Magnocaricion communities have an unlimited supply of water and high N and P availability in the sediment at high light levels. Hence, they also maintain high photosynthetic rates and develop large leaf areas. Photosynthesis is high partly because reed plants can use considerable amounts of root-borne CO2 supplied through their aerenchyma. Light-saturated net photosynthesis rates in Phragmites can reach 25 μmol CO2 m−2 s−1 (Lessmann et al. 2001). Submerged macrophytes are less productive than emergent helophytes, as the light levels and supply of inorganic carbon in the water are lower than in the air. Biomass  Reported figures for the biomass of Phragmites, Scirpus and Typha reedbeds vary between around 500 and 4000 g dry mass per m2 for the above-ground parts (see Table  4.14). Phragmites is a highly polymorphic species with variable ploidy (n = 3 to 8), which can produce very variable stem size and plant biomass depending on the depth of the water, its salinity and the sediment type (Rodewald-­ Rudescu 1974; Lissner and Schierup 1997; Vretare et al. 2001). More calcareous sediment leads to softer reeds, and humus-rich sediment similarly produces tall but weak stems, whilst silica-rich sediment causes stronger stems to form, and brackish water leads to low-growing but strong stems (Rodewald-Rudescu 1974; Hanganu et al. 1999). The above-ground biomass of Magnocaricion communities, which have a constant water supply, is closely correlated in the temperate zone with the summer temperature (Gorham 1974). This ranges from around 1500 g m−2 in the relatively warm lowlands to 200 g m−2 in montane or subarctic zones. However, these figures are only for relatively fertile habitats (cf. Ambroz and Balatova-­ Tulackova 1968).

Table 4.14  Maximum recorded biomass (dry mass) of submergent and emergent aquatic plant communities in Central Europe Species Elodea canadensis, above-ground (CZ) Phragmites australis, above-ground (CZ) Phragmites australis, above-ground (CZ) Phragmites australis, above-ground (CZ) Typha angustifolia, above-ground (CZ) Schoenoplectus lacustris (CZ) Emergent species, above-ground (PL)

Biomass (g m−2) 450 2980 1100–2200 6000–8560 4040 3000 440–830

Author Pokorny et al. (1984) Dykyjova (1971) Dykyjova and Hradecka (1973) Dykyjova and Hradecka (1973) Dykyjova (1971) Dykyjova (1971) Szczepanska (1973)

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Above-ground biomass alone is, however, a poor measure of productivity, as many species grow new leaves several times a year. For example, Typha and Glyceria produce new leaves roughly three times a year, as found by Mathews and Westlake (1969) using demographic models. A large proportion of the primary production is also invested in the roots and rhizomes: of the total biomass, around 30–60 % is found below-ground in Typha, Glyceria maxima and Carex rostrata stands, 40–95 % in Phragmites and Schoenoplectus lacustris stands, 50–80 % in water lily communities and even 20–45 % in Potamogeton stands (based on the literature review by Wetzel 2001). A detailed study of a reedbed in the Czech Republic found 6000–8560 g m−2 below-ground dry mass (Dykyjova and Hradecka 1973), which can be extrapolated to a total biomass of 7000–11,000 g m−2 in relatively productive Phragmites stands. Studies of submerged macrophyte stands have shown highly variable, but consistently low biomasses of less than 1 g m−2 in a Myriophyllum spicatum stand in Lake Neusiedl up to 450 g m−2 in an Elodea canadensis community in a eutrophic pond in the Czech Republic (see Table 4.14). Productivity  Numerous authors have published information about the productivity of reedbeds. Sieghardt (1973) found that Phragmites in Lake Neusiedl had an annual above-ground productivity of around 1700 g dm m−2, which comes close to that of a productive wheat field, and uses around 5 % of the photosynthetically active radiation. According to Dykyjova (1971), up to 6 % of the PAR received during the growing season is used in photosynthesis (only above-ground production), corresponding to a productivity (above- and below-ground) of 8740 g m−2 year−1. In the Danube Delta, Rodewald-Rudescu (1974) even recorded production levels of 14,300 g m−2 year−1. At the other end of the spectrum, Phragmites in the littoral of the relatively nutrient-rich Lake Balaton produced only 5–30 g m−2 year−1 (Karpati and Karpati 1971). Generally, it can be said that the denser the stand, the greater the leaf area index (up to a maximum of 7 m2 m−2, Ostendorp 1993), and the greater the productivity. The stem density and the production decrease from this maximum both further inland and into the deeper water (Ondok 1970). However, few reliable average production values are available, as the yield varies with the weather conditions from year to year. In a reedbed in England, for example, Mason and Bryant (1975) recorded 1080 g m−2 in 1972, but only 550 g m−2 in 1973. In the Danube Delta and other wetlands, reedbeds are intensively used for harvesting cellulose. According to Rodewald-Rudescu (1974), these stands have been cut for over 100 years without experiencing a decrease in productivity. In contrast, Granéli (1990) states that the above-ground biomass doubles if the stands are cut, as the shoots are not shaded by the old stems. The pronounced plant-internal reallocation of carbon in Phragmites also plays a role in its constant high productivity, as in autumn (before cutting) around two thirds of the mobile carbohydrates and a large proportion of the stored nutrients are transported to the rhizome (Sieghardt 1973; Westlake 1975; van der Linden 1986). The rhizomes themselves always contain more N, P and K than the stems and leaves (Dykyjova and Hradecka 1976). To a certain extent, Phragmites is a natural monoculture, as it can dominate over many square kilometres without the presence of any other similarly tall plants

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(e.g. in the mostly shallow Lake Neusiedl). However, according to Straskraba (1963), Phragmites was only responsible for 70 % of the primary production of a moderately eutrophic fish pond in the southwest of the Czech Republic, whilst a further 21 % was provided by the algae covering its submerged stems and 7 % by the phytoplankton in the water. In the large reedbeds of Lake Neusiedl, the submerged bladderwort (Utricularia vulgaris) also provides a certain proportion of the total primary production (Maier 1973). The proportion of net primary production contributed by the roots is often very high in littoral communities, e.g. 88 % in a Carex rostrata stand (Saarinen 1996). However, the C allocation to the roots and rhizomes decreases at low redox potentials in the sediment (Day and Megonigal 1993; Vretare et al. 2001, cf. also Coops et al. 1996). Wetzel (2001) reviewed the published productivity data from temperate lakes and ponds, finding that submerged macrophytes in oligotrophic lakes produce around 500 g dm m−2 year−1, those in eutrophic lakes up to 1000 g m−2 year−1, and emergent helophytes produce between 2000 and 4500 g m−2 year−1 (only above-­ ground production). The productivity of Magnocaricion communities does not reach the upper extremes of the Phragmition, but on average it is about the same. Baradziej (1974) recorded above-ground net primary production for a Carex acuta community in Poland of 550 g m−2 year−1, whilst stands with a large proportion of Iris pseudacorus achieved an NPP of 800 g m−2 year−1. Other Carbon Fluxes  The earlier assumption that aquatic plants are less affected by herbivory than terrestrial plants has been disproved. Lodge et al. (1998) found that aquatic macrophytes lose on average 30 % of their annual production to herbivory, mainly to birds and fish. A large proportion of the primary production of the phytoplankton is released into the water by the algae in the form of dissolved organic carbon (DOC). This can often be up to 20 % of NPP (Søndergaard and Schierup 1982; Brock 1984) and is used by bacteria as a source of energy. These are then consumed by protozoa, the exudates and faeces of which are then in turn taken up by the bacteria again. Most of the carbon in freshwater water bodies therefore does not circulate within a long food chain from phytoplankton to vertebrates as top predators, but is ‘trapped’ in a closed loop of autotrophic and heterotrophic plankton (microbial loop, Fig. 4.28). The ratio of productive (carbon assimilation) to decomposing (respiration) processes follows a characteristic sequence along the length of a river from the source to the mouth. The turbulent headwaters receive large amounts of external organic material (e.g. the leaf litter of riverbank trees), so that more is decomposed than is produced by the aquatic plants themselves. The import of dead organic matter from the surroundings provides 50–99 % of the carbon accumulated in the ecosystem of the upper reaches of streams, whilst the aquatic macrophytes usually provide less than 20 % (Allan 1995). In the middle reaches, it is primarily the autochthonous production by the dense riverbank vegetation that drives the flux of carbon to the biotic community in the river. The lower reaches are instead dominated by carbon

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Fig. 4.28  Simplified schematic diagram of limnic food webs in a lake, focusing on the ‘microbial loop’ (thick arrows), in which heterotrophic bacteria consume dissolved organic matter from the water and make them available for higher trophic levels (Modified from Schwoerbel 1999)

loss through heterotropic respiration, fuelled by deposited organic material in the fine sediment. The carbon cycle of rivers is thus tightly linked to their banks and surrounding terrestrial areas. This linkage led Vannote et al. (1980) to formulate the river continuum concept, i.e. that a stream or river is an open system with a strong gradient along its length in terms of production and decomposition. Nutrients do not circulate within a small area as in lakes and terrestrial ecosystems. Instead, transport by the flowing water means that uptake, incorporation and re-mineralisation are spatially separated, and the nutrient fluxes spiral down the river (‘nutrient spiralling’, Newbold et al. 1981).

4.6.2  Water Cycling Productive plant communities like reedbeds also have a high water consumption. In a Czech reedbed, maximum transpiration rates of 6.9 mm d−1 were recorded (Smid 1975). Koerselman and Beltman (1988) found maximum values of 3.9 mm d−1 in

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the Netherlands, whilst Kiendl (1953) and Herbst and Kappen (1993) even found 16 and 15  mm d−1, respectively, in northern Germany. Phragmites regularly shows transpiration rates up to 1.1 to 3.4, and in extreme cases even up to 12 times higher than the evaporation from open water (Rodewald-Rudescu 1974, Price 1994  in Wetzel 2001). This is despite the fact that the leaf area index is lower than that of many forests (see Sect. 4.7.1.1  in Volume I), namely usually only 4 to 5. In the Danube Delta, the transpiration of a reedbed between March and November was 770 mm and the evaporation from the water’s surface was a further 580 mm, i.e. much higher than the precipitation falling there. Phragmites, Schoenoplectus and Typha have a similar functional organisation to littoral helophytes (or amphiphytes), in that their above-ground parts can be highly xeromorphic and their stomata are sensitive to air humidity, like those of many terrestrial plants (Krolikowska 1975). Water lilies transpire somewhat less than some helophytes (Gessner 1951), but as much or even more than most terrestrial plants. It is interesting to note that the water transpired from the upper surface of their leaves is usually replaced by that taken up by hydathodes on the underside of the leaf, or from the rhizome supplied by root pressure. A water potential gradient caused by leaf water deficits is clearly less important here. If the stem of a water lily leaf is cut and its leaf surface is lifted from the surface of the water, it will wilt even if the stem is still in the water. If it is still connected to the rhizome, then it will retain its turgor, as long as the air is not too dry. Thus, leaves of the Nymphaeaceae can gradually change between floating and fully emergent forms without sustaining damage if, for example, the water level sinks rapidly. Many other aquatic plants, particularly those with submerged leaves, would quickly dry out if exposed to the air.

4.6.3  Nutrient Cycling Phosphorus  The availability of P, the most important growth-limiting nutrient in fresh water bodies, is highly dependent on the oxygen content of the sediment. If the surface of the sediment is oxidised, then it acts as an effective phosphorus trap, in that P is precipitated as Fe(III) hydroxophosphate (Fe(OOH)~P). In contrast, anoxic conditions in the hypolimnion and reducing conditions at the surface of the sediment lead to the reduction of Fe3+ and the release of orthophosphate (see Sect. 4.8.1). If the redox potential decreases even further, then Fe(II) precipitates as FeS taking the bound phosphorus with it, and the P concentration in the water then decreases again. The availability of P in water is therefore controlled by the interplay between P release from the sediment and P precipitation. Precipitation is caused not only by FeS formation but also calcite formation as part of biogenic decalcification (see Sect. 4.2.2). These interactions regulate P availability in lime-­ poor water bodies, whilst in calcareous water with low redox potential, P is bound as apatite and therefore its availability is always low. Aquatic plants have a large influence on the phosphorus cycle within water bodies. P uptake by phytoplanktion reduces the soluble reactive phosphorus (SRP) in

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many lakes in summer to below the detectable limit (around 1 μg l−1), showing its importance as a growth-limiting factor. The P turnover through the planktonic food web is therefore extremely quick, and the entire SRP pool in a lake can circulate within a few hours (Mulholland et  al. 1985). This turnover is accelerated by the microbial loop in aquatic food webs (see Fig. 4.28). Plant-Induced Change in Trophy  Aquatic macrophytes can have a lasting influence on nutrient levels, and particularly the phosphorus cycle, in the water. Once rooted macrophytes have established in a lake, then they can in some cases accelerate its eutrophication, as in contrast to the phytoplankton, they exploit the rich nutrient supply in the sediment. After the leaves have died off, the phosphorus and other nutrients are released into the water. This macrophyte effect can be described as ‘autogenic eutrophication’ (Granéli and Solander 1988). However, the production of large amounts of biomass also means that the macrophytes fix a lot of P, resulting in a reduction in nutrient availability, as shown by the occurrence of clear water stages in some lakes (see Sect. 4.7.1). Macrophytes of oligotrophic waters such as Isoëtes spp. (quillworts) and Juncus bulbosus can lead to an ‘autogenic oligotrophication’, as they oxidise the sediment through the release of oxygen by their roots (Smits et al. 1990; Pedersen and Sand-­ Jensen 1992). According to Tessenow and Baynes (1978), Isoëtes communities can oxidise the sediment up to 10 cm deep. This leads to the sediment retaining its function as a sink for phosphate, whilst the aerobic decomposition of detritus is promoted and the formation of organic lake floor sediment is hindered, which would otherwise produce reducing conditions and promote P release from the sediment (Vahle 1990).

4.7  Vegetation Dynamics 4.7.1  Seasonal and Interannual Fluctuations The seasonal change in phytoplankton communities in the lakes of the temperate zone is one of the best understood examples of vegetation dynamics, therefore we will briefly illustrate it here. After the massive disturbance of the winter cold, there will be a succession of 30 to 100 generations of different phytoplankton species. The sequence of these generations can be seen as an annual secondary succession. In the eutrophic lakes of Central Europe, there is usually a spring maximum of small algae followed by a clear-water stage caused by the predation pressure of the zooplankton. This is then replaced by a second, midsummer maximum of larger and better defended phytoplankton forms (see Fig.  4.29). Sommer et  al. (1986) have provided a causal explanation for these dynamics in Lake Constance and other similar lakes using the Plankton Ecology Group (PEG) model. Changes in the water level are the most important cause of the strong interannual fluctuations in biomass and composition of macrophyte vegetation observed

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Fig. 4.29  Seasonal succession of major phytoplankton groups in the mesotrophic Lake Erken in southern Sweden. From Blomqvist et al. (1994, in Wetzel 2001). 1 and 2 = mid- and late-winter phase below ice, 3 = diatom blooms during the spring mixing, 4 = the beginning of the summer stagnation with low P and Si levels and blooms of small flagellates, 5 = summer clear-water phase during the stagnation with intensive grazing by zooplankton, 6 = late summer blooms of Chrysophyceae, Cryptophyceae, green algae and cyanobacteria under reduced grazing pressure, 7 = autumn mixing with diatom blooms, 8 = beginning of the winter stagnation and ice

in some shallow lakes. The water level influences the light levels and wave action, as well as the chemical composition of the water, which are all factors that have a significant impact on macrophyte growth (Neuhuber 1971, Schiemer 1979). Some shallow lakes undergo regular fluctuations every 5–20 years from clear-water phases with rich macrophyte vegetation to turbid, plankton-rich phases with few macrophytes without any changes in nutrient input (e.g. Blindow et  al. 1993; Scheffer et al. 1993). These fluctuations in macrophyte biomass are probably primarily caused by competition for light between macrophytes and phytoplankton. Brönmark and Weisner (1992) explain the two quasi-stable ecosystem states as the result of a cascade of trophic interactions: a cold winter with high mortality of predatory fish promotes planktivore fish populations. These decimate the zooplankton, thereby promoting the phytoplankton and leading to severe reductions in macrophyte vegetation. This hypothesis predicts that strong reduction in planktivore fish populations in year 1 is enough to cause an increase in grazing pressure from zooplankton and lead to clear water with low phytoplankton density, steering the system back towards macrophyte dominance. Dense macrophyte stands stabilise the clear water stage and suppress new phytoplankton blooms via several mechanisms. These include especially shading, reduction in water movement and thus increased plankton sedimentation and facilitation of plankton grazers, but also probably allelopathic effects of macrophytes (such as Myriophyllum spicatum) on the plankton (Körner and Nicklisch 2002; Hilt 2006). Changes in lake water levels presumably play a central role in the switch between the two alternative stable states (Blindow et al. 1997).

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Changes in water level and winter ice drift often lead to significant interannual changes in riparian vegetation (Kopecky 1969; Brux et al. 1988). However, all common macrophytes have developed successful adaptations to compensate for losses of biomass after catastrophic events, including turions, shoot fragments capable of rapid regeneration or long-lived rhizomes. Communities of flowing water receive new species via vegetative dispersal of helophytes from the river bank, or if hydrophytes are washed down from the upper reaches (Wiegleb et  al. 1989). Drift is an important process not only for river animals but also for macrophytes, especially as many river phanerogams do not have a permanent seed bank (Hughes and Cass 1997).

4.7.2  Long-Term Dynamics and Succession in Lakes The majority of Central European lakes are relatively young, and are dynamically developing into raised bogs or similar semi-terrestrial ecosystems. In the course of this succession, the lake becomes enriched in nutrients that collect from the inflows into the lake basin. At the same time, the lake bed builds up, moving closer to the epilimnion and its primary producers, causing the lake to undergo natural eutrophication. This process is slower in large and deep lakes than in small and shallow ones, as a larger area is available to absorb the incoming particles and nutrients. The terrestrialisation of a lake is also driven by the lake organisms themselves, as their dead biomass gradually fills the lake bed. The frequent lack of oxygen in the lake sediments hinders the productivity of aquatic plants less than the decomposition of dead plant and animal material, so that the organic matter accumulates on the lake bed (Xiong and Nilsson 1997). The silting up of lakes is thus a case of both autogenic and allogenic succession. If the transition from lake to bog begins when the area is free of forest, then inorganic sediment layers of allochthonous origin dominate, which are then covered by authochthonous organic layers (Lang 1975). Biotic activity can also contribute to the accumulation of inorganic sediments; the most important of these processes is calcite precipitation caused by the intensive photosynthesis of phytoplankton during the summer (see Sect. 4.4.1), which leads to the formation of lime gyttja and lake marl in the sediment of calcareous lakes. The clear zonation of the littoral vegetation in many lakes may lead to the impression that the spatial sequence of vegetation types is caused by a temporal change, i.e. a directional succession from floating-leaved plant communities through reedbeds to swamp forests. This is, however, usually not the case, as is shown by numerous palaeo-limnological studies of sediment cores from silted-up Central European lakes (e.g. Lang 1975; Grosse-Brauckmann 1974; 1976b). Human modification of the landscape as well as natural changes in climate and hydrology have repeatedly affected the succession in lake vegetation. Long periods of eutrophication are often

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followed by periods of oligotraphent plant communities. For example, reed peat can follow alder swamp peat, as the water level rose and killed the forest vegetation (Barth and Pott 2000).

4.7.3  S  elf-Purification of Water Bodies as Secondary Succession If organic matter (e.g. from sewage) leaches or flows into a water body, then the nutrient supply for heterotrophic aquatic organisms will improve and cause a succession to decomposing and mineralising organisms. Fresh wastewater is dominated by carbohydrate-decomposing bacteria, followed by protein-degrading bacteria and, with sufficient oxygen, finally nitrifying bacteria (Schwoerbel 1999). This type of heterotrophic succession occurs as a temporal sequence after the inflow of wastewater into a lake or pond; in flowing water, the temporal sequence is replaced by a stretch of river or stream that consists of the various stages of decomposition and acts as a self-purification stretch. During the course of the succession, the decomposers change the quality of the organic matter and the concentration of oxygen and nutrients in the water, thereby also influencing the aquatic plants. These stages are assigned to water quality classes (see Table  4.6), which are defined according to chemical parameters of the water and indicator organisms (Friedrich 1990). Figure 4.30 shows important chemical and biological parameters that characterise the decomposition of organic wastewater flowing into the Mettma, a low mountain stream in the Black Forest, over an 8  km self-purifying stretch. After the introduction of the wastewater, rapid aerobic decomposition of the organic matter begins, mainly by the bacterial species Sphaerotilus natans (‘sewage fungus’), characteristic for the overly polluted (‘polysaprobic’) stage. Intensive respiration leads to the formation of an anoxic zone close to the wastewater inflow, accompanied by the release of large amounts of NH4+ as the first product of protein decomposition. Dissimilative nitrate reduction (denitrification) in the anoxic water reduces the NO3− content of the water. After 2 or 3 km, the oxygen consumption has dropped enough for nitrification (oxidation of ammonium) to take place again (highly polluted stage, α-mesosaprobic). The high nutrient levels promote the growth of macrophytes, that also contribute via photosynthetic production of O2 to a rapid re-oxygenation of the water. However, it is rare that this process will finally achieve good water conditions (no or minimal pollution, oligosaprobic) further downstream, as a proportion of the nutrients from the wastewater will remain in the system and permanently increase the nutrient and organic matter levels. The self-purification process therefore usually ends at a moderately polluted (β-mesosaprobic) stage.

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Fig. 4.30  The process of self-purification along a 7 km stretch of the Mettma, a low mountain stream in the Black Forest, after the inflow of wastewater from a brewery. Showing concentrations of ammonium, nitrite and nitrate in the water, O2 content and O2 deficit and the biomass of the ‘sewage fungus’ Sphaerotilus natans and epilithic algae (without Sphaerotilus) (From Schwoerbel 1999)

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4.8  Human Influence 4.8.1  Eutrophication of Water Bodies Stillwater Bodies  The anthropogenic eutrophication of still and flowing water bodies, i.e. the increase in primary productivity due to increased nutrient inputs, is one of the most serious disturbances to aquatic ecosystems caused by man. Lakes and ponds have undergone natural eutrophication over centuries and millennia as they have silted up into fens (see Sect. 4.7.2), but it is only in the last 60 years or so that phosphorus inputs have caused sudden eutrophication of many Central European lakes and rivers. The submerged macrophytes of the littoral zone suffer from increased epiphyte growth and higher densities of phytoplankton in the water column under eutrophication, which both reduce the light availability (e.g. Jupp and Spence 1977). The diverse macrophyte vegetation has been replaced in many mesotrophic to eutrophic ponds and small lakes by duckweed and species-poor Ceratophyllum demersum stands, indicating a transition to a hypertrophic stage (Pott 1995). The steep increase in biomass production in eutrophic lakes leads to higher oxygen consumption in the hypolimnion and the sediment, as well as the formation of methane, and increased denitrification and desulphurisation. This frequently leads to the mass death of fish from lack of oxygen or ammonium toxicity, as well as death of birds (e.g. from botulism). Increased release of P from the sediment also enhances the eutrophication in a process known as internal loading. Eutrophication increases not only the total primary production of a lake, but also influences the relationships between the major groups of primary producers (cf. Philipps et al. 1978; Wetzel 2001). In shallow, oligotrophic lakes, nutrient inputs first promote the submerged macrophytes, but with increasing eutrophication it is the floating algae and phytoplankton that profit more than the submerged macrophytes rooted to the sediment. It is often the epiphytes that are the most productive group under mesotrophic conditions, whilst the phytoplankton reaches its maximum productivity only under eutrophic conditions. Further increases in nutrient levels reduce productivity due to the shading effects of dense phytoplankton populations. The enrichment of organic matter in the sediment promotes tall, emergent hydrophytes and helophytes, which can shade out the submerged macrophytes almost entirely in highly eutrophic lakes (see Fig. 4.31). Agricultural fertilisers, wastewater and P-containing detergents represent the main phosphorus inputs. Wagner (1976) calculated the P load in Lake Constance from these sources over the period 1930 to 1974, i.e. before water quality regulations came into force (see Fig. 4.32). As a result of the steep increase in inputs, the total dissolved phosphorus concentrations in the water increased from less than 10 μg l−1 before 1950 to around 90 μg l−1 in 1980 (see Fig. 4.33). At the same time, the phytoplankton biomass doubled and cyanobacterial blooms occurred. Microbial fixers of atmospheric N2 (e.g. Anabaena and Aphanizomenon) can serve as indicators of whether the ecosystem has changed from being P-limited to being N-limited. Lake Constance changed over this time period from an oligotrophic to a

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Fig. 4.31  Relative changes in the primary production by phytoplankton, epiphytic algae and submerged and emergent macrophytes in lakes with increasing nutrient contents (Schematic diagram based on Wetzel 2001)

Fig. 4.32  The most important factor in the rapidly increasing nutrient enrichment of Central European water bodies was phosphorus. The source of phosphorus inputs was determined for the eastern part of Lake Constance from 1930 to 1974, the period of rapid eutrophication of many Central European waters (From Wagner (1976) in Elster (1977))

­ eso-­eutrophic lake (Kümmerlin 1998), and between 1967 and 1978, the species m composition of macrophytes on its shoreline changed accordingly (Lang 1981). Successful measures to reduce water pollution and the introduction of replacements for phosphate in detergents led the P content of Lake Constance to drop below 20 μg l−1 again by 1998 (Güde et  al. 1998). This impressive ‘re-­ oligotrophication’ allowed species of the Characeae to recolonise Lake Constance. However, the phosphate levels in the sediment remained high and no clear improvement in the nitrogen content was observed (see Fig. 4.33).

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Fig. 4.33  Concentrations of total P and N, minimum O2 levels and phytoplankton biomass in Lake Constance at the end of winter mixing between 1950 and 1998 (From Güde et al. 1998). The P concentrations have since reduced again down to the levels in the 1950s

Rivers and streams are similarly affected by eutrophication. For example, the total P concentration (dissolved and particulate) in middle reaches of the Rhine near Koblenz rose to nearly 700 μg l−1 in 1970, dropping back to 300 μg l−1 by 1994. The NH4-N concentration followed a similar pattern (up to 1500 μg l−1), whilst the NO3-N concentration continued to increase (1994: 3600 μg l−1, Hamm 1996). The chloride load in the Rhine has, as in many other rivers, multiplied in the last 100 years through wastewater and potash mining (Gerken and Schirmer 1995; Tittizer and Krebs 1996).

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Table 4.15  Group indicator values of selected macrophytes for lakes and ponds in southern Germany Group indicator value 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0

Species Chara hispida, C. polyacantha, C. strigosa, Potamogeton coloratus, Utricularia ochroleuca Chara aspera, C. intermedia, Utricularia minor Chara virgata, C. tomentosa, Potamogeton alpinus Chara contraria, C. globularis, Nitella opaca, Nitellopsis obtusa, Potamogeton gramineus, P. natans, P. × zizii Chara vulgaris, Myriophyllum spicatum, Potamogeton filiformis, P. perfoliatus, Utricularia australis Myriophyllum verticillatum, Potamogeton berchtoldii, P. lucens, P. praelongus, P. pusillus Hippuris vulgaris, Lagarosiphon major, Potamogeton pectinatus Elodea canadensis, E. nutallii, Potamogeton crispus, P. obtusifolius, P. compressus, Ranunculus circinatus, R. trichophyllos Ceratophyllum demersum, Lemna minor, Potamogeton mucronatus, P. nodosus, Sagittaria sagittifolia, Spirodela polyrhiza, Zannichellia palustris

From Schneider and Melzer (2005) 1.0 very clean, 5.0 highly polluted

As a result of this eutrophication, the upper and middle reaches of the Rhine lost many of its indicators of mesotrophic conditions, such as Potamogeton gramineus, Callitriche obtusangula, Hottonia palustris, the aquatic mosses Fissidens rufulus and Pachyfissidens grandifrons, as well as species of Chara (Krause 1981; Philippi 1982). Still areas of nutrient-rich water are now dominated by Sagittaria sagittifolia, Ceratophyllum demersum and increasingly also the introduced water fern Azolla filiculoides, and in flowing water by Ranunculus fluitans and Potamogeton pectinatus (Westermann and Scharff 1987/88). Streams and small rivers are becoming choked with vegetation due to the mass development of submerged macrophytes and the establishment of dense reedbeds of Phalaris arundinacea and Glyceria maxima. According to a literature review by Punzel (1993), this development can be best checked by shading from riverbank through the plantation of floodplain tree species. Macrophytes as Bioindicators  Various authors have attempted to define aquatic plant species as biological indicators for the nutrient or organic matter levels in water bodies (e.g. Husak et al. 1989). Particularly the submerged species are highly sensitive to changes. Melzer (1994) and Schneider and Melzer (2005) have classified a variety of submerged macrophytes from Bavarian lakes into nine categories from oligotrophic (1) to eutrophic (5) (see Table 4.15). Examples of indicators for nutrient-poor, calcareous flowing water are Chara hispida and Potamogeton coloratus (Kohler 1976; Melzer 1988). The community of clear streams characterised by P. coloratus used to be very frequent in the northern Prealps and the Swiss Plateau

4.8  Human Influence

263

(Mittelland), but has become rare in the last few decades. Kohler (1976) and Trémolières et al. (1994) state that a decisive factor for this was the concentration of NH4+, whilst the NO3− content of the water can be very high in Potamogeton coloratus streams without affecting the plant. Large quantities of ammonium did, in fact, reach the streams via wastewater, mainly from settlements and livestock farms. Planting experiments by Kohler et al. (1973) and laboratory experiments by Glänzer et  al. (1977) suggest that Potamogeton coloratus and other species sustain direct damage from ammonium (or ammonia) particularly in alkaline waters, rather than being outcompeted by other species. Groenlandia densa (opposite-leaved pondweed) is less sensitive to the effects of ammonium and is found mainly in moderately polluted rivers, but dies if it is planted in more strongly polluted sections of the same river. In contrast, Callitriche spp. (water-starworts) appear to be highly resistant to river pollution, but are difficult to identify to species level and as such are poor indicators. Nevertheless, a high density of this genus points to a polluted water body, and they are often the last higher aquatic plants found in eutrophic streams (see Grube 1975). In still water, almost all floating duckweed species can be considered as indicators of eutrophication, especially Lemna gibba and L. minor (see Table 4.7). However, there are numerous other natural chemical and physical factors that affect the species composition in addition to wastewater, so that it is difficult to identify reliable indicators for water quality classes. Even the classification by Melzer (1994) from the foothills of the Alps is probably only regionally valid. For example, different classifications were developed by Klosowski (1985, 1990) and Klosowski and Tomaszewicz (1989) in Poland and Trapp (1995) in Bremen. In Lower Saxony, Weber-Oldecop (1969); Wiegleb (1978) and Steffen et  al. (2014) were only able to identify broad groupings even after detailed and comprehensive studies. Nevertheless, Weber-Oldecop (1969) and Wiegleb (1978) found a relatively good correlation between the species composition of still water communities and alkalinity, and the closely related conductivity of the water. Under otherwise similar conditions, this factor varies almost in parallel with the pollution level, since wastewater contains not only organic matter but also basic nutrient ions. Wiegleb (1978) summarised his findings in an ecological characterisation of multiple species. He identified the following communities as highly reliable indicators in Lower Saxony: –– the Lemnetum gibbae for phosphate- and mineral nitrogen-rich, and particularly ammonium-rich waters; –– the Ceratophylletum demersi for nitrate-rich waters; –– the Zannichellietum for very calcium-rich waters affected by phosphate; –– the Hottonietum and Stratiotetum for waters rich in carbon dioxide but poor in N and P and with neutral to acidic pH; –– the Potamogeton lucens community for ammonium-poor waters; –– the Juncus bulbosus-Sphagnum community for very acidic, HCO3−-free bog and open-cast mine pools.

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4  Vegetation of Freshwater Habitats

Sulphate inputs from farming and via atmospheric deposition can cause eutrophication of water bodies in that it releases phosphate from the sediment. Chemical reduction in the sediment also releases S2−, which leads to sulphide toxicity in aquatic plants such as Stratiotes aloides, Elodea nuttallii and Potamogeton pectinatus (e.g. Van Wijck et al. 1992). As long as there is sufficient iron in the substrate, then this toxic effect is weakened for Elodea but not for Stratiotes (van der Welle et al. 2007). In water bodies rich in sulphate and iron, Elodea can therefore outcompete Stratiotes and might lead to its decline.

4.8.2  Acidification Acid input with atmospheric deposition as well as from ecosystem-internal sources has caused long-lasting changes not only in many Central European forests, but also in its lakes and rivers in the last 50 years. Water bodies on carbonate- and clay-poor bedrocks are particularly sensitive to acidification, such as those in the crystalline massifs of the Black Forest, the Bavarian Forest, the Ore Mountains, the Harz and the central Alps (e.g. Bauer et al. 1988; see Fig. 4.34). The lowest pH values were measured in these waters after rapid snowmelt, as these had collected the acid deposits throughout the winter. Close to the source of the Schedebach stream in the Harz Mountains, for example, its water had a pH of below 3.0 in April 1985, which would normally be between 3.5 and 4.5, with highest values in autumn (Wasserwirtschaftsamt Göttingen, 1986). Low buffering capacities, shown by an alkalinity of below 100 μmol l−1, can also be found in many water bodies in the north German Pleistocene landscapes and large areas of northern Europe. Acid input first reduces the alkalinity of the water, and then causes a drop in pH, which can fall e.g. over granite to below 4.5. Carbonate-rich water bodies are largely unaffected by this process. Most water bodies in Switzerland are therefore immune to acidification through acid rain or snow, although the alpine reservoirs with crystalline bedrock are affected. Other limestone regions of Central Europe are also sufficiently buffered. According to Arts (1990), submerged moss species are good indicators of pH, as they tend to live in either acid or alkaline conditions, e.g. Sphagnum cuspidatum Drepanocladus fluitans Sphagnum denticulatum Riccia fluitans Fontinalis antipyretica Calliergonella cuspidata

3.6–4.0–5.2 3.6–4.1–5.8 3.6–4.1–5.8 5.9–7.2–9.7 6.0–7.6–8.5 7.4–8.2–8.3

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265

Fig. 4.34  Acidification of streams in western Germany during snowmelt in 1983, before SO2 emissions were greatly reduced (Modified from Schoen et al. (1984) in Arndt and Kohler (1984)). The input from atmospheric sulphuric and nitric acid was greatest close to the large industrial areas, and occurred in waves as the collected deposits in the snow were released

Decreasing pH is accompanied by increasing mobility of Al3+ ions, which are not only toxic for some plant and animal species but can also cause the precipitation of phosphate and humic substances (Baker and Christensen 1991). Acidification of lakes is therefore linked to the oligotrophication of the open water (Lampert and Sommer 1999; Dokulil et al. 2001).

266

4  Vegetation of Freshwater Habitats

In acidic lakes, Krause (1997) found that macrophytes such as Chara delicatula and Nitella flexilis disappear with a reduction in pH from 6.5 to 5.0 (cf. also Kohler and Tremp 1996). Lakes that have always been acidic, such as the Großer Arbersee in the Bavarian Forest, also experienced an increase in indicators of acid conditions, particularly those with EIV-R values of 2 and 3 (Melzer et al. 1985). Particularly Sphagnum and Juncus bulbosus have profited from acidification. According to Grahn (1977), Isoëtes species die off in acidified lakes as they increasingly become covered in epiphytic algae and cyanobacterial mats that starve them of light. The algae, in contrast, profit from the reduced herbivory in acidic water (Lazarek 1985). Lack of magnesium, enhanced by the high cation exchange capacity of Sphagnum, might also play a role in causing the yellowing of the Isoëtes rosettes (Grahn et al. 1974). Decreased acid deposition since the 1980s has resulted in a slow rise in pH values and in the alkalinity of the water in small streams and ponds of mountain ridges on acid bedrock in Central Europe. For example, the pH in the small stream Lange Bramke in the Harz Mountains increased from around 4.3 to 5.0 between 1984 and 2011 (Bittersohl et  al. 2014). This has resulted in a slow recolonisation by acid-­ sensitive biota. However, much of the sulphate remains in the soil and river and lake bed sediments and is only slowly released, so that de-acidification will take a very long time (Raspe et al. 1998).

4.8.3  Reedbed Dieback The decline of Phragmites in many Central European lakes during the last 60 years or so has been the effect of both natural and anthropogenic influences (Ostendorp 1989, Dokulil et al. 2001). For example, the reedbeds on the German shoreline of Lake Constance reduced in size by 22 % between 1954 and 1978, but since then have been increasing in extent again (Ostendorp 1991). In the Bavarian Chiemsee, this decline was around 40 % between 1960 and 1990, and even 90 % around the Ammersee (Kubin and Melzer 1997). Along the Havel in western Berlin, the reedbed reduced by 68 % between 1962 and 1982 (Sukopp and Markstein 1989). According to Ostendorp (1990, 1991), the reedbeds in Lake Constance were primarily affected by hydrological and meteorological extreme events such as high water levels in the summer, flooding of the young shoots, hail storms, and smothering by algal mats. Ostendorp (1993) also discusses the effect of increased herbivory damage to the aerenchyma of the rhizomes by the larvae of leaf beetles (Donacia clavipes), although it is unclear whether the herbivory was only a secondary effect following earlier damage to the reeds. In other areas, such as Berlin, increased wave action due to motor boat traffic, the formation of eddies by embankments, swimming and the deposition of flotsam and algal mats were considered as causes (e.g. Sukopp and Markstein 1989). Eutrophication promotes the formation of algal mats, which can bend and break young shoots, cutting off the oxygen transport to the rhizomes so that they die.

4.8  Human Influence

267

Klötzli (1971) and others also blame the reduced formation of sclerenchyma rings in the stem due to higher nutrient conditions, so that some reedbeds are more sensitive to mechanical damage than they used to be. Palaeolimnological studies indeed show for southern Germany that a reduction in reedbeds coincides with eutrophication (Kubin and Melzer 1997). Bornkamm and Raghi-Atri (1986) state that nitrogen has a particularly large effect on some reed ecotypes, and that phosphate is less important in this respect. Klötzli and Züst (1973) found large amounts of NH4+ in collapsing Phragmites stands on damp soil (see also Grünig 1980 and Bogenrieder 1990). The once common practice of reed cutting in winter, which removed the dead stems and caused dense and robust stands to form, also stopped several years before this (Leippert 1978). Armstrong et al. (1996a, b) consider the main cause to be the callus formation caused by the exposure to toxins, causing blockage of the gas and vascular transport in the reed stem. Doubtless many, often locally acting factors played a role in the reduction in the extent of reedbeds, but eutrophication probably increased the susceptibility of Phragmites to these stressors (cf. van der Putten 1997). In the expanding reedbeds of Lake Neusiedl, Burian and Sieghardt (1979) found numerous gaps within the stands, which resembled the gap dynamics of temperate old-growth forests caused by natural senescence. It should also be mentioned that Phragmites spreads in many wet meadows and marsh complexes if these are not regularly mown, causing problems for nature conservation (e.g. Güsewell et  al. 2000).

4.8.4  Threats to and Conservation of Freshwater Habitats The widespread eutrophication and stream basin modification are the greatest threats to the Central European freshwater vegetation. Both acid and calcareous oligotrophic lakes and ponds and their characteristic vegetation have become rare across Central Europe. All of the communities of the Littorelletalia are now highly threatened (Dierßen 1983). During the last few decades, the number of lakes in northern Germany and the Netherlands dominated by Littorellion species has seen a steep decline. Particularly Lobelia dortmanna, Isoëtes lacustris and Littorella uniflora have undergone large population decreases. Roelofs (1983) studied 68 water bodies dominated by Littorella in 1950. By 1980, the species was very rare or absent from 53 of these. 41 had become dominated by submerged Juncus bulbosus or Sphagnum species, mainly due to the acidification of the water. In the remaining 12, the plant communities had changed, partly due to the eutrophication of the water and the sediment, mainly by phosphate inputs. They had in 1980 large populations of e.g. Myriophyllum alternifolium and Ranunculus peltatus, which, in contrast to Isoëtes, profit from higher bicarbonate concentrations (Roelofs et al. 2005). In small, shallow lakes, floating plants such as Riccia fluitans and Lemna minor dominated, shading out the Littorellion species. In deeper lakes, plankton and epiphyte densities increased, reducing the growth

4  Vegetation of Freshwater Habitats

268

Table 4.16  Floristic changes in various streams and rivers in four different regions of Lower Saxony between the 1950s and 2010/2011 based on repeated surveys of 338 semi-permanent plots according to Steffen et al. (2013)

Decreased species Callitriche hamulata Callitriche palustris agg. Ceratophyllum demersum Elodea canadensis B Fontinalis antipyretica Juncus bulbosus fluitans Lemna trisulca B Leptodictyum riparium Myriophyllum alterniflorum Nuphar lutea Potamogeton alpinus Potamogeton crispus Potamogeton friesii Potamogeton lucens Potamogeton natans Potamogeton obtusifolius Potamogeton pectinatus Potamogeton perfoliatus Potamogeton pusillus Ranunculus aquatilis agg. Increased species Elodea nuttallii Lemna gibba Lemna minor Myriophyllum spicatum Spirodela polyrhiza Species without overall decrease or increase Sparganium emersum B Chiloscyphus polyanthos

Constancy in the 1950s

Constancy in 2010/2011

Significance of change

23.1 41.4 5.9 39.9 6.8 2.7 15.1 5.9 9.5 34.6 5.3 18.3 9.5 10.9 30.6 1.5 16.0 12.4 6.8 23.4

12.1 32.0 7.1 20.4 3.3 0 9.5 0 4.4 24.0 0 7.4 0 3.0 10.4 0 10.4 3.0 3.6 8.0

*** *** n.s.a *** * n.s.a ** *** *** *** *** *** *** *** *** * ** ** ** ***

0 7.1 34.0 0.9 10.7

19.8 9.8 61.2 5.6 32.5

*** n.s.a n.s.a *** n.s.a

52.1 1.1

52.1 1.4

n.s. n.s.

The historical data were collected between 1936 and 1969 (summarised as ‘1950s’) by Roll (1939), Weber-Oldecop (1969) and R. Alpers (unpubl. material in the R.-Tüxen-Archive in Hannover). Percent frequency (constancy) values are given (1950s vs. 2010/2011); *: P 7, cf. Lötschert and Georg 1980). Sandy grasslands in continental eastern Central Europe usually have higher pH values than those in the oceanic west due to the climatic water balance. Glacial sands here can also reach pH (KCl) values of 6–7 with high base saturation (Kozlowska and Wierzchowska 1985). This is also the case for the dry Corynephorus grasslands of the east, particularly if these grow on freshly deposited substrate. The carbon and nitrogen contents of dune and aeolian sands below dry grasslands are very low, probably as a result of the lack of phosphorus in these highly acidic substrates. Lache (1976), Storm et  al. (1998) and Jentsch and Beyschlag (2003) recorded only 0.1–0.6 % C and less than 0.04 % N in the Ah horizon of northwest German inland dunes, i.e. six to ten times lower than in nutrient-rich meadows. The C and N contents in the topsoil of the dry calcareous grasslands, however, are usually similar to those in neighbouring forests. This may at first glance be surprising, as high summer temperatures generally accelerate litter decomposition. However, summer drought decreases the activity of the mineralising soil organisms, so that some dry grasslands have very high C and N pools, such as in the Chernozem steppe grasslands of eastern Central Europe. On very steep slopes, the humus-rich topsoil can be lost through erosion, as suggested by the low humus contents measured by Quantin (1935) in a sloping Anthylli-Teucrietum in the French Jura (see Table 7.1). Dry grasslands thus develop on a wide range of soils, which share the common characteristic that the biological activity in the soil is slowed for several weeks or months by drought. This climatic drought is compounded by the low or very low capacity for plant-available water in most dry grassland habitats, such as in the case of the rocky soils only a few centimetres thick as well as the deep sandy soils with large proportion of coarse sand that reduces the available water.

7.3  Vegetation 7.3.1  Classification of the Major Habitat Types In order to give an overview of the variety of vegetation units within dry grasslands, we will first look at the major habitat factors without going into detail. The Ellenberg indicator values (EIV) of the numerous character species can be helpful in this respect, as is shown in the ecograms in Fig. 7.13 (see the explanation of the EIV

Fig. 7.13  (continued) grasslands are light-demanding (EIV-L 7–9) and adapted to extremely low nutrient levels (N < 3). See also Table 7.5. It should be emphasised that the alliances are placed in the ecograms according to their character species (from Ellenberg et al. 1992), and not to the average values for the whole stand. These show similar tendencies but less clear differences. The largest amplitudes in T and C values are seen in the Mesobromion character species on calcareous soils, and in those of the Sedo-Scleranthion on silicate bedrock, and of the Corynephorion on sandy soils. All alliances have extreme M and R values, but particularly the Festucion valesiacae and the Xerobromion. Further details are given in the text

7.3 Vegetation

515

Fig. 7.13  Ecograms of character species of major Central European dry grassland alliances in the classes Festuco-Brometea and Koelerio-Corynephoretea (see Table 7.4), according to temperature and continentality values or moisture and reaction values. All character species of dry

516

7  Nutrient-Poor Dry Grasslands

scores in the Directions for use in this volume). This portrayal is somewhat unusual, as instead of showing the average indicator values it instead gives the ranges of the relevant character species (see Table 7.4). Only few studies analysed the relationships between measured abiotic site factors and dry grassland vegetation types. One example is the study of Jandt (1999) in a variety of xerothermic habitats in northern Thuringia. The over 100 dry grassland communities so far identified for Central Europe can  be divided roughly into two classes, the Festuco-Brometea (calcareous dry  grasslands) and Koelerio-Corynephoretea (sandy and rocky dry grasslands). The diversity of the Central European dry grassland communities means that although the two classes are floristically easy to distinguish, the lines are somewhat blurred regarding soil chemistry and vegetation structure. As we will see, sandy or rocky grasslands can also occur on carbonate substrates, and may produce dense swards. In turn, some Festuco-Brometea grasslands can be found on acidic soils or have particularly sparse vegetation. Nevertheless, the majority of communities in this class are found on calcareous soils, whether from carbonate rock, loess at an early stage of weathering, till or other substrates. The pH indicator values of their character species are therefore generally high (R 6–9), except for in one rather marginal alliance (Koelerio-Phleion phleoidis, see Fig. 7.13, left; EIV-R 3.5–5.5). The whole class is relatively thermophilic (T 6–8). The class Koelerio-Corynephoretea of sandy and rocky grasslands (previously called the Sedo-Scleranthetea) is in some respects quite a contrast to the Festuco-Brometea. Two groups of communities can be clearly distinguished within this class in terms of their flora and habitat. The order Sedo-Scleranthetalia colonises shallow Protorankers or Protorendzinas (Lithic Leptosols) on both silicate and carbonate solid bedrock in rocky or cliff habitats, whilst numerous communities within four other orders (see the overview of communities in Chap. 14: no. 5.2) grow on sandy and usually deep unconsolidated substrates. The majority of the communities in the class Koelerio-Corynephoretea grow on weakly to highly acidic soils (EIV-R 6 to R 2), although some are found on carbonate rocks and basic sands (the alliances Alysso-Sedion: R 6–9, Koelerion albescentis: R 3–7, Koelerion glaucae: R 7–8 and Seslerio-Festucion pallentis: R 5–8; see Table 7.4). With the exception of the Koelerion glaucae (continentality values: C 6–8) most communities in the Koelerio-Corynephoretea have a generally suboceanic character (C 1–C 6) (Table 7.4, 7.4a and 7.4b). The two dry grassland classes are also similar in that their character species of class, orders and also alliances are reasonably drought-tolerant (M 2–3 or 4) and limited to moderately to very warm habitats (usually T 5–8 or 9). It is only the communities of the Sedo-Scleranthion (see Table  7.4b: bottom section) that also occur at high elevation, where they are dominated by succulent species. In the following, we will discuss the major dry grassland communities of both classes in more detail, although the multitude of described associations means that we can only give a general overview. We will focus more on the large-scale relation-

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Table 7.4  Character species of calcareous and sandy dry grasslands, particularly in the suboceanic regions of Central Europe, showing the ecological characteristics of the species

Classes: Festuco-Brom. (FB) Koel.-Cor. (KC) FB KC

On lime-rich (calcareous) soils Calcareous dry grasslands FestucoBrometea LTC MRN Ajuga genevensis 8xx 372

On non-calcareous soils Sandy dry grasslands Koelerio-­ Corynephoretea LTC MRN Acinos arvensis 963 251

Allium carinatum

854 382

g Agrostis stricta

973 221

Allium oleraceum

764 374

Androsace septentrionalis Artemisia campestris B Brachythecium albicans → B Ceratodon purpureus Helichrysum arenarium Herniaria glabra

877 252

Hieracium echioides Holosteum umbellatum Lactuca perennis →

866 261 865 3x2

865 3x2 874 381 877 352 917 21–

L 7–9 T 5–8

7–9 5–8

Allium sphaerocephalon Asperula cynanchica

985 382 7x5 383

C 3–7

2–7

Aster linosyris

875 282

M 2–4

2–4

976 383

R 7–9

1–6

→

g Bothriochloa ischaemum g Brachypodium pinnatum Campanula glomerata g Carex humilis

837 362

g = graminoid l = legume B = bryophyte o = orchid

Centaurea jacea subsp. subjacea Centaurea scabiosa Eryngium campestre Euphorbia cyparissias g Festuca rupicola Filipendula vulgaris Galium verum Gentiana cruciata Minuartia setacea Odontites lutea Ornithogalum kochii

765 382 76x 473 764 383 976 271 775 392 985 281

Orobanche caryophyllacea g Phleum phleoides Pimpinella saxifraga g Poa pratensis subsp. angustifolia Polygala comosa

865 392

Minuartia viscosa g Poa badensis → g Poa bulbosa B Polytrichum piliferum Potentilla argentea Potentilla collina Potentilla inclinata Potentilla recta Potentilla rhenana B Racomitrium canescens → Rumex tenuifolius

867 382 7x5 3x2 76x xx3

Scleranthus perennis Sedum acre Sedum forsteranum

864 241 863 2x1 871 341

866 382

Sedum rubens

772 3x3 (continued)

N 1–4 1–3 = Species with partly deviating ecological behaviour

655 474 7x7 47x 765 283

7x3 384 975 383 8x4 3x3 977 382

965 252 935 2x– 8xx 2x– 867 251 865 342

974 282

963 231 974 221 975 261 975 352 984 221 936 16– 965 321

7  Nutrient-Poor Dry Grasslands

518 Table 7.4 (continued)

Classes:

On lime-rich (calcareous) soils Calcareous dry grasslands FestucoBrometea LTC MRN Potentilla heptaphylla 754 392 B Rhytidium rugosum 9x6 37 – Salvia pratensis Sanguisorba minor

864 384 765 382

Scorzonera austriaca Stachys recta B Thuidium abietinum

777 382 764 392 8x6 27–

Thymus praecox

856 381

l Trifolium montanum Veronica austriaca Veronica spicata

8x4 382 866 392 776 372

On non-calcareous soils Sandy dry grasslands Koelerio-­ Corynephoretea Sedum sexangulare Sempervivum tectorum B Tortula ruralis Taraxacum laevigatum → l Trifolium arvense l Trifolium campestre Valerianella carinata → Valerianella dentata → Veronica verna

LTC MRN

754 261 8x2 34x 9x5 26− 865 372 863 321 863 463 773 48x 762 47x 875 241

From Oberdorfer et al. and Ellenberg et al. up to 1992. Further information is given in the text, and the table continues on the following pages Table 7.4a  Character species of calcareous and sandy dry grasslands

Orders: Bromet. er. L 7–9 T 5–8 C 2–4

Suboceanic calcar. dry grasslands Brometalia erecti Anthyllis vulneraria

LTC MRN

863 372

Arabis hirsuta g Avenochloa pratensis g Bromus erectus

753 48x 764 3x2 852 383

M 1–4 R 7–9

g Carex caryophyllea Dianthus carthusianorum

8x3 4x2 854 372

N 1–3

Globularia punctata →

865 292

Festuc. vales. L 7–9

Helianthemum ovatum

854 392

Hieracium wiesbaurianum

763 381

T 6–7 C 5–8 M 1–3

Hippocrepis comosa Iris germanica g Koeleria pyramidata

752 372 883 382 764 472

Continental calcar. dry grasslands Festucetalia valesiacae Achillea collina Achillea setacea Allium flavum l Astragalus onobrychis Centaurea stoebe Dianthus gratianopol. → Euphorbia seguierana Hieracium bauhinii → Orobanche coerulescens l Oxytropis pilosa Pimpinella nigra Potentilla arenaria

LTC MRN

966 272 978 271 –– 976 291 976 282 974 271 976 281 974 371 977 281 977 171 966 281 976 181 (continued)

7.3 Vegetation

519

Table 7.4a (continued)

Orders: R 7–9 N 1–2

Sedo-­ Scler. L 6–9 T 6–8 C 2–5 M 2–4 R 3–6 N 1–4

Suboceanic calcar. dry grasslands Brometalia erecti Linum leonii →

LTC MRN

975 391

Linum viscosum Ononis natrix

7x4 48? 883 381

Pulsatilla vulgaris →

765 272

Scabiosa columbaria

852 382

Teucrium montanum l Trifolium ochroleucon

854 191 774 482

Silicate rock debris and rock ledge communities Sedo-Scleranthetalia Allium montanum Arabidopsis thaliana Arenaria leptoclados g Festuca pannonica → Jasione montana Petrorhagia saxifraga → Sedum album Teucrium botrys → l Trifolium ornithopodioides

Continental calcar. dry grasslands Festucetalia valesiacae Potentilla pusilla → Pulsatilla grandis Pulsatilla montana → Ranunculus illyricus → Scorzonera purpurea Silene otites g Stipa joannis Verbascum phoeniceum Veronica prostrata

LTC MRN

964 281 96? 391 845 281 876 474 876 282 877 272 878 272 766 372 876 281

LTC MRN

9x5 262 663 444 983 3x2 976 2x1 763 332 973 271 9x2 2x1 964 282 962 3x?

L light, T temperature, C continentality, M moisture, R acidity/alkalinity, N nitrogen, x indifferent, − unknown Table 7.4b  Character species of calcareous and sandy dry grasslands On lime-rich (calcareous) soils Suboceanic calcar. Alliances: dry grasslands Xerobrom. Xerobromion 976 281 764 291

Continental calcar. dry grasslands Festucion valesiacae g Carex supina Erysimum odoratum

973 291 865 292 885 261

g Festuca duvalii 987 181 g Festuca valesiaca 877 272 Hieracium rothianum 986 281

LTC MRN

L 7–9 T 6–8 C 2–5 M 1–2 R 7–9

Allium pulchellum → Dorycnium germanicum Fumana procumbens Globularia punctata Gypsophila fastigiata →

LTC MRN

777 272 976 282

Alliances: Festucion v. L 7–9 T 7–8 C 6–8 M 1–2 R 7–9 (continued)

7  Nutrient-Poor Dry Grasslands

520 Table 7.4b (continued) On lime-rich (calcareous) soils Suboceanic calcar. Alliances: dry grasslands N 1–2 Helianthemum 872 281 apenninum Helianthemum canum 874 291

Continental calcar. dry grasslands Nepeta pannonica

876 27?

Alliances: N 1–2

986 291

Helianthemum nummularium g Koeleria vallesiana Linum tenuifolium Orobanche amethystea → Orobanche teucrii Phyteuma tenerum g Stipa bavarica Thymus froelichianus Trinia glauca Alliances: Continental calcar. semi-dry grasslands Mesobrom. Mesobromion

764 371

Seseli hippomarathrum g Stipa capillata

985 291 984 292 785 3x1

g Stipa pulcherrima 987 181 Thymus oenipontanus – – (in Tyrol endemic)

L 7–9

772 483

Suboceanic calcar. semi-dry grasslands CirsioBrachypodion Adonis vernalis

872 392

l Astragalus danicus

877 392

L 7–9

754 473

Danthonia alpina

96? 3?2

T 6–7

9x4 4x2

Inula spiraeifolia

876 392

C 5–8

753 473 954 382 877 481

Linum austriacum Linum flavum Linum perenne

976 382 876 483 7x6 382

M 2–4 R 6–9 N 1–3

957 482 862 393 834 492 7x4 382 754 483

l Onobrychis arenaria l Ononis arvensis Scabiosa ochroleuca Senecio integrifolius Seseli annuum →

777 291 867 472 876 382 767 48? 875 392

757 582

g Stipa tirsa

878 362

765 382

Thesium linophyllon →

875 281

o Aceras anthropophorum T 4–7 o Anacamptis pyramidalis C 2–6 g Brachypodium rupestre M 3–4 Carlina acaulis subsp. simplex R 7–9 Carlina vulgaris N 2–3 Cirsium acaule Koelerio-P. Equisetum ramosissimum → L 7–9 Erigeron acris → T 6–8 Euphorbia verrucosa C 2–6 Gentiana aspera → M 2–3 Gentianella ciliata R 4–6 Gentianella germanica N 1–2 o Herminium monorchis → Hieracium cymosum o Himantoglossum hircinum

878 282

862 291 864 281 96? 171 874 281 985 181 LTC MRN

Alliances: LTC MRN

Cirsio-­ Brach. 767 372

772 392 (continued)

521

7.3 Vegetation Table 7.4b (continued) On lime-rich (calcareous) soils Suboceanic calcar. Alliances: dry grasslands l Medicago lupulina

75x 48x

Nonea pulla l Onobrychis viciifolia l Ononis repens l Ononis spinosa

766 392 876 383 852 472 865 473

o Ophrys apifera

762 492

o Ophrys holosericea 874 492 o Ophrys insectifera 754 493

o Ophrys sphecodes o Orchis militaris o Orchis morio o Orchis simia o Orchis tridentata o Orchis ustulata Polygala chamaebuxus → Prunella laciniata Ranunculus bulbosus o Spiranthes spiralis → On non-calcareous soils Silicate rock debris and rock ledge communities Sedo-Veronicion Alliances: dillenii Sedo-Vero. Androsace elongata → L 8–9 Gagea bohemica

884 493 765 392 753 473 882 382 974 392 755 4x3 644 383

T 6–9

M2

Scleranthus polycarpos → Scleranthus verticillatus Spergula pentandra

R 3–6

Tuberaria guttata

972 251

N 1–2

Veronica dillenii

876 252

C 2–6

Continental calcar. dry grasslands Acid semi-dry grasslands Koelerio−Phleion phleoides [in order Brometalia erecti] Armeria elongata Armeria arenaria g Festuca cinerea g Festuca trachyphylla g Koeleria macrantha → Lychnis viscaria etc. (not yet clear, especially for the oceanic communities)

Alliances:

763 362 882 362 974 2x1 866 3x2 767 372 764 342

772 392 863 373 862 552

Alpine succulent communities on silicate rock Sedo-Scleranthion

966 462

Alliances: SedoScle. L 8–9

943 231

Cerastium arvense subsp. strictum Plantago serpentina → Sedum annuum

933 341

T 2–6

974 261

Sedum vulgare

942 342

C 2–6

964 261

Sempervivum arachnoideum Sempervivum montanum Silene rupestris

934 221

M 2–3

822 321

R 2–6

932 331

N 1–2

LTC MRN

987 261 985 252

LTC MRN

936 3x2

522

7  Nutrient-Poor Dry Grasslands

ships rather than specific regional types. Many phytosociologists have studied Central European dry grasslands from a regional perspective, leading to the creation of conflicting classifications (cf. Dierschke 1986a, 1997; Chytrý 2007). Here, we mainly follow Rennwald (2000), as well as Korneck (1993), Pott (1995) and – for the Alps – Mucina and Kolbek (1993a, b), who give more of an international perspective, even if they still come to contradictory conclusions in some cases.

7.3.2  Calcareous Dry Grasslands (Class Festuco-Brometea) The class Festuco-Brometea (Chap. 14: no. 5.3) contains grasslands with topsoil of pH 6.0–7.5, and rarely any lower. This class differs from the Koelerio-­ Corynephoretea (acid sandy and rocky grasslands) in its unusually large number of character species, and the two are linked by only a few intermediate community types (see Table 7.4). As implied in the introduction to this section, this is mainly due to the fact that the chemically intermediate soils, i.e. the weakly calcareous to weakly acidic, are mainly deeper than the limestone or dolomite Rendzinas on the one hand and the silicate Rankers on the other. In addition, deep soils generally do not support dry grasslands, but rather arable fields, fertilised pastures or forest. The chemical differences between calcareous and silicate soils have thus been extended by centuries of human use and the resulting soil erosion. 7.3.2.1  Submediterranean Calcareous Dry Grasslands The calcareous dry grasslands can be divided into those with more continental, and those with more oceanic or submediterranean species. At lower elevations, the climatic and floristic contrasts are more pronounced than higher up in the mountains. The overview in Table 7.5 is therefore limited to colline areas, but uses examples from across Central Europe to show how the extremes are linked by intermediate vegetation types. This is somewhat similar to the gradient of thermophilic broadleaved forests in Central Europe (see Table 6.3 in Vol. I), but even clearer in the case of dry grasslands. Only older vegetation records were used to avoid the distorting effect of eutrophication. The suboceanic-submediterranean calcareous grasslands are placed within the order Brometalia erecti (see Table 7.4). Their character species typically have low continentality values (EIV-C 2–5, or 6 at most). These communities are best developed in limestone mountain areas in the submediterranean region with warm summers and mild winters. In Central Europe, these communities are found in particularly species-rich forms in western Switzerland, southwestern Germany and in the Alsace region. There are only relatively species-poor forms in the Netherlands, the limestone region of the Eifel and in the Lower Saxony uplands, and those in Denmark and southwestern Sweden have little of the character of the Brometalia erecti left (Dierßen 1996a, Bruun and Ejrnaes 2000).

7.3 Vegetation

523

Table 7.5  Dry grasslands on base-rich bedrock in various areas of Central Europe. From tables in Braun-Blanquet, Preis, Oberdorfer, Quantin, Bornkamm and othersa Order: Alliance: Association no.: C Carex liparocarpos F Potentilla pusilla P Poa alpina carniolica P Koeleria vallesiana F Pulsatilla montana S Petrorhagia saxifraga P Asperula cynanchica Scabiosa gramuntia K Scorzonera austriaca F Festuca valesiaca Elymus hispidus C Veronica praecox C Astragalus excapus F Achillea setacea F Centaurea stoebe S Myosotis stricta C Thymus glabrescens F Achillea collina F Pulsat. pratensis nigra F Stipa pennata & joannis Tortula ruralis S Poa bulbosa F Oxytropis pilosa S Sedum rupestre Melica transsilvanica F Carex supina K Veronica spicata S Erysimum crepidifolium B Festuca cinerea Sesleria albicans Buphthalmum salicifol. F Erysimum odoratum K Thymus praecox K Festuca rupicola F Thesium linophyllon F Stipa capillata K Allium sphaerocephalon F Potentilla arenaria Galium glaucum

Festucetalia valesiacae Stipo- Festucion Poion valesiacae 1 2 3 4 5 5 4 4 4 4 3 3 3 5 5 1 3 1 3 • 3 1 2 5 5 4 4 3 3 2 3 2 3 4 1 3 1 1 2 1 2 1 1 3 1 • 3 2 2 1 1 1 4 5 2 2 1 2 5 3 2 1 5 1 1 3 • 2 5 4 5 4 2 3 5 5 3 4 5 3

Brometalia erecti XeroMeso­bromion bromion 5 6 7 8 9

5 5 3 2 3 1

3 3 3 1 (continued)

7  Nutrient-Poor Dry Grasslands

524 Table 7.5 (continued) Order: Alliance: Association no.: S Alyssum montanum F Seseli hippomaratrum F Stipa pulcherrima S Artemisia campestris K Stachys recta B Koeleria macrantha Medicago minima K Bothriochloa ischaemum F Silene otites C Globularia punctata K Lactuca perennis S Melica ciliata Teucrium chamaedrys K Aster linosyris K Carex humilis K Phleum phleoides K Asperula cynanchica K Odontites lutea Peucedanum oreoselinum F Euphorbia seguierana K Euphorbia cyparissias S Arenaria leptoclados S Acinos arvensis S Erophila verna Hieracium pilosella K Centaurea scabiosa B Hippocrepis comosa B Helianthemum ovatum B Bromus erectus K Brachypodium pinnatum K Sanguisorba minor B Anthyllis vulneraria B Linum tenuifolium S Alyssum alyssoides K Eryngium campestre B Dianthus carthusianorum K Pimpinella saxifraga S Sedum album S Sedum acre Medicago falcata Scabiosa canescens

Festucetalia valesiacae Stipo- Festucion Poion valesiacae 1 2 3 4 1 5 1 3 1 1 2 5 4 5 3 2 3 4 3 1 5 3 2 2 2 1 2 2 4 3 1 2 3 1 1 2 1 3 3 5 4 3 4 5 2 3 4 1 5 3 3 4 4 1 2 5 5 4 1 1 4 2 4 1 5 5 5 2 4 5 2 1 2 1 2 1 5 1 1 1 1 3 3 3 1 2 2 5 5 3 • 4 • • 1 1 • 2 1 1 • 2 1 2 1 1 2 2 5 5 5 5 1 • 2 2 4 2 2 5 2 2 1 2 1

Brometalia erecti XeroMeso­bromion bromion 5 6 7 8 9 1 2 1 2 2 3 4 1 2 2 3 2 3 2 3 1 3 1 2 5 4 1 2 3 4 5 1 1 5 5 1 5 5 5 4 3 4 3 4 4 4 1 1 3 1 1 4 5 4 5 1 3 2 1 4 1 3 1 3 5 4 4 4 1 2 5 2 3 3 1 5 1 • • 4 5 1 • 5 5 5 3 3 • 3 5 5 3 1 • 3 5 4 1 3 5 3 5 3 4 1 1 3 1 3 2 3 1 2 1 3 4 4 2 4 5 3 1 1 2 1 1 3 1 1 1 (continued)

7.3 Vegetation

525

Table 7.5 (continued) Order:

Festucetalia valesiacae Stipo- Festucion Poion valesiacae 1 2 3 4 1 • 3 1 1 1 1 1 3 5 4 5 3 1 5 1 5 3 5 1 5 4 2

Alliance: Association no.: K Salvia pratensis B Avenochloa pratensis C Cerastium pumilum K Galium verum Thymus pulegioides B Festuca trachyphylla B Pulsatilla vulgaris K Potentilla verna B Teucrium montanum B Helianthemum nummul. B Fumana procumbens Anthericum ramosum B Scabiosa columbaria B Arabis hirsuta K Prunella grandiflora B Koeleria pyramidata M Ranunculus bulbosus M Cirsium acaule M Carlina vulgaris M Ononis spinosa B Carex caryophyllea M Medicago lupulina C Trinia glauca C Onobrychis arenaria D Cladonia furcata L D C. convoluta L C Ononis natrix S Petrorhagia prolifera 2 C Coronilla minima C Micropus erectus S Cerastium brachypetalum Linum catharticum Trifolium pratense D Briza media D Plantago media M Onobrychis viciifolia M Anacamptis pyramidalis D Potentilla heptaphylla D Silene nutans D Anthoxanthum odoratum D Polygala amara D Leucanthemum vulgare

1

Brometalia erecti XeroMeso­bromion bromion 5 6 7 8 9 4 5 3 4 1 5 3 1 4 1 4 1 3 1 1 2 4 4 3 4 5 5 2 5 1 4 3 1 4 1 3 4 1 1 2 5 5 5 1 1 3 1 1 3 5 4 1 2 1 5 3 2 1 3 1 3 1 2 5 3 3 1 4 2 1 1 2 2 2 2 • • 5 4 • 3 5 4 5 4 4 2 4 4 4 3 3 3 4 5 4 4 1 3 3 5 4 2 5 4 1 5 1 3 5 5 5 4 4 1 (continued)

7  Nutrient-Poor Dry Grasslands

526 Table 7.5 (continued) Order: Alliance: Association no.: D Lotus corniculatus D Plantago lanceolata D Campanula rotundifolia D Daucus carota D Leontodon hispidus D Dactylis glomerata D Centaurea jacea D Carex flacca D Knautia arvensis K Poa angustifolia D Achillea millefolium D Agrimonia eupatoria M Gentianella ciliata Several bryophytes: K Rhytidium rugosum K Thuidium abietinum K Homalothecium lutescens K Pleurochaete squarrosa

Festucetalia valesiacae Stipo- Festucion Poion valesiacae 1 2 3 4

•

4

5 4 2

Brometalia erecti XeroMeso­bromion bromion 5 6 7 8 9 5 5 5 4 5 3 5 3 4 3 5 2 4 2 2 3 2 3 4 3 3 2 3 2 2

2 1

5 5 3

1 1

4

Several less constant species were omitted, including the orchids Aceras anthropophorum, Himantoglossum hircinum, Ophrys apifera, holosericea und sphecodes, Orchis militaris, morio, simia and ustulata that occur occasionally in no. 8 and 9. The numbers after the plant names refer to constancy classes. Records prior to 1960 were intentionally used here, as the dry grasslands were largely still in active use then and not affected by eutrophication or abandonment C character species of the community in which it occurs at high constancy, D differential species of a community or an alliance with few character species, P character species of the alliance Stipo-­ Poion xerophilae, F character species of the Festucetalia or Festucion valesiacae, B character species of the Brometalia, M character species of the Mesobromion or differential species separating it from the Xerobromion, K class character species of the Festuco-brometea, S character species of the class (or lower ranking unit) of the Koelerio-Corynephoretea, which in some cases also occur in K No. 1: Inner Alpine Stipo-Koelerietum vallesianum in the lower Valais. From Braun-Blanquet (1961). Stipo-Poion = Stipo-Poion carniolicae No. 2 to 4: Examples of ± continental Festucion valesiacae: 2: Festuco valesiacae-Erysimetum crepidifolii in the low mountains in western Czech Republic. From Preis (1939) 3: Erysimo-Stipetum in the Nahe valley and in Rhenish Hesse. From Oberdorfer (1957) 4: Seslerio-Festucetum rupicolae in the Franconian Jura. From Gauckler (1938) in Oberdorfer (1957). With similarities to the Brometalia! No. 5 to 7: Examples of ± suboceanic Xerobromion: 5: Trinio-Caricetum humilis near Würzburg. From Volk (1937) in Oberdorfer (1957). With similarities to the Festucetalia! 6: ‘Xerobrometum rhenanum’ in the Kaiserstuhl. From von Rochow (1951) in Oberdorfer (1957). 7: ‘Xerobrometum lugdunense’ in the western Swiss Jura. From Quantin (1935) No. 8 and 9: Examples of suboceanic Mesobromion: 8: ‘Mesobrometum collinum’ in Kraichgau. From Oberdorfer (1957) 9: Gentianello-Koelerietum close to Göttingen. From Bornkamm (1960) a

7.3 Vegetation

527

7.3.2.2  Continental Calcareous Dry Grasslands In mid- and eastern Central Europe, a gradual transition occurs to the continental calcareous grasslands of the order Festucetalia valesiacae, which itself then transforms further east into the cold steppes of Ukraine and the Hungarian Plain. The character species of this order have a more continental range (EIV-C 6 or 5 to C 8, see Table 7.4). The westernmost occurrences of these communities are found in the dry valleys of the Valais Alps (see Table 7.5), in the Kaiserstuhl in southwest Germany and around the Nahe River west of Mainz, with many species of the Brometalia present in the latter two regions. There is a wide transitional zone between the Brometalia erecti and the Festucetalia valesiacae from the central German dry zone in Thuringia over Mainfranken and the Franconian Jura to Tyrol. This subcontinental-suboceanic transition zone includes the dry grasslands around Würzburg in northern Bavaria (see Table 7.6, no. 5). Both their species composition and the climate of their habitat could easily belong to the continental order of the Festucetalia valesiacae. The climate data in Table 7.6 and the vegetation table (Table  7.5) therefore clearly show how the two extremes are linked. Well-­ developed continental calcareous dry grasslands are found in the central German Saale-Unstrut region, on the slopes of the Oder valley, in the dry zone in the western Czech Republic, in the Pannonian area of Austria, at the eastern edge of the Alps, at the edge of the Hungarian Plain in western Slovakia, and in the dry areas of southern Poland with less than 550 mm, and in places with even less than 500 mm precipitation per year (Braun-Blanquet 1936; Klika 1939; Niklfeld 1964; Mahn 1965; Kolbek 1975; Toman 1992; Chytrý 2007). It is often the substrate that is the decisive factor in this transitional zone whether communities of the Festucetalia or of the Brometalia form. The vegetation on limestone soils contains more species of the Brometalia, whilst that on loess soils and base-rich silicate soils contains more Festucetalia species. The Stipo-Koelerietum vallesianum in Valais (no. 1 in Table 7.5) and the semi-­ dry grasslands around Göttingen (no. 9) share only a few class character species and accompanying species. The flora of the Valais grassland has more in common with the submediterranean Xerobrometum of the southwestern Jura (no. 7), but is still surprisingly dissimilar considering how close regions 1 and 7 are to each other. These inner Alpine dry grasslands (no. 1) are much more similar to the, geographically very distant, steppe grasslands in the low mountains of the northern Czech Republic near Litoměřice (no. 2) and the rocky steppe grasslands at the edge of the Rhine valley near Mainz (no. 3). It is only in these three communities that feathergrasses (Stipa capillata and S. pennata with subsp. joannis) play an appreciable role. Other species of the order Festucetalia valesiacae (F in Table 7.5, see also Fig. 7.14) also only occur here, whilst the species of the Brometalia (B) become more important in the right hand side of the table. These floristic similarities and differences are partially explained when considering the indicator values describing the regional climates of the grassland examples (see Table 7.6). In regions 1, 2 and 3, the climate is relatively continental, with low rainfall during the growing season. The submediterranean summer depression in

Stipo-Poion carniolicae 1. Valais: Sitten  Siders Festucion valesiacae 2. Bohemia: Litomericec 3. Rhine Hesse: Mainz  Oberlahnstein 4. Franconian Jura: Nuremberg  Amberg (Oberpf.) Xerobromion 5. Mainfranken: Würzburg

Grassland alliance Community no. (see Table 7.5) Region, location

638 536

502

512

590 585

677

560

177

94

77 320

525

179

Year

540 532

m a.s.l.

318

399

332 356

290

318

318 268

IV– IX

Precipitation (mm)

tends to Fest.

tends to Brom.

30

Quotientb July temp.: annual precip.

6.1

5.8

6.2

6.5

6.5

mT

4.7

4.6

4.7

5.3

4.9

mC

2.6

2.8

2.5

2.6

2.2

mM

Mean EIV valuesa

Table 7.6  Climatic data and indicator values for the dry grassland examples in Table 7.5. From data in Maurer et al. (1910), Reichsamt f. Wetterdienst (1939) and Quantin (1935), indicator values calculated from Ellenberg et al. (1992)a

528 7  Nutrient-Poor Dry Grasslands

747

734 728 607 802

398

214 258 155 242

417 431 378 436

443

480

422

>350 Brome-­ talia 8.8 8.6 8.5 7.6

10.0

9.5

9.7

9

18.1 17.5 17.2 17.3

18.5

19.5

18.7

24.5 24 28.5 21.6

25

22.5

28

50

>50

30–40

20–40

30

20

No

Never/ temporarily

170–>210

170–210

170–210

130–170

130–210

130–170

Max. amount of plant-­extractable water (0–70 cm) (mm) 130–170

< 60

< 60

100–>150

100–150

60–100

100–150

Depth of soil dryingb (cm) 100–150

From data in von Müller (1956) Each value is the range for 1–5 stands a Maximum water loss from the main root zone over 1 month in % of plant-available water; b Depth of dry soil maintained for 1 month; c In summer

Acid purple moorgrass meadow Bromus racemosus-­ Senecio aquaticus assoc.

Community Semi-dry manured pasture Semi-dry manured meadow Damp manured pasture Damp manured meadow Wet meadow

Degree of soil dryinga (%) < 20

Community dependence on Groundwater table >2 m groundwater Never No

No

Yes/partly

Partly

Yes/partly

Yes

Yes

Permanently

Temporarily/ permanently Permanently

Never/ temporarily

Temporarily

Never/ occasionally

Capillary fringe Water shortage at reaches root horizonc time of 2nd cut Yes Never

Table 8.4  Several hydrological parameters of meadow and pasture communities in the middle Weser valley

No

Yes

No

(Partly)

No

(Partly)

Alternating moist/dry No

8.2  Environmental Conditions and Habitat Classification 611

612

8  Agricultural Grassland on Mesic to Wet Soils

vides a quantitative measure of the frequency and duration of drought and flooding episodes in grassland communities (Niemann 1963). Numerous authors (e.g. Tüxen 1954a; von Müller 1956; Klötzli 1969b; Hundt 1970b, 1972; Balátová-Tulácková 1972; Pfrogner 1973; Meisel 1977; Grootjans and Ten Klooster 1980; Rosenthal 1992; Scholle and Schrautzer 1993; von ­Ruville-­Jackelen 1996; Leyer 2002) have studied the temporal dynamics of the groundwater regime in agricultural grassland communities using hydrographs and determined their typical maximum, minimum and average values as well as annual amplitudes (see Fig.  8.6 and Table  8.4). Although these communities will be ­discussed in detail in Sect. 8.3, we will look at the basic differences in water regime here, because it forms the basis of the ecological gradients in the grassland vegetation. The Bromo-Senecionetum aquaticae (marsh ragwort meadows) and Angelico-­ Cirsietum oleracei (cabbage thistle meadows) typically have maximum groundwater levels of only a few centimetres below the soil surface, or are even flooded for short periods. In summer, the groundwater in the former community can drop to around 80 cm, and in the latter community to around 140 cm below the surface. The Cirsium oleraceum-Arrhenatheretum is not only floristically but also hydrologically quite similar to the Angelico-Cirsietum, and also occasionally has standing water at the surface. The Arrhenatheretum in its wet (Deschampsia cespitosa-), typical, and moderately damp (Ranunculus bulbosus-) forms instead colonise increasingly dry soils with maximum water table levels of 75, 130 and 210 cm below the surface, and minimum levels of 190–230, 330–410 and 505 cm or lower (Tüxen 1954b). The deeper the average groundwater level, the greater the annual amplitude in its fluctuation in Arrhenatheretum communities (Meisel 1977; see Fig. 8.7). In contrast to the Cirsium oleraceum-Arrhenatheretum, which belongs to the damp meadows, Tüxen’s (1954b) typical groundwater form and the Ranunculus bulbosus-form of this community belong to the mesic meadows, which are little affected by groundwater or stagnant water. The strong desiccation of the topsoil of Arrhenatheretum communities in summer is shown by the measurements of Hundt (1970) in the middle Elbe valley. The plant-available water reserves dropped to 18 mm water in the soil profile from 0 to 120 cm, and the soil water content was at times only 5 vol%, similar to a dry grassland. Importance of Alternating Soil Moisture  Floodplain meadows are characterised by occasional flooding as well as strong summer droughts. They can therefore be described as periodically wet, i.e. the soil experiences alternating hypoxia and drought. The cause of the alternating soil moisture in this case is the combination of seasonal flooding and a climate with warm summers and relatively high evaporation rates. In oceanic climates with cooler summers, particularly waterlogged clay soils often alternate between wet and dry. Klapp (1965) considers the close proximity of species with low, and species with high water demand (e.g. of Brometalia and Molinietalia species) to be a good indicator of alternating soil moisture. It is occasionally stated that Molinia grassland habitats are always periodically wet, whilst the Angelico-Cirsietum oleracei and other fertilised moist meadows

8.2  Environmental Conditions and Habitat Classification

613

Fig. 8.6 (Parts a and b) Changes in the soil water regime in 1953 in six grassland communities with damp to dry soils in the middle Weser valley near Stolzenau (Lower Saxony). Precipitation, the groundwater level (black), the capillary fringe (dense hatching) and height of the vegetation are also shown. The whiter patches show drier soils (Modified from von Müller 1956)

614

Fig. 8.6 (continued)

8  Agricultural Grassland on Mesic to Wet Soils

8.2  Environmental Conditions and Habitat Classification

615

Fig. 8.7  The average amplitude of changes (bar) and lowest groundwater levels (horizontal line) in 15 grassland communities in northwestern Germany (number of measured locations in brackets, from which the average maximum and minimum was calculated). From Tüxen (1954a, in Dierschke and Briemle 2002; with permission of Ulmer Verlag, Stuttgart). >0 flooding (dark); 0–30 cm wet main root zone (light grey); GF groundwater form

have a more constant water supply. This is wrong in both cases. Some soils colonised by Calthion communities have extremely variable water levels, and both the water content and the water table can fluctuate to the same extent as the soils of some Molinion communities (see Table 8.5 and Fig. 8.8). In turn, some Molinia grasslands grow on soils with very little variation in water levels (e.g. Eskuche 1962; Balátová-Tulácková 1972; Rosset 1990). Pfrogner (1973) studied the groundwater levels in the Inn floodplain in Bavaria to determine the ecologically relevant hydrological thresholds (see Table 8.5). The results showed that Arrhenatheretalia (i.e. damp meadow) species were promoted when the summer water table was below 80 cm, whilst sedge species profited from one to two months of groundwater levels above 30 cm. Molinietalia (moist meadow) species had a competitive advantage when the groundwater remained between 30–60 cm below the surface for up to one month. The nutrient-rich Lolio-Cynosuretum pastures can also form clearly differentiated types depending on the groundwater regime. These are the Glyceria fluitans type with periodic surface water and minimum levels of around 80 cm below the surface, the moist Lotus pedunculatus type (with an amplitude of roughly 20–110 cm) as well as the damp (typical) type and the dry Plantago media type (Tüxen 1954b). The last two types are not affected by the groundwater and are less clearly

8  Agricultural Grassland on Mesic to Wet Soils

616

Table 8.5 Maximum and minimum groundwater levels in various agricultural grassland communities and in a sedge community Highest Community n level Tall-sedge swamps and reed Caricetum gracilis 1 +10

Lowest level

Caricetum gracilis

1

Caricetum elatae

1

Annual mean

Region

Author

30

Brandenburg

+10

15/80

Switzerland

+25/+5

45/140

Switzerland

Hundt (1972b) Klötzli (1969) Klötzli (1969) Hundt (1972b) Hundt (1972b) Balat.-Tul. (1972) Ruville-J. (1996)

Phalaridetum 1 +10 < 150 arundinaceae Glycerietum 1 +10 < 150 maximae Caricetum 6 +12/0 13/29 appropinquatae Caricetum 1 +10? acutiformis Small-sedge swamp and similar communities Caricetum diandrae 1 0 45

Brandenburg Brandenburg Silesia 19–35

Münsterland

Switzerland

Caricetum diandrae

1

+15/+24

13/16

Silesia

Caricetum vulpinae

1

+100

150

Silesia

Caricetum davallianae Orchido-Schoenetum

1

0

45/95

Switzerland

1

0/15

20/120

Switzerland

‘Schoenetum’

1

5

37

Tall-forb communities Valeriano-­ 1 0/15 20/120 Filipenduletum Filipendulo-­ 12 +3/15 36/60 Geranietum Temporarily flooded river valley grassland Cnidio-Violetum 1 +30 240 SanguisorboSilaetum Cnidio-­ Deschampsietum Ranunculo-­ Alopecuretum geniculati

29

Switzerland

Switzerland Silesia

1

+10

115

1

+10

< 150

Northwest Germany Northwest Germany Brandenburg

2

+10

17–97

Münsterland

Klötzli (1969) Balat.-Tul. (1972) Balat.-Tul. (1972) Klötzli (1969) Klötzli (1969) Rosset (1990) Klötzli (1969) Balat.-Tul. (1972) Meisel (1977) Meisel (1977) Hundt (1972b) Ruville-J. (1996) (continued)

8.2  Environmental Conditions and Habitat Classification

617

Table 8.5 (continued) Community n Wet meadows (Calthion) Angelico-Cirsietum 5 heracleetosum Angelico-Cirsietum 1 oleracei 1 BromoSenecionetum aquaticae BromoSenecionetum aquaticae Cirsietum rivularis 9

Highest level

Lowest level

Annual mean

Region

Author

34

144

92

South Bavaria

+10

70

Saxony

10

75

Northwest Germany

Pfrogner (1973) Hundt (1970) Meisel (1977)

+10

50–135

Münsterland

Ruville-J. (1996)

+5/10

40/60

Silesia

Balat.-Tul. (1972)

Münsterland

Ruville-J. (1996)

Silesia

Balat.-Tul. (1972) Klötzli (1969) Klötzli (1969) Rosset (1990) Klötzli (1969)

Fertilised pastures (Cynosurion) Lolio-Cynosuretum 2 typicum Purple moor grass meadows Selino-Molinietum 6 2/6 Saturejo-Molinietum, 2 various subass. Stachyo-Molinietum, 9 various subass. Molinietum 1 caeruleae Junco-Molinietum, 4 various subass. False oatgrass meadows Arrhenatheretum 36 cirsietosum Arrhenatheretum 1 lychnidetosum Arrhenatheretum 9 typicum 1 Arrhenatheretum typicum (with Sanguisorba officinalis)

12/21

100/170

31/41

8/16

20/45

180

Switzerland

+10/45

30/170

Switzerland

35

60

10/30

40/190

4/53

121/141

70

120

82

235

+10

275

52

Switzerland Switzerland

56

South Bavaria

147

Northwest Germany South Bavaria Northwest Germany

Pfrogner (1973) Meisel (1977) Pfrogner (1973) Meisel (1977)

The numbers refer to cm below the soil surface or cm above the surface (with +). If there is more than one value then these were measured in different stands or different years Balat.-Tul. Balátová-Tulácková, SAss. Subassociation

618

8  Agricultural Grassland on Mesic to Wet Soils

Fig. 8.8  Not all Molinietum communities have variable soil moisture! Modified from Eskuche (1962). In the soil of a Junco-Molinietum (Carex fusca variety) in the Erft lowlands (western Germany), the plant-available water content (in vol%) and groundwater level varied surprisingly little in the period between 1956 and 1958. The water supply to Molinia grasslands is only very variable when on marl soils; on all other soils, the water levels are largely similar to fertilised moist meadows

characterised by their soil hydrology (amplitude: 27–469 cm below the soil surface) than the other meadow and pasture types discussed above (see Fig. 8.6). Comparing the groundwater types found in northwestern Germany with those in southern or eastern Central European grasslands shows some variation. These differences are to some extent caused by the climate, but sometimes simply by the specific physical properties of the grassland soils. For example, the typical Arrhenatheretum studied by Pfrogner (1973) in the Inn plain in southernmost Germany had higher maximum groundwater levels (up to 82 cm below the soil surface) and lower minimum levels (235 cm) than the Arrhenatheretum studied by Tüxen (1954b) in northwestern Germany (see Fig. 8.7), i.e. the amplitude was much larger. Role of Capillary Rise  As well as the groundwater, the capillary zone in the soil profile is also very important for the water supply to the grassland plants, particularly in damp meadow and pasture communities on soils rich in loam and clay with groundwater levels below 1 m. The thickness of the capillary zone depends on the particle size of the soil. In coarse sands it only reaches up to around 20 cm above the water table, whilst in floodplain loams and strongly decomposed fen peat it can be over 2 m thick (Wohlrab 1966). In loam soils, crop plants may profit from groundwater through capillary rise even from a depth of 3.3 m (Scheffer/Schachtschabel 2010). Communities with their roots in a thick capillary zone are therefore generally well-supplied with water, even when the groundwater is deep below the soil surface (see Table 8.4 and Fig. 8.6). Soil Water Reserves  Grasslands that are not affected by the groundwater are hydrologically best characterised by the seasonal changes in their reserves of plant-­ available water (PAW), i.e. the maximum amount of water held at matric potentials between −0.01 and −1.5 MPa in the root zone of the soil profile (cf. Fig. 4.16 in Vol. I) and its exhaustion during summer. These values are largely determined by the grain size of the soil. According to Voigtländer and Jacob (1987), clay soils have a PAW over twice as high as sandy soils, and for peaty soils this is even five times higher (see Fig. 8.9). Hundt (1972b) recorded the seasonal changes in PAW in wet meadows, sedge communities, damp meadows and dry grasslands in eastern Germany over a period of 2 years. After winter saturation, the moist to wet soils of

8.2  Environmental Conditions and Habitat Classification

619

Fig. 8.9  The field capacity, capacity of plant-available water and storage of unavailable water of different soils in mm water in the soil profile from 0 to 100 cm (From Czeratzki and Korte (1961) and Kuntze (1981, in Ehlers 1996))

the Glycerietum maximae, Caricetum gracilis and Angelico-Cirsietum oleracei contained 300–400 mm water in the soil profile down to 120 cm below the surface, and only dropped below 200 mm very occasionally in summer. The wet and to some extent alternating wet and dry Arrhenatherion and Cnidion communities achieved a maximum of 200 mm in spring, of which just under 100 mm was still available in mid-summer. Both stands will have profited particularly from capillary rise transporting water from the groundwater in summer. Despite the same amount of rainfall in spring, the Festuco-Stipetum capillatae steppe grassland and the Corynephoretum canescentis on exposed slopes or sandy soils only retained around 100 mm available water to a depth of 120 cm, and both stands consumed this water to leave only a few mm by mid-summer (see Fig. 8.10).

8.2.4  Soil Chemical Properties Grasslands are found on soils across the full pH spectrum in Central Europe, from highly acidic Nardus grasslands and nutrient-poor pastures to dry grasslands and calcareous Salvia-Arrhenatheretum with a pH >7. Increasing soil acidity is expressed not only in an increasing concentration of exchangeable acidic (H, Al, Fe and Mn) cations, but perhaps more importantly in a decrease in the percentage base saturation (see Table 8.6). This parameter provides the best classification of agricultural

620

8  Agricultural Grassland on Mesic to Wet Soils

Fig. 8.10  Seasonal changes in the storage of plant-available soil water over the course of 2 years in eight grassland communities with moist to dry soils in eastern Germany. The soil moisture storage increases from the Corynephoretum canescentis to the Arrhenatheretum elatioris and Cnidio-­ Deschampsietum to the Caricetum gracilis (From Hundt 1970, 2001)

8.2  Environmental Conditions and Habitat Classification

621

Table 8.6  Exchangeable cations, base saturation and acidity in soils of different grassland communities

Community Dry meadows on limestone Bromus-­ Arrhenatheretum Caricetum davallianae Dry Lolio-Cynosuretum Lolio-Cynosuretum Damp Lolio-Cynosuretum Nutrient-poor Lolio-Cynosuretum Montane Arrhenatheretum Trisetetum Scirpus sylvaticus swamp Nardetum Carex nigra fen

Cation exchange capacity (μmolc g−1) 373

pH (H2O) 7.1–8.0

Base saturation (%) 94.7

Exchangeable base cations (μmolc g−1) 355

7.1–8.0

89.2

163

20

183

7.1–8.0

87.8

323

44

367

6.1–7.0

83.7

161

32

193

6.1–7.0 5.1–6.0

70.3 62.4

117 181

52 117

161 298

6.1–7.0

56.4

89

68

157

5.1–6.0

46.3

152

172

324

5.1–6.0 5.1–6.0

32.1 25.5

79 96

172 252

251 349

4.1–5.0 4.1–5.0

22.6 11.2

62 40

218 353

280 393

Exchangeable acid cations (μmolc g−1) 18

From data in Klapp (1965) pH values: most common values

grasslands according to their soil chemical properties. The relative proportion of exchangeable basic cations and their concentration in the soil mainly influence the species composition of the grassland community, whilst the productivity is more affected by soil moisture than soil chemistry (see Sect. 8.5.2). The above is mainly true for unfertilised grasslands. High levels of fertiliser application causes the homogenisation of the flora and reduces productivity differences, whether the soils are sand, loam or peat, i.e. intensive grasslands are largely unaffected by the soil properties (Klapp 1950). It is only in the upper montane belt that the effects of soil and climatic factors on the species composition become clearer (see Chap. 5). Grassland soils typically have much higher carbon and nitrogen pools than arable fields on the same substrate. Klapp (1965) and Scheffer (2002) provide typical values of 0.8–1.5 % for the carbon content in the uppermost horizon of Central

622

8  Agricultural Grassland on Mesic to Wet Soils

European arable fields on mineral soils, compared to 2–4 % for pastures and meadows. The soil N content in arable fields on mineral soils is around 0.10–0.15 % and for agricultural grassland 0.2–0.6 % (Scheffer/Schachtschabel 2010). The higher concentrations in grassland soils are mainly the result of the lack of mechanical disturbance of the soil, but perennial grassland plants also possess larger root systems that enrich the soil C and N pools through root necromass production and rhizodeposition. The C and N contents rise with increasing elevation. For example, Hachmöller (2000) found C contents in the topsoil of acid montane grassland in the Ore Mountains of 3.1–4.5 % (maximum 7 %) and N contents of 0.29–0.52 % (maximum 0.72 %), which is in the upper range of the concentration values given above.

8.3  T  he Vegetation of Agricultural Grasslands and Roadside Verges 8.3.1  O  verview of the Agricultural Grassland Communities of Central Europe In addition to mechanical factors such as mowing, grazing and trampling, the species composition of grassland communities is also influenced by edaphic and climatic factors, in particular the temperature and duration of the growing season and the water and nutrient availabilities. We will first consider the true meadows, i.e. grassland communities that are only or mainly managed by mowing, and are not affected by selective grazing or soil compaction from livestock trampling. In order to provide a brief overview of the diversity of habitats occupied by Central European meadow communities, as well as a comparison with the forest communities that they replace, Fig. 8.11 displays the meadow communities within the same ecogram as was used in Chap. 4 (Vol. I) for the forest communities. As in the discussion of the forests, we will also base the discussion here on the submontane belt, of which agricultural grassland covers a particularly large area. One fundamental distinction is between unfertilised and fertilised meadows, in which nutrient loss through mowing is replaced by the application of fertiliser (see Fig. 8.11a, b). Another basic distinction is the water supply. The wettest meadows are found in the terrestrialisation zones of lakes and other still water bodies, the naturally treeless parts of which are discussed in Chap. 4. Only reed beds, tall-sedge communities and some fens can be seen as natural grasslands in a geobotanical sense. In an agricultural sense, some wet and damp meadows that could support tree growth but are too wet to plough are also termed ‘natural’ grassland (see Fig. 8.11b), as are very dry, stony or sandy soils prone to erosion that are also unsuitable for arable cultivation. Sedge marshes are created by low-intensity mowing in swamp forest habitats, the higher lying parts of fens and around springs in flushes. These have also been discussed in the context of natural mire communities in Chap. 3.

8.3  The Vegetation of Agricultural Grasslands and Roadside Verges

623

Fig. 8.11  An ecogram of the alliances of unfertilised (top, a) and moderately fertilised (bottom, b) grassland communities in the submontane belt of Central Europe. Almost all stands on soils that are moderately damp and not too acidic are intensively used. There are no fertilised meadows in the driest habitats because the addition of nutrients has little effect, or on highly acidic soils because fertilisation would mean an input of bases here thus reducing the acidity (Modified from Ellenberg 1996)

624

8  Agricultural Grassland on Mesic to Wet Soils

Instead of the mixed broadleaved forests that would naturally grow on damp mineral soils and drained peat soils, nutrient-poor Molinion litter meadows or nutrient-­ rich Calthion communities form under mowing management. The Molinion and Calthion alliances are floristically so similar that they are combined in a single order (Molinietalia). If unfertilised, all wet and damp meadows produce fibre-­ rich and protein-poor, tough forage, but which is well suited for animal bedding. They were therefore used until around 1950 almost exclusively as litter meadows. For the same reason, semi-dry grasslands (Mesobromion) were also often used to produce bedding, particularly in southern Germany and Switzerland (see Chap. 7). If nutrient-poor meadows are regularly fertilised and mown, then species with high fodder value, particularly those of the Arrhenatherion, will quickly start to dominate if the soil is damp enough. Without the addition of nutrients, the Arrhenatheretum or similar communities could only develop in the floodplains of rivers at the level of the hardwood forests (see Fig. 9.12 in Vol. I). The softwood floodplain forests are replaced by the Angelico-Cirsietum oleracei or other fertilised moist meadow types (Calthion) if they are drained. The ecogram of meadow communities is based on the situation in the submontane belt of Central Europe, but may be interpreted for the colline and lowland zones with the caveat that the moisture levels in Fig. 8.11 relate to the combined effects of climatic and edaphic factors. A deep loess soil, for example, can support a Mesobromion grassland in a dry and warm climate, or a wet grassland if the precipitation is higher. Similarly to the situation in forests, the montane and subalpine climate restricts the growth of so many plant species that the submontane and colline communities are replaced by different ones at these elevations. Rocky communities as well as dry and semi-dry grasslands contain more montane and subalpine species and fewer thermophilic species with increasing elevation. Eventually, the Brometalia communities lose their character as described in Chap. 7. Nardus grasslands and small-­ sedge communities, in contrast, only reach their optimum development in the montane belt. Fertilised grasslands are dominated here by montane ­Polygono-­Trisetion communities instead of the Arrhenatherion. Arrhenartherum, Trisetum, Cirsium oleraceum and Molinia grasslands share so many species that they are combined in the class of agricultural grasslands (Molinio-Arrhenatheretea). Nutrient-rich and fertilised pastures are floristically so homogenous that they are placed in a single alliance, the Cynosurion cristati. Only the pastures of the upper montane to alpine belts on moist loam soils are given their own alliance, the Poion alpinae. The nutrient-poor pastures merge with the Violion caninae and Mesobromion. In the following sections, we will describe the most important communities of Central European meadows and pastures in more detail. An overview of the orders and alliances of agricultural grassland vegetation in the class Molinio-­ Arrhenatheretea is given in Chap. 14: no. 5.4. However, it should be noted that there is still disagreement in the literature about the classification of this vegetation formation, and there are several differences of opinion e.g. in Oberdorfer (1993a),

8.3  The Vegetation of Agricultural Grasslands and Roadside Verges

625

Mucina et al. (1993), Pott (1995), Schaminée et al. (1996), Dierschke (1997a), Berg et  al. (2004), Burkart et  al. (2004), Chytrý (2007) and other overviews. As most grassland across Central Europe has been undergoing a process of transformation driven by management intensification and climatic changes in recent years, phytosociologists will have to come to terms with the fact that there will never be a generally valid and satisfactory systematic classification of agricultural grassland.

8.3.2  Mesic Meadows We will use the term mesic meadows here to discuss all mown grasslands that occur in mesic habitats. With low to moderate fertilisation, these soils support various types of Arrhenartherum (false oat grass) and Trisetum (golden oat grass) meadow, and without fertilisation probably the Festuca rubra-Agrostis grassland and FestucoCynosuretum (see Fig. 8.11a). Grasslands on wet soils (i.e. Molinion, Calthion, see Sect. 8.3.3, or Caricion nigrae, Caricion davallianae, see Sects. 3.3.4 and 3.3.5), on dry soils (Mesobromion, see Chap. 7) and on highly acidic and base-poor soils (Violion caninae, see Chap. 6) are discussed in other sections. 8.3.2.1  A  rrhenartherum Meadows from Lowland to Submontane Elevation 8.3.2.1.1  Arrhenartherum Meadows and Their Habitat Types Arrhenartherum meadows are the most recent type of meadows to develop in Central Europe, but in southwestern Germany and the lower areas of the Alps, they are by far the most common, even today when most areas have already been homogenized by intensive mowing and grazing (see Sect. 8.7.1). They have long played a central role in hay production. Like Arrhenatherum elatius, the Arrhenatherum meadows have a suboceanic-­ submeridional distribution that stretches from the northern edge of the Mediterranean region to northern Central Europe. They are most diverse in southwestern Germany, where Schreiber (1962, and references contained within) studied their habitat and geographic variability, and in the neighbouring Swiss Plateau, where they were identified as early as the nineteenth century by Stebler and Schröter (1887–1892) as a unique community. The continental parts of Central and Eastern Europe do not support the growth of typical Arrhenatheretum communities, and they are rare in the dry and warm basins. Dierschke (1997a) considers that it is most practicable to include all Central European Arrhenatherum meadows in a broad association (Arrhenatheretum elatioris). Like beech, these communities grow best in the submontane belt of the subatlantic region, although they prefer warmer temperatures and do not occur at such high elevations or latitudes as beech.

626

8  Agricultural Grassland on Mesic to Wet Soils

The Central European Arrhenatheretum communities are found in areas that would naturally support beech, oak, hornbeam and floodplain forests, and under moderate fertilisation and one to three cuts per year are reasonably intensive agricultural habitats. Dierschke and Briemle (2002) state that early Arrhenatherum meadows developed without the application of fertiliser on naturally nutrient-rich floodplains after forest clearance. The most species-rich and best characterised Arrhenatheretum meadows are cut twice a year, as was common over a century ago, and mainly fertilised with manure. More frequent cutting and intensive fertilisation plus periodic grazing may increase yields, but decrease species richness as well as reducing the number of character species. As more intensive management (or abandonment) has replaced the low-­ intensity agricultural grassland types, so the ‘typical’ Arrhenatheretum meadow has slowly disappeared. Similarly to the once widespread Calluna heaths of northwestern Germany, the remaining fragments of Arrhenatherum meadows of southwestern Germany and northern Switzerland must be placed under nature conservation measures in order to ensure their survival for future generations (Nowak and Schulz 2002). However, most regions still contain some Arrhenatheretum communities under less intensive management, and it is these that we will focus on in the following. All character species with high constancy and fidelity to the Arrhenatherion alliance belong to the upper and middle structural layer of typical Arrhenatherum meadows, namely Arrhenatherum elatius itself, Tragopogon pratensis, Crepis biennis, Campanula patula, Galium album, Anthriscus sylvestris, Heracleum sphondylium and Geranium pratense (see Table 8.7). Even species less tightly linked to the Arrhenatherion tend to be upper- and middle-layer grasses and tall forbs and legumes (see Fig. 8.12). The low-growing white clover (Trifolium repens) and low grasses such as Lolium perenne are suppressed, unless they are shade-tolerant. Low-­ growing species such as Bellis perennis, Ajuga reptans and Lysimachia nummularia can survive by making use of the high light intensities after mowing, ensuring that all available niches in the grassland are filled (see Fig.  8.13). Even the grasses ­produce dense layers of leaves close to the soil surface and use the available light to the maximum, similarly to forest herbs on fertile soils (see Figs. 4.56 and 4.60 in Vol. I), i.e. down to 10 % or less of the incident light (see Fig. 8.4). The lower layers of the meadow vegetation therefore make a considerable contribution to the total yield, particularly in spring and after mowing, when light levels are higher for them. Differences in the water regime and increasingly also the level and type of fertilisation produce different forms of Arrhenatherum meadow (with 2 or more cuts per year). Several Arrhenatheretum subassociations or facies dominated by certain species can be distinguished using groups of differential species (see Fig. 8.11b). The typical Arrhenatherum meadows (Arrhenatheretum typicum) form on mesic soils. Fertilisation only with manure promotes tall Apiaceae, particularly Anthriscus before the first cut and Heracleum before the second, because their broad leaves quickly rise above and suppress the grasses. These Apiaceae facies (see Fig. 8.14) produce less valuable fodder than the grass-rich meadows, as their dried leaves crumble and their stems become woody. The tall Apiaceae only completely domi-

8.3  The Vegetation of Agricultural Grasslands and Roadside Verges

627

Table 8.7  Arrhenatheretum and Bromus racemosus-Senecio aquaticus meadows in southwestern and northwestern Germany Region: No.: Arrhenatheretum (sensu lato): Arrhenatherum elatius Trisetum flavescens Galium mollugo Heracleum sphondylium Leucanthemum vulgare Dactylis glomerata Bellis perennis Veronica chamaedrys Achillea millefolium Vicia sepium Lolium perenne Avenochloa pubescens Lolium multiflorum Knautia arvensis Anthriscus sylvestris Pimpinella major Tragopogon pratensis Campanula patula Crepis biennis Pastinaca sativa Geranium pratense Indicators of dryness: Salvia pratensis Thymus pulegioides Scabiosa columbaria Bromus erectus Koeleria pyramidata Lotus corniculatus Festuca ovina agg. Luzula campestris Ranunculus bulbosus Briza media Silene vulgaris Medicago lupulina Plantago media Campanula rotundifolia Indicators of moisture: Glechoma hederacea Ajuga reptans Silene dioica

A Danube 1 2 5 5 5 4 5 4 2 5 5 3 1 4 4 5

3 2 •

5 5 5 4 4 4 4 4 5 3 4 5 4 5 1

5 5 5 4 5 5 5 4 4 3 4 4 4 4 3 1 2 • •

B North Rhine 1 2 3

3

4

5 5 5 5 5 5 1 5 3 3 1 4 3 5 2 1 1 1 • • 1

5 5 5 5 4 4 2 2 2 2 1 4 2 5 5 4

5 5 3 4 3 5 2 3 3 2 3 2

5 4 2 5 3 5 3 2 2 3 3 1 1

2 2 1

3 3 1

• • 1

2 2

3 2

1 1 2 2 3 2 3 1

4

1

1

1

1

1 2 1 2 2

2 2

1 5 2

4 5 4 4

1 1 5

1

3 4 3

3 4 5

4 4 4

1 1

1

1 1

3 1

1

1

2 3

1 2

(continued)

628

8  Agricultural Grassland on Mesic to Wet Soils

Table 8.7 (continued) Region: No.: Festuca pratensis Deschampsia cespitosa Alopecurus pratensis Poa trivialis Cardamine pratensis Ranunculus repens Rumex crispus Lysimachia nummularia Indicators of moisture and flooding: M Lychnis flos-cuculi M Filpendula ulmaria M Angelica sylvestris C Cirsium oleraceum M Geum rivale M Sanguisorba officinalis C Polygonum bistorta Alchemilla vulgaris agg. Phalaris arundinacea M Lythrum salicaria Carex gracilis C Myosotis palustris C Bromus racemosus C Senecio aquaticus C Lotus uliginosus Juncus articulatus C Caltha palustris M Equisetum palustre Carex disticha Widespread meadow plants: Festuca rubra subsp. rubra Rumex acetosa Ranunculus acris Holcus lanatus Cerastium holosteoides Poa pratensis Trifolium pratense Centaurea jacea Lathyrus pratensis Vicia cracca Prunella vulgaris

A Danube 1 2 1 1 1 2

3 1 5 3 2 2 3 1

4 1 5 4 5 4 5 2 2

1 3 1 5 5 5 3 4 1

2 5 5 5 5 4 5 2 4 4 4 2

B North Rhine 1 2 3 4 5 5 1 1 2 1 5 3 2 5 5 2 5 5 5 2 1 1 3 1 1

5 4 1

5 5 3

1

3 3 4 2 2 2 1 3 2

5 3 4 4 4 4 3 3 3

5 5 5 5 4 3 2 2 3 3 1

5 5 5 5 5 2 3 2 4 4 1

1

1 1 4 4 4 4 5 5 5 5 4 5 4

5 4 4 4 4 5 5 4 3 1

4 5 5 4 3 5 5 3 5 1 2

5 5 5 5 4 4 3 2 4 1 1

5 4 5 5 4 2 3 4 4 2 1

4 5 1 3 5 5 4 1 3

5 5 5 5 4 4 2 3 2 3 1

(continued)

8.3  The Vegetation of Agricultural Grasslands and Roadside Verges

629

Table 8.7 (continued) Region: No.: Phleum pratense Daucus carota Cynosurus cristatus Trifolium dubium Bromus hordeaceus Other species: Plantago lanceolata Taraxacum officinale Trifolium repens Anthoxanthum odoratum Elymus repens Agrostis capillaris

A Danube 1 2 2 3

3

4 2 1

3 1 5 4 3 3

5 5 5 2 2

5 5 3 2 2 3

5 5 2

B North Rhine 1 2 3 1 1 1 2 1 2 1 3 3 2 3 2 3 3 5 4 3 2

2 3

5 5 3 2 1 1

5 • 4 5

4 1 4

4 3 4 5

From data in Eskuche (1955) and Meisel (1960) Further species: Danube no. 1: Carex caryophyllea (3), Linum catharticum (3), Euphrasia rostkoviana (3) Species that do not have a constancy value greater than 2 in any of the vegetation units were mostly omitted No. A 1–4 and B 1 and 2 are Arrhenatheretum communities, and therefore belong to the Arrhenatherion alliance and Arrhenatheretalia order. The damp meadows no. B 3 and 4 contain numerous plants of wet and damp habitats, but only a few species of the Calthion (C) alliance and Molinietalia (M) order. Some of these avoid highly acidic soils, and are therefore absent from the Bromus racemosus-Senecio aquaticus meadows The groundwater does not reach the root zone of the vegetation units A 1 and B 1, and only periodically in A 2 and B 2. In A 3 and B 3 the groundwater does not reach the topsoil, whilst in A 4 and B 4 it is often saturated A Danube valley near Herbertingen (from Eskuche) No. 1: Salvia-Arrhenatheretum (‘Ranunculus bulbosus subassociation, Salvia pratensis variety, Bromus erectus subvariety’) No. 2: typical Arrhenatheretum, described by Eskuche as ‘Alopecurus pratensis subassociation, Deschampsia cespitosa variety, Silene cucubalus (= vulgaris) subvariety’ No. 3: Cirsium-Arrhenatheretum (‘Cirsium oleraceum subassociation, typical variety, pure form’) No. 4: Carex-Arrhenatheretum (‘Cirsium oleraceum subassociation, Carex gracilis variety’) B Lower Rhine near Moers (from Meisel) No. 1: Briza-Arrhenatheretum (‘Briza media subassociation, typical variety, typical subvariety’) No. 2: Alopecurus-Arrhenatheretum (‘Alopecurus pratensis subassociation, typical variety, typical subvariety’) No. 3: Bromus racemosus-Senecio aquaticus association, Trifolium dubium subassociation (with soil moisture levels similar to those of the Cirsium-Arrhenatheretum in the Danube valley) No. 4: Bromus racemosus-Senecio aquaticus association, typical subassociation (with soil moisture levels similar to those of the Carex gracilis variety of the Arrhenatheretum in the Danube valley)

Fig. 8.12 Cross-section through a typical Arrhenatheretum meadow (mesic subassociation) in the Saale valley above Wörmlitz (eastern Germany) (From Hundt 1958) From left to right: Arrhenatherum elatius, Pastinaca sativa, Poa pratensis, A. e., Vicia sepium, A. e. flowering, Poa p., Daucus carota, Galium mollugo, Geranium pratense, Crepis biennis. Shown to 75 cm below the soil surface

Fig. 8.13  Extensively managed, forb-rich Arrhenatheretum on sandy soil on the Baltic Sea island of Hiddensee (eastern Germany) which is mown once a year and not fertilised for nature conservation purposes. Visible are among others Galium x pomeranicum (G. album x G. verum), Daucus carota, Senecio jacobaea, Armeria maritima, Cichorium intybus, Vicia cracca, Cirsium arvense, Arrhenatherum, Dactylis and Festuca rubra

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Fig. 8.14  Top: Cross-section through Arrhenatheretum, Polygono-Cirsietum and Caricetum gracilis stands. Series of hay meadow communities with increasing moisture levels in fen soils in Auetal west of Brunswick (Lower Saxony), under high nutrient levels. The dry-mesic type is the typical Arrhenatheretum. The mesic type is referred to in northern Germany as the Alopecurus pratensis subassociation (see Fig. 8.16 top) Bottom: Forms of the Arrhenatheretum caused by different levels of fertilisation and management. Modified from Ellenberg (1952a). All of these cross-sections depict the vegetation before the first cut. The numbers refer to the average ranges in hay yield in 10 kg per hectare. The variation in yield caused by fertilisation is greater than that caused by water supply. The yields in the top diagram are for the second cut, and in the lower diagram for the first cut. The lower the nutrient levels, the greater the role of the low-growing grasses, which lose less of their phytomass through mowing

nate where their fruits have time to ripen, i.e. in meadows that are regularly mown later in summer. Anthriscus and Heracleum produce only leaves and a storage root or rhizome in the first year, but in the second year produce a tall flowering stem and die after the seeds ripen. Such species cannot spread in the modern, frequently mown grasslands, even under high nutrient conditions, as their stems are not left intact long enough for the seeds to ripen. More disturbance-tolerant N-demanding species such as Taraxacum officinale take over.

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If the nutrient inputs are too low to compensate for the losses from mowing, then the tall grasses and forbs are less productive than in the situation just described. Grasses in the middle layer such as Trisetum flavescens or Holcus lanatus and low-­ growing grasses tolerant of low nutrient levels such as Poa pratensis, Festuca rubra or even Agrostis capillaris can spread with the higher light availability. They dominate together with Plantago lanceolata and other low-growing forbs in the most nutrient-poor Arrhenatheretum communities. Such meadow communities would be difficult to assign to an association without the occasional stunted individual of Arrhenatherum or other character species of the Arrhenatherion. There are many intermediate types between the Poa pratensis facies and the Arrhenatherum facies (see Fig. 8.14). On damper soils, some species tolerant of wet conditions mix with the characteristic species combination of the Arrhenatherum meadows, e.g.: Angelica sylvestris Cirsium oleraceum Deschampsia cespitosa

Filipendula ulmaria Geum rivale Sanguisorba officinalis

Silene flos-cuculi Bistorta officinalis

The better water supply (see Fig.  8.14) means that these damp Cirsium -Arrhenatheretum communities (see Table  8.7, A 3) are higher-yielding than the typical form under the same management. As they mostly consist of good fodder plants and their soil is unsuitable for arable cultivation, these are considered particularly valuable agricultural grasslands. This damp Arrhenatheretum is also found in various forms depending on its nutrient levels. The Arrhenatherum facies of the typical Arrhenatherum meadow is replaced here by an Alopecurus pratensis facies, with both Arrhenatherum elatius and Festuca pratensis. Low levels of fertilisation cause stands rich in Holcus lanatus and Festuca rubra to form. Apiaceae facies are rare here and generally an indicator of abandonment. They are usually dominated by Angelica silvestris, which also grows well in unfertilised litter meadows (see Fig. 8.23). In dry and warm regions such as the Upper Rhine Plain, areas with high water tables support Arrhenatherum meadows that, apart from the above-mentioned species of damp soils, also contain tall sedge species and other marsh plants, e.g.: Caltha palustris Carex acutiformis Carex gracilis

Lythrum salicaria Phalaris arundinacea

These Carex-Arrhenatheretum communities (see Fig.  8.14 and Table  8.7, A4) produce low-quality hay, although often in large quantities. If well managed, meadow foxtail (Alopecurus pratensis) or reed canary grass (Phalaris arundinacea) can come to dominate. Even wetter soils are occupied by the Angelico-Cirsietum oleracei or other moist meadow types (see Sect. 8.3.3.1). Moving from typical Arrhenatherum meadows to drier soils, more Mesobromion species start to occur. The first indicator of drier conditions is usually Ranunculus

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bulbosus, followed in southern Germany by Salvia pratensis. Well-developed Salvia-Arrhenatheretum communities (see Table  8.7, A 1) include the following species: Bromus erectus Ranunculus bulbosus

Salvia pratensis Scabiosa columbaria

This subcommunity of the Arrhenatheretum is usually particularly rich in character species. It contains numerous species with striking flowers, and its colourful mixture and changing aspect make it one of the most beautiful plant communities in Central Europe. Unfortunately, however, its low yields mean that this community is one of those in continuous decline. On lime-free and often sandy soils, the Hypochaeris-Arrhenatheretum merges with Nardus grassland and sandy dry grasslands (Lisbach and Peppler-Lisbach 1996). 8.3.2.1.2  Geographical Variation in Arrhenatherum Meadows Southwestern Central Europe  It is only in the centre of its range, i.e. in southwestern Germany and the neighbouring Swiss Plateau, that the Salvia pratensisand typical Arrhenatheretum (each of which with typical and periodically wet varieties) as well as the Cirsium oleraceum-Arrhenatheretum and Carex-­ Arrhenatheretum are all found in well-developed forms. They are still found here and there in close proximity to one another, particularly in the foothills of the Swabian Jura and the Black Forest, as long as large enough areas of calcareous soil are available. In these areas it was easy to distinguish all the subassociations, varieties and even subvarieties (see e.g. Ellenberg 1952a; Eskuche 1955; Schreiber 1962). In particular, it was easy here to find large areas of the typical subassociation, i.e. an Arrhenatheretum without species of wetter or drier habitats (Nowak and Schulz 2002; see Fig. 8.15). In the Upper Rhine Plain and other landscapes with relatively low levels of precipitation, the typical Arrhentheretum shrinks to a narrow transitional zone between drier and wetter sub-communities (Oberdorfer 1952, 1957, 1983b). Almost all meadow vegetation where the roots do not reach the groundwater, even on deep and non-sloping soils, belongs to the Salvia-Arrhentheretum or similar ‘dry’ Arrhenatheretum communities. In contrast, soils with high water tables support numerous species of damp or wet conditions, which may even mix with Mesobromion species on sandy topsoils. On heavy marl, in which the water supply does not depend on the groundwater, the typical Arrhenatheretum is sometimes found over large areas as long as it is sufficiently fertilised, but as a variety of alternating wet and dry conditions that contains some indicators of damp conditions such as Silene flos-cuculi and Ajuga reptans. Even drier areas of southwestern Central Europe are characterised by the domination of the Salvia-Arrhentheretum on all fertilised but not grazed grassland that is

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Fig. 8.15  Locations of Arrhenatheretum communities in southwestern Germany, i.e. regions in which more or less all six subunits of the Arrhenatheretum occur (see the explanation top left) (Modified from Schreiber 1962). The habitat of the typical variety of Salvia-Arrhenatheretum is relatively dry, whilst that of the typical variety of the typical Arrhenatheretum is mesic. The varieties of alternating moisture levels colonise clay soils, or other areas that are periodically flooded. The Cirsium-Arrhenatheretum is found on damper, and the Carex-Arrhenatheretum on relatively wet soils (cf. Fig. 8.11). The six segments of the outer ring show whether each subunit is absent (white) or how well developed it is (hatched or black). The inner circle shows the proportion of montane species Locations: 1a all 6 subunits are well developed and contain no montane species (intermediate elevation with a moderately humid climate), 1b typical variety of the typical Arrhenatheretum is less well developed (drier climate); 2 only Salvia-Arrhenatheretum (relatively dry climate, free-­ draining soil). 3 only the first 4 units occur with montane species (Swabian Jura with free-draining soil); 4a Salvia-Arrhenatheretum is absent due to the humid climate (Black Forest and Odenwald), 5 uplands without Arrhenatheretum

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mown at most three times a year. Even the deepest loess loam soils with a high water storage capacity support species of dry conditions, e.g. on the Tauber Gäu Plateau and in the driest and warmest parts of the Upper Rhine Plain. Typical Arrhentheretum communities are completely absent here, and even the damper varieties are only found on a small scale. In the Swabian Jura, the typical subassociation is often found alongside the Salvia-Arrhentheretum, because the climate here is relatively cool and damp. In the higher parts of this flat-topped mountain range, the Arrhenatheretum contains several montane species (see Sect. 8.3.2.2), but this does not affect its differentiation into sub-communities based on soil moisture. Cirsium- and Carex-Arrhenatheretum communities are very rare in the Jura uplands due to the highly permeable carbonate substrate. The Salvia form and other dry types of Arrhenatheretum, which played and still play such a large role in southwestern Germany, are largely absent from the coolest parts of this region with the highest precipitation levels. At most, they can be found sporadically on steep, calcareous south-facing slopes, whilst other areas are dominated by the typical or damp forms. Even the Carex form does not occur, apparently because the character species are not competitive enough to dominate on constantly wet soil. Similarly to northwestern Central Europe, the Arrhenatheretum only occurs on mesic soils, and at higher elevations its character species are replaced by montane species. The Arrhenatheretum does not occur in the cool higher elevations of the Black Forest. Northwestern Central Europe  The species composition of the Arrhenatheretum as well as its subdivision according to the soil properties changes with the regional climate not only in southwestern Germany, but also in other parts of Central Europe. In particular, it loses characteristic species of the southern German communities towards the edges of its range, such as Bromus erectus, Campanula patula, Colchicum autumnale, Leontodon hispidus and Pastinaca sativa. Whilst at least three subassociations and numerous varieties are found in western Central Europe, only two subassociations can be distinguished in northwestern Germany and the Netherlands. These are namely the Briza-Arrhenatheretum and the Alopecurus-­ Arrhenatheretum, whilst forms with fewer (character) species should instead be assigned to the Ranunculus repens-Alopecurus pratensis association (Dierschke 1997b: see Fig. 8.16 top). There are so few differential species for these subassociations that they would at most be considered varieties in southwestern Germany. Tüxen (1937) lists Helictotrichon pubescens, Pimpinella saxifraga, Briza media, Plantago media and Luzula campestris for the dry Briza-Arrhenatheretum, which is now extremely rare in Germany. These species are more indicators of nutrient-poor rather than dry conditions. The Briza subassociation of the northwestern Arrhenatheretum thus corresponds to a species-poor form and moderately dry variety of the typical Arrhenatheretum of southwestern Germany. The only Mesobromion species to occur here is Ranunculus bulbosus (cf. B 1 with A 1in Table 8.7), although not in all

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Fig. 8.16  The classification of Arrhenatheretum and moist meadows according to increasing height of the average groundwater level in well-drained soil in northwestern (top) and southwestern (bottom) Germany (From Ellenberg 1954a) Top: Dry forms of the Arrhenatheretum are rare in the oceanic northwest Bottom: Dry forms are the most common form in the south, but there are also large areas of damp to wet forms

Briza-Arrhenatheretum stands. The Alopecurus-Arrhenatheretum of northern Germany can be compared to a moderately damp variety of the typical Arrhenatheretum (cf. B 2 with A 2). The Arrhenatheretum does not occur in areas of northwestern Central Europe with higher groundwater levels (see Fig. 8.16). Whilst all character species of the Arrhenatheretum together with numerous species of damp and wet habitats can still be found in the Donauried in western Bavaria (e.g. in columns A 3 and A 4  in Table 8.7), they mutually exclude each other in the Lower Rhine region (B 3 and B 4). Soils that would be colonised by a Cirsium-Arrhenatheretum in the southwest (A 3), support an Angelico-Cirsietum oleracei or other damp meadow community, e.g. a Bromus racemosus-Senecio aquaticus association (B 3) in the northwest. Only a few of the Arrhenatheretum character species spill over into such damp meadows. The wet soils on which tall sedges such as Carex acuta can grow are hostile to the other Arrhenatherion species in the northwest, so that there are no parallel communities to the Carex-Arrhenatheretum (A 4) here. Instead, damp meadows such as the typical Bromus racemosus-Senecio aquaticus association (B 4) develop.

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Why is the amplitude of soil moisture of the lowland Arrhenatheretum so much narrower in northern than in southern Central Europe? Ellenberg (1954a) believes that the drying of the topsoil may play a role in this phenomenon. In a relatively warm and continental climate, the water stored in the uppermost soil layer with its dense root networks is more frequently and strongly depleted. As a result, the topsoil does not remain saturated with water and hypoxic for long, even at very high groundwater levels, so that grasses with root systems that require well-aerated soils can spread. In a cooler and more humid oceanic climate, in contrast, only marsh plants can survive on soils with high groundwater levels where anoxia is frequent. More Contintental Regions  Like beech, the range of the Arrhenatheretum is mostly concentrated in subatlantic regions. In the oceanic climate of the outer Alps these meadows persist relatively far to the east. In Styria (Eggler 1958) and in the Karawanks mountain range (Aichinger 1933), for example, they are well developed with numerous subtypes, whilst they are completely absent from the inner Alpine valleys (see Braun-Blanquet 1961). North of the Alps, their range ends in central Germany. In the Unstrut valley near Straußfurt (Thuringia, central Germany), Hundt (1957) recorded clearly distinguishable dry, typical and damp subassociations of the Arrhenatheretum. He named these three sub-communities the ‘­ Bromus-­ Arrhenatheretum’, ‘pure Arrhenatheretum’ (see Fig. 8.12) and ‘Cirsium oleraceumArrhenatheretum’. By the middle reaches of the Elbe and lower Mulde (eastern Germany), the Arrhenatheretum is rarer and Hundt (1954) distinguished only ‘dry’ and ‘damp’ forms. Alopecurus pratensis is common in both subassociations, so is even less suitable to differentiate between the dry and damp types here than in southwestern Central Europe. The Ranunculus repens-Alopecurus pratensis community is characteristic for the periodically flooded river meadows in the northeastern Central European lowlands (Freitag and Körtge 1958; Hundt 1958). Arrhenatherum is replaced in this cooler climate with more frequent late frosts by Alopecurus, which also quickly colonises nutrient-rich but somewhat wetter habitats and is more cold-tolerant (Dierschke and Briemle 2002). Most of the soils that support the Ranunculus repens-Alopecurus pratensis community are used as permanent pasture, rotational pastures with mowing or arable fields in the northern German lowlands. Hay meadows are only found in places where grazing is problematic, i.e. very wet grasslands with sedge communities or on some dykes. The easternmost occurrences of the Arrhenatheretum are in northern Romania and northern Poland, where they are largely similar to the more western types (Passarge 1963b). In Belarus and Fennoscandia, but also in Slovakia, they are replaced by other meadow communities, e.g. those dominated by Deschampsia cespitosa (tufted hairgrass). In the Balkans, the best-developed Arrhenatheretum stands are found in Croatia, particularly in the surroundings of Zagreb, which of all the landscapes in the Balkans has a climate and vegetation most similar to Central Europe (see Horvat et al. 1974).

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Whilst the Arrhenatherion alliance does not reach the eastern and northern edges of Central Europe, it far exceeds the western and southern edges. In the southern Cévennes, i.e. the border to the submediterranean region, as well as in the Cantabrian Mountains, their species composition is largely similar to that of the Central European Arrhenatheretum (Klesczewski 2000). Overall, the Arrhenatheretum (or the Arrhenatherion alliance in the narrower sense) has a similar distribution to the beech-dominated forest communities. In contrast, no areas of natural spruce or pine forests have been found with well-developed Arrhenatheretum stands. Thus, together with the Fagion, the Mesobromion, the Nanocyperion (see Sect. 10.1.2) and several other alliances, the Arrhenatherion is one of the characteristic vegetation types of Central Europe. This is why the Arrhenatherum meadows and their regional varieties are described, as they were present a few decades ago, in such detail. The remaining grassland communities will be summarised more briefly. 8.3.2.2  Mountain Arrhenatherum and Trisetum Meadows 8.3.2.2.1  Changes in Arrhenatherum Meadows with Increasing Elevation The diversity of the Arrhenatherum meadows in Central and Southern Europe is further increased by the fact that they also change with increasing elevation. The shortening of the growing season, the reduction in summer temperatures, the increase in precipitation and thereby also leaching, as well as the reduction in use intensity have a negative effect on the competitiveness of Arrhenatherum elatius and its accompanying species (see Fig. 8.17). They are replaced by less demanding species that are suppressed by the taller, more competitive species in the lowlands, such as Trisetum flavescens and Poa chaixii, grasses of medium height and species of low-nutrient soils like Festuca rubra, Agrostis capillaris, Hypericum maculatum, and other low-growing species. In the more northerly mountains of Central Europe, e.g. on the Vogelsberg and in the Rhön Mountains in central Germany, Arrhenatherum occurs sporadically up to 600 m a.s.l., but its association is replaced by Trisetum meadows from 350–400 m upwards unless intensively fertilised (Speidel 1972). In fertilised grassland, the Alchemilla upland form of the mountain Arrhenatheretum (Dierschke 1997a; other authors describe this as a submontane Poo-Trisetetum) dominates up to around 500 m (and up to 1000 m in the southern Black Forest). Above this, it is replaced by the Geranio-Trisetetum (see below and Table  8.8). In the Alps, in contrast, Arrhenatherum can dominate in places up to 1200 m a.s.l., particularly on base-rich soils, and can occasionally reach above 1500 m. As other character species of the lowland Arrhenatheretum sometimes also reach the montane belt, Baeumer (1956) and Oberdorfer (1957 and later) also term these the mountain Arrhenatherum meadows (‘Arrhenatheretum montanum’, see Fig.  8.17). Theurillat (1992c) describes such intermediate types in Valais as Trisetum meadows, i.e. the Anthrisco-Trisetetum. Other mountain ranges have intermediate conditions, e.g. in the Polish western

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Fig. 8.17  The occurrence of meadow plants as a function of elevation in the Harz Mountains and the Thuringian Forest (in % of all stands surveyed at that elevation) (From data in Hundt 1966)

Table 8.8  Climatic data for fertilised mesic meadows at different elevations

Community Trisetetum  Rumici-T. (upper montane)

Altitude (m a.s.l.)

Precipitation (mm year−1)

900–1500

1240–1720

 Geranio-T. (montane)

600–1400

1100–1700

 Poo-T. (submontane)

2001–600

600–1050

1001–800

575–900

20–150

475–650

Arrhenatheretum  Alchemillo-A. (submontanemontane)  Dauco-A. (lowland-collin)

Mean air temperature (°C) (January) (year) –6.7 bis −3.4 –3.8 bis −2.0 –3.0 bis −0.4 –4.0 bis +1.7 –0.8 bis +0.3

2.7–5.7 5.2–6.3 6.3–7.0

5.5–8.7 8.6–9.4

From data in the literature from northern Poland to Belgium and from Slovakia to Switzerland; simplified from Passarge (1969a) 1 Lowest values in northeastern Poland; otherwise 400 (Poo-T.) and 350 (Alchemillo-A.)

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Fig. 8.18  Cross-section of a Trisetetum meadow in the lower Harz Mountains, which developed from a Nardus grassland after fertiliser application and two cuts per year (From Hundt 1964) From left to right: Nardus stricta, Ranunculus acris, Meum athamanticum, Trisetum flavescens (3 flowering stems), Lathyrus linifolius, Trollius europaeus, Hypericum maculatum, Festuca rubra, Hypericum perforatum, Heracleum sphondylium, Alchemilla vulgaris coll.; shown to 68 cm below the soil surface

Carpathians. These support montane mesic meadows up to around 500–600 m, which Kornas (1967) described as the Gladiolo-Agrostietum. In the Central Bohemian Uplands, the Arrhenatheretum occurs at somewhat higher elevations on south-facing slopes (see Fig. 8.24). 8.3.2.2.2  Montane and Subalpine Trisetum Meadows Typical Trisetum (golden oat grass) meadows (Geranio-Trisetetum, see Figs. 8.18, 8.19, 8.20) are only found in the upper montane belt, as described by Marschall (1947) in Switzerland. The transition from Arrhenatherum- to Trisetum-dominated communities increases in elevation from north to south from 400 to 800  m a.s.l. Most authors assign these meadows to their own alliance, the Polygono-Trisetion within the Arrhenatheretalia order (see also Dierschke 1981a). The character species of the Geranio-Trisetetum are: Alchemilla vulgaris agg. Centaurea pseudophrygia Crepis mollis Crocus albiflorus

Narcissus radiiflorus Phyteuma nigrum Pimpinella major subsp. rubra Viola tricolor subsp. subalpina

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Fig. 8.19  A Meo-Trisetetum meadow in the upper Thuringian Forest (Vesser valley, central Germany) at around 800 m a.s.l. with abundant Meum athamanticum and Geranium slyvaticum

Alongside these, the Trisetetum is characterised by numerous relatively light-­ demanding montane and subalpine species that differentiate it from the Arrhenatherion, namely: Primula elatior (common) Silene dioica (common) Astrantia major Campanula scheuchzeri Geranium sylvaticum

Meum athamanticum Phyteuma ovatum Phyteuma orbiculare Poa chaixii Rumex alpestris

Trisetum flavescens itself cannot be used as either character species or a differential species, because it is too common in the lowland Arrhenatheretum (see Fig. 8.17; see also Baeumer 1956). The role of Polygonum bistorta (bistort) in this community is uncertain. Its range is mainly boreal and montane, but it also occurs in lowland areas and can be found in fertilised mesic meadows particularly in northern Germany, and less so in other parts of Central Europe (Dierschke 1997a), but it is in steep decline here. It is less tightly linked to damp habitats in mountain areas as it is in the lowlands, but is largely absent in the dry types of the Trisetetum. Some species characteristic of the humid herb layer of lowland beech forests (e.g. Anemone nemorosa, Luzula luzuloides or Phyteuma spicatum), occur outside of the forest shade in montane meadows, presumably because evaporation rates are lower in the humid mountain climate.

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Fig. 8.20 Daffodils (Narcissus radiiflorus) in spring in a mesic Trisetum meadow below the Moleson in the western Swiss Prealps (Photo: Perrochet)

The Trisetetum also mixes with other grassland communities, especially as the relief often varies at small scales at higher elevations. With high levels of fertiliser application, the Trisetetum meadows are replaced by tall herb communities (see Sect. 8.3.4) with a larger proportion of forbs. Similarly to the Heracleum-­ Arrhenatheretum, the large-leaved forbs suppress the growth of the grasses and legumes through shading (see Fig. 8.12). Some mountain meadows are intensively fertilised because the harsh mountain climate means that livestock must be housed and fed for long periods (Dierschke and Briemle 2002). In mountains with silicate bedrock, e.g. in the Black Forest, unfertilised mountain meadows rich in Meum athamanticum (baldmoney) are found that are distinguished by some authors as the Meo-Festucetum. These also merge with Nardus grasslands with many intermediate types. The less intensive the cutting and the fertiliser application, the more Nardetum species replace the taller and more demanding species of the Trisetetum. Intermediate combinations of species are often dominated by Agrostis capillaris or Festuca nigrescens. Such intermediate communities can cover larger areas than the pure Nardetum or Trisetetum combined in some mountain regions. In limestone mountains, Trisetetum communities are often interspersed with Seslerion or related communities. However, as grassland is usually grazed here, combinations with pasture communities with large numbers of Leontodon and Crepis (see Sect. 8.3.6) are more frequent. Following Dierschke and Briemle (2002), several subtypes of damp mountain meadows can be distinguished, mainly caused by differences in nutrient supply and soil pH:

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(a) acid, nutrient-poor mountain meadows with e.g. Arnica montana, Deschampsia flexuosa, Galium harcynicum, Nardus stricta, Veronica officinalis; (b) moderately base-rich mountain meadows with e.g. Helictotrichon pratense, Pimpinella saxifraga, Primula veris, Sanguisorba minor; (c) nutrient-rich mountain meadows with e.g. Alopecurus pratensis, Anthriscus sylvestris, Festuca pratensis, Heracleum sphondylium; (d) damp mountain meadows fed by spring water with Cirsium palustre, Geum rivale, Filipendula ulmaria and other species of wet soils. In the low mountain meadows, a western German and Belgian Phyteuma nigrum type can be distinguished from an eastern German and Czech Centaurea pseudo­ phrygia type (Dierschke and Briemle 2002). Kliment (1994) and Chytrý (2007) give overviews of the mesophilic mountain meadows in Slovakia and Czech Republic, respectively. Compared to the Alpine Trisetum meadows (e.g. the Astrantio-Trisetetum of Oberdorfer 1957, 1983b), those of the low mountains are species-poorer. This is particularly shown in the literature review by Hundt (1964, 1966) of the meadows in the Thuringian Forest and the Harz Mountains (see also Hundt 1980; Bruelheide 1995; Waesch 2003; see Figs. 8.18 and 8.19). The Alpine Trisetetum contains not only species of the Poion alpinae alliance, but also those of the subalpine Nardus grasslands, e.g. Crepis aurea, Phleum alpinum, Plantago alpina, Polygonum viviparum and Rumex alpestris. 8.3.2.3  Unfertilised Grasslands of Intermediate Habitats In intermediate habitats, particularly those with mesic (moderately moist to moderately dry) soils, almost no unfertilised grassland remains in the modern cultural landscapes of Central Europe. Their high productivity means that these areas are usually under arable cultivation or forest, and where grassland exists, it is fertilised and used as intensively as possible. As a result, it is now almost impossible to study the species composition of unfertilised, low-intensity meadows on mesic and moderately base-rich soils. Probably it would be a mixture of species of damp litter meadows (Molinion) and semi-dry grasslands (Mesobromion), as well as of acid Nardus grasslands (Violion caninae), as is found here and there in the foothills of the Alps on dry hilltops in the regions with Molinia litter meadows. Studies of the vegetation in grasslands that have not been fertilised for decades and only occasionally mown or grazed, such as found on military training grounds or airstrips (Glavac and Raus 1982; Manz 1997), suggest that Arrhenatheretum communities would transform into relatively unproductive, nutrient-poor grasslands and meadows of the Festuca rubra-Agrostis capillaris community if fertiliser application ceased for a long period of time. Probably the taller grasses promoted by fertilisation would fail to grow, including Arrhenatherum, allowing the less demanding grasses such as Festuca rubra, Agrostis capillaris and Poa pratensis and low-

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growing herbs to spread. This grassland resembles the Poa pratensis facies of the Arrhenatheretum (see Fig. 8.14). It is also possible that the Festuca rubra-­Meum community (Meo-Festucetum) of low-intensity nutrient-poor montane meadows forms, as an unfertilised type of the Geranio-Trisetetum (Bartsch and Bartsch 1940; Dierschke and Briemle 2002).

8.3.3  Wet Meadows Between the Arrhenatheretum and the Trisetum on the one hand, and the tall- and small-sedge communities on the other, there are large areas with soils influenced by groundwater or trapped surface water that support the moist meadow communities (see Fig. 8.11). These productive meadow communities probably existed as early as the Bronze Age in river valleys where the floodplain and swamp forests had been cleared (Küster 1995), as these soils would have been too wet for grazing or crop cultivation. Moist meadows contain a range of vegetation types with a large range in soil moisture levels, acidity and nutrient regimes. Naturally mesotrophic to eutrophic or fertilised areas support Calthion and Cnidion communities, whilst nutrient-­ poorer moist meadows that are only mown late in autumn support Molinion communities. The Filipendulion alliance of tall herbs contains forb-­ rich and rarely mown moist meadow communities. 8.3.3.1  Wet Meadows of Mesotrophic to Eutrophic Habitats 8.3.3.1.1  Calthion Communities and Other Fodder Meadows of Base-Rich Wet Soils The wet and nutrient-rich meadows are placed within the Calthion palustris alliance, although the name is somewhat misleading as the marsh marigold has a relatively low fidelity to these communities. Indeed, this alliance otherwise contains few character species that are reliably present throughout (at most those listed in Table  8.7). The Calthion is best characterised by the absence of species of the Arrhenatherion or Trisetion on the one hand, and of species of the Molinion (see Sect. 8.3.3.2) on the other (see also Burkart et al. 2004). Many of its communities are characterised by remnants of sedge communities, with which they often form a mosaic of habitats (see also Chap. 4). Together with a basic set of generally widespread meadow plants (i.e. the class character species, see Table  8.7), the Calthion communities contain numerous helophytes, particularly character species of the Molinietalia order also listed in Table 8.7. In contrast to the small-sedge communities, the soils here are constantly moist, but never flooded (i.e. anoxic) for long periods. Depending on the base-­ richness of the soil and the drainage, as well as the climatic conditions, various association character species appear. The management also has a large influence on

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the species composition, making the Calthion alliance particularly rich in communities (Havlová et al. 2004; Chytrý 2007). The Angelico-Cirsietum oleracei is widespread in Central Europe (see Meisel 1969b; see Figs. 8.21 and 8.22). It mainly occurs on relatively base-rich wet mineral soils or drained peat soils with groundwater levels 30–60  cm (at most 140 cm) below the surface. Two cuts per year and occasional fertiliser application produces relatively grass-rich stands, especially with Alopecurus pratensis, Festuca pratensis, Holcus lanatus and Poa trivialis. This community has similar nutrient, and particularly nitrogen, levels to the Arrhenatheretum (Williams 1968). Even wetter meadows, e.g. the Caricetum gracilis, are always poor in nitrogen due to their relatively high denitrification rates (Kovacs 1964; León 1968). This is also often the case in the wettest form of the Angelico-Cirsietum, the Angelico-­ Cirsietum caricetosum, in which the groundwater is usually just under the soil surface. Phalaris arundinacea and Glyceria maxima and in some places also Carex disticha can dominate in these meadows. In contrast, the Angelico-Cirsietum heracleetosum tends more towards the Arrhenatheretum, and also produces the highest yields of all the wet meadow communities. With increasing elevation, the Angelico-Cirsietum oleracei is replaced on calcareous soils in eastern and southern Central Europe by the Cirsietum rivularis (as described by Oberdorfer et al. 1983b; see also Chytrý 2007): in southern Germany this occurs above around 700 m. This temperate-continental community can also be

Fig. 8.21  Cross-section of a wet Cirsium oleraceum-­ Polygonum bistorta association, Carex acutiformis subassociation, on glacial sands south of the Fläming uplands (eastern Germany) (From Hundt 1958) From left to right: Cirsium oleraceum, Carex acutiformis, Cirsium palustre, Lotus uliginosus, Holcus lanatus, Galium palustre, Filipendula ulmaria (leaf), C. o., H. I., C. a., Geum rivale (leaves), Angelica sylvestris, Ranunculus acris, Poa trivialis. Shown to 70 cm below the soil surface

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Fig. 8.22  A moderately fertilised Angelico-­ Cirsietum oleracei near Clausthal-Zellerfeld (Harz Mountains, northern Germany) at around 500 m a.s.l. with Cirsium oleraceum, Bistorta officinalis and Cirsium palustre in the foreground

divided into multiple subtypes, and borders with many other grassland communities. In eastern and southern Central Europe, Deschampsia cespitosa plays such a major role in wet hay meadows that it is included in the name of the Calthion communities there, e.g. in the Stellario-Deschampsietum near Warsaw (according to Traczyk 1968) and in the Deschampsietum or Cirsietum rivularis deschampsietosum (according to Spanikova 1971; see also Fig.  8.25). This dominance of Deschampsia is at least partly due to the influence of the livestock that graze the meadows after mowing and avoid eating this tough grass. Traczyk (1968) found that the proportion of dead plant material in wet Deschampsia meadows is unusually high compared to other meadows, mainly produced not only by the dominant grass but also sedges such as Carex nigra and C. panicea. Under higher fertiliser application and more frequent mowing, these meadows turn into Angelico-Cirsietum oleracei stands even in central and southern Poland. A wide variety of transitional types between the Deschampsietum and the Angelico-Cirsietum oleracei can be found even today e.g. in the Spreewald (Müller-Stoll et al. 1992).

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Fig. 8.23  Cross-section of a Trollius europaeus-Polygonum bistorta meadow in the southeastern Harz Mountains, here in its mainly vegetative early summer state; grasses come to prominence later (From Hundt 1964) From left to right: Trollius, Polygonum, Filipendula ulmaria, Trisetum flavescens (culm with buds), Anthriscus sylvestris, Alchemilla vulgaris coll., Cirsium palustre, Heracleum sphondylium. Shown to 70 cm below the soil surface

Fig. 8.24  Distribution of meadow communities according to elevation and aspect in the Bohemian Highlands (Modified from Moravec 1965). The Arrhenatheretum, Trisetetum and Agrostietum communities are similarly distributed in the mountains of western Central Europe, although here the lower limits of the montane meadows are often at higher elevations

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8.3.3.1.2  Wet Meadows on Base-Poor Soils Fertilised meadows of base-poor soils are found in such a multitude of forms, both in silicate mountains and in the sandy lowlands, so that we can only mention a few examples here. The majority of these have only developed in the last 100 years from Molinia litter meadows, whilst others were created by the drainage of sedge communities. Similar meadow-like forb and sedge communities can naturally form only in a few small areas, e.g. around springs (see Sect. 4.3.4). Some of the acidic wet meadows can be extremely species-poor. The Bromo-­ Senecionetum aquaticae, which was widespread in the lowlands of northwestern Germany from the Dutch border to Schleswig-Holstein, has only the single character species of Senecio aquaticus (Tüxen and Preising 1951). As noted by Meisel (1969b), the Bromus racemosus referred to in the name of this community does not actually occur in true Bromo-Senecionetum stands. They are in some respects similar to the Deschampsietum, and like these, are intermediate between the Angelico-­ Cirsietum oleracei and the Molinietum, although the boundaries between the three are often blurred. They can, however, easily be classified into a dry, a typical and a wet subassociation (see Burkart et al. 2004). As can be seen in Table 8.7, these types differ on the same soils based on the distance of the water table from the surface. The fluctuations in soil moisture levels over the year and with different weather conditions are quite considerable in all sub-communities. The once common Bromo-Senecionetum aquaticae has been in decline in northwestern Germany over the last 50 years due to increasing fertilisation and mowing frequency (Schrautzer and Wiebe 1993), and will soon be restricted in this region to nature reserves (Preising et  al. 1997). In the valleys of the acid, subatlantic low mountains, the Bromo-Senecionetum occurs in small fragments with a somewhat different species composition (Bergmeier et al. 1984; Schwabe 1987).

Fig. 8.25  The distribution of meadow communities in a floodplain in western Czech Republic under low-intensity grazing (left) and under regular fertilisation and mowing after drainage (right) (Modified from Moravec 1965) Almost all communities also occur (or occurred) in southwestern Central Europe under similar environmental conditions. Instead of the Sanguisorbo-Deschampsietum, a Polygono-Cirsietum or damp Arrhenatheretum would dominate here

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8.3.3.1.3  River Valley Meadows with Low-Intensity Management The communities of the Cnidion alliance in the east of Central Europe are exposed to greater fluctuations in the water level than the other wet meadows in the Molinietalia. In contrast to the artificially fertilised Calthion meadows, these are species-rich ‘naturally’ fertilised forms. The most influential factor for these meadows is the sediment-rich floodwater that delivers nutrients and moisture several times each year. The characteristic species of this alliance listed by Balátová-­ Tulácková (1969) are mainly found in continental climates, namely: Cnidium dubium (C 6) Allium angulosum (C 7) Carex praecox var. suzae (C 7) Gratiola officinalis (C 5)

Juncus atratus (C 8) Leucojum aestivum (C 4) Lythrum virgatum (C 8?) Oenanthe silaifolia (C 5)

Burkart et  al. (2004) list the differential species in Germany as being, among others, Elymus repens, Poa palustris, Potentilla reptans and Symphytum officinale. As these meadows are often mown late due to the flooding, they also have a certain floristic similarity to the Molinion (see Sect. 8.3.3.2.1). Balátová-Tulácková (1969) lists several associations for eastern Germany, southeastern Czech Republic, Austria and northeastern Croatia (see also Chytrý 2007 for the Czech Republic). The Cnidium dubium meadows of eastern Germany, the Czech Republic and Poland are combined by Burkart (1998) and Chytrý (2007) into the Cnidio-Deschampsietum. In the valley of the middle Elbe and the middle Upper Rhine, wet depressions are colonised by the Cnidio-Violetum with character species such as Viola persicifolia, Cnidium dubium, Lathyrus palustris and Gratiola officinalis (Philippi 1960; Walther 1977). The Sanguisorba officinalis-Silaum silaus community occurs in river valleys with moderate nutrient levels in southern and central Germany, but in contrast to the Cnidio-Deschampsietum has a more subatlantic range (Thomas 1990; Dierschke and Briemle 2002; Hundt 2007). In terms of its nutrient supply, the Cnidion is intermediate between the fertilised Calthion and nutrient-poor Molinion. All Cnidion communities are, like the Molinietum communities, sensitive to fertilisation, which promotes the more productive species of the Calthion. 8.3.3.2  Wet Meadows of Oligotrophic to Mesotrophic Habitats 8.3.3.2.1  Molinia Litter Meadows Together with the Festuco-Brometea class, Molinion meadows are among the most species-rich and most colourful grassland ecosystems of Central Europe, with up to 80 species per 20 m2 (see Fig. 8.26). They colonise damp soil with relatively low nitrogen availability. In contrast to the Calthion, the Molinion character

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Fig. 8.26  A somewhat schematic cross-section through a calcareous Molinia grassland on the Swiss Plateau (From M. Mayer 1939) From left to right: Serratula tinctoria, Carex panicea, Molinia caerulea, Selinum carvifolia, Potentilla erecta, Gentiana pneumonanthe, M.c., Thalictrum flavum, M.c., Sanguisorba officinalis (leaf), Succisa pratensis, Allium angulosum, Epipactis palustris, Iris sibirica, M.c. Shown to 10 cm below the soil surface; the roots are only indicated, but would really reach several tens of centimetres down

species generally have low leaf concentrations of N, P and K (Balátová-Tulácková 1993). Their habitats are intermediate between nutrient-poor mesic meadows and small-sedge communites. Their development is linked to a single cut per year or even every 2 years without fertiliser application. The eponymous Molinia caerulea (purple moor grass) and its taller and larger-flowered sister species M. arundinacea are widespread in these communities, but not limited to them. They also occur in fens, swamp forests, damp heaths and Nardus grasslands. Molinietum communities are sensitive to fertilisation, which causes the species richness to decline and most of the character species to disappear. Dierschke and Briemle (2002) noted that some of these meadows formed on nutrient-poor gley and peat soils early in the history of low-intensity mowing. However, most developed less than 200 years ago in the Prealps, where litter meadows were once the most widespread grassland types (Konold and Hackel 1990). When year-round housing and feeding of livestock became widespread at the beginning of the nineteenth century, litter raking in the forest threatened to cause serious damage to forest production. In response, many wet meadows were transformed into litter meadows, especially as grassland litter produced more nutrient-rich manure than forest litter. It was only in the twentieth century that cereal straw was increasingly used for bedding, which due to the mineral fertilisers became available

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in larger quantities. As a result, the litter meadows of Central Europe gradually declined, so that they are now almost exclusively found in the foothills and edges of the Alps. Even here they are becoming rarer, as modern livestock systems house the animals without as much bedding. Molinietum communities are usually only mown towards the end of the growing season, when the plants have died back and become straw-like. Many species are adapted to autumn mowing and flower late, such as Serratula tinctoria, Selinum carvifolia, Succisa pratensis etc. In some regions (e.g. in the southern Black Forest and in Hesse), there are also Molinietum communities that are mown once to twice a year for fodder (Bergmeier 1990; Nowak and Schulz 2002) linked to the once widespread practice of meadow irrigation (see Sect. 8.7.2). Molinia-rich wet meadows are found on base-rich as well as acid soils, as Molinia species are largely indifferent to soil pH and base-richness (Balátová-­ Tulácková 1968, 1972). Grabherr (1942b) states that Molinia is restricted to very humus-rich to peaty soils. It is promoted by fertilisation, but soon overgrown by other, faster-growing grasses, so that both Molinia caerulea and arundinacea only achieve high coverages in nitrogen-poor habitats. They are sensitive to early mowing because they only translocate nutrients to their storage organs relatively late in summer, and to long periods of flooding. Molinietum communities can be divided into those dominated by calcicole and those dominated by calcifuge species (Tüxen and Preising 1951). Only the former, the Molinietum on calcareous soil (or ‘true’ Molinietum) is species-rich (see Table  8.9), with alliance character species such as Serratula tinctoria, Cirsium tuberosum, Dianthus superbus, Betonica officinalis, Silaum silaus and Carex tomentosa. These communities include the ‘pure’ form of the Molinietum caeruleae, which was once widespread on base-rich fen soils mainly in the surroundings of Lake Constance and southern Bavaria (see Figs. 8.26 and 8.27), but reached up to the northern low mountains in its diverse forms (Philippi 1960; Buchwald 1996; Nowak and Schulz 2002). Among the numerous associations of the calcareous Molinietum, other major ones include the Cirsio tuberosi-Molinietum arundinaceae and the Allio suaveolentis-Molinietum with a mainly Prealpine distribution (cf. Oberdorfer 1983b; Pott 1995), and the Succiso-Molinietum in the Pannonian region (Wagner 1950). Acid Molinietum communities developed in place of birch swamp forests and damp birch-oak forests. They are comparatively species-poor and usually take the form of the Junco-Molinietum or Juncus conglomeratus-Succisa pratensis ­community. As they lack many or even all of the character species of the Molinion alliance, their synsystematic position is contested. Some stands are best assigned to the Calthion (Burkart et al. 2004), and some with species of nutrient-poor habitats like Nardus or Danthonia have more in common with the Violion caninae (Nowak and Schulz 2002). Both calcareous and acid Molinietum communities contain a number of subassociations that reflect the soil moisture levels. The typical subassociations (e.g. A 2 and B 2 in Table 8.9) do not contain any differential species. The relatively wet sub-­ communities (A 1 and B 1) sometimes contain the same marsh species on both

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Table 8.9  Calcareous and acid Molinia meadows on wet to dry soils Region: O

O All All A All

O All All

A A O All O

O

A = Upper Rhine valley B = Spreewald   No.: Silaum silaus Sanguisorba officinalis Centaurea jacea Selinum carvifolia Betonica officinalis Galium boreale Cirsium tuberosum Serratula tinctoria Galium verum Lathyrus pratensis Lotus corniculatus Carex tomentosa Inula salicina Tetragonolobus maritimus Dactylis glomerata Ononis spinosa Dianthus superbus Filipendula vulgaris Juncus subnodulosus Equisetum palustre Carex flacca Allium angulosum Molinia caerulea Succisa pratensis Deschampsia cespitosa Potentilla erecta Danthonia decumbens Poa pratensis Anthoxanthum odoratum Holcus lanatus Plantago lanceolata Festuca rubra Briza media Agostis stolonifera Ranunculus acris Prunella vulgaris Leontodon hispidus Filipendula ulmaria Linum catharticum Vicia cracca Lysimachia vulgaris Symphytum officinale

A 1 5 5 5 4 4 4 4 4 4 4 3 2 2 2 2 2 3 2 2 1 2 5 5 5 4 2 1 1 3 2 3 4 5 3 3 2 4 1 2 5 3

2 5 5 4 5 4 4 3 3 2 3 3 3 4 1 1 1 1 1 3 2 1 5 4 3 3 3 3 2 3 2 4 2 2

1 2 1 3 3

3 5 4 5 4 4 4 5 4 4 3 5 4 2 5 4 4 1 3 2 2 3 1 5 3 3 4 3 2 2 3 4

B 1 2

5 5 4 3 3 4 1

1 5 4 5 3 3 5 3 3

3 2 1 3 3 4 1 3

5 5 5 5 4 5 4 5 4 4 2 2 5 4 4 5 2 1 1

3

5 5 4 5 5 4 1 5 5 2 3 3 2 1 2 2 1

(continued)

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Table 8.9 (continued) Region:

A = Upper Rhine valley B = Spreewald   No.: Symphytum officinale Achillea millefolium O Achillea ptarmica Thalictrum flavum Valeriana dioica All Gentiana pneumonanthe Cerastium fontanum holost. B? Viola persicifolia O Galium uliginosum Luzula multiflora B Ophioglossum vulgatum Rumex acetosa B Salix repens Cardamine pratensis O Cirsium palustre Potentilla anserina Taraxacum officinale B Inula britannica Leontodon saxatilis Lotus uliginosus Carex pallescens Campanula patula Indicators of moisture:a Mentha aquatica Caltha palustris Lythrum salicaria Phragmites australis Carex panicea Carex acutiformis Hydrocotyle vulgaris Galium palustre Potentilla palustris Ranunculus flammula Phalaris arundinacea Iris pseudacorus Ranunculus repens Indicators of dryness and nutrient deficiency:b Bromus erectus Koeleria pyramidata Brachypodium pinnatum Plantago media Trifolium montanum

A 1 3 2 1 2 1 2

2 3

3 3

1 1

2 4 1 1 1 5 4 2 1 1 3 4 3 2 3 1 1 1

1

3 2 3 4 4 1

B 1 2 3

4

1 3 1 1

5 4 1 1 5 4 3 5 4 4 4 4 5

4 1 4 1 2 5 5 5 4 5 2 2 2 3 2 1 1 4 2 2

3 1 3 5 1 2 4 4 5 3 5 1 1 2 2 1 2 2 5 3

1 4 3

5 1 1

3

5 5 5 3 3 (continued)

8  Agricultural Grassland on Mesic to Wet Soils

654 Table 8.9 (continued) Region:

A = Upper Rhine valley B = Spreewald   No.: Pimpinella saxifraga Festuca ovina coll. Nardus stricta Polygala vulgaris Hypochoeris radicata Agrostis capillaris Dianthus deltoides Viola canina Calluna vulgaris Hypericum perforatum Carex pilulifera Armeria maritima

A 1 2

3 3 4

B 1 2 2 4 3 3

3 2 5 4 3 5 5 5 5 5 3 3

From relevés by Philippi (1960) and Passarge (1956b) In no. B 1 also with 3: Juncus effusus, Carex vesicaria, Lycopus europaeus and Stellaria palustris b In no. A 3 also with 3: Phyteuma tenerum Species that do not have a constancy greater than 2 in any vegetation type were mostly omitted Letters before the species names: A = Character species of vegetation type A B = Character species of vegetation type B All = Alliance character species of the Molinion O = Order character species of the Molinietalia A: Calcareous Molinia meadows (Molinietum medioeuropaeum) in the southern Upper Rhine Plain, from Philippi (1960, Table 1) No. 1: wet variety of the typical subassociation No. 2: typical variety of the typical subassociation No. 3: moderately dry (Brachypodium) variety of the subassociation of Bromus erectus B: Acid Molinia meadows (Viola persicifolia-Molinia caerulea association) in the Lübbenauer Spreewald, from Passarge (1956b, Table XII) No. 1: typical variety of the subassociation of Potentilla palustris (wet) No. 2: typical variety of the typical subassociation No. 3: moderately dry (Calluna) variety of the subassociation of Dianthus deltoides a

calcareous and acid substrates. Unsurprisingly, it is the dry subassociations (A 3 and B 3) that differ most. The dry calcareous Molinietum shares some species with calcareous Mesobrometum communities, and is particularly species-rich. The lists in Table 8.9 are extreme examples of calcareous and acid Molinietum communities, but both types can intermingle. Wagner (1950) noted that the Molinietum communities cannot be classified as clearly throughout Central Europe as Koch (1926) did in his first description of this vegetation type in northeastern Switzerland. This is particularly shown in the small-scale studies e.g. of Korneck (1962/63), Rodi (1963), Fritsch (1962), Braun (1983) and others.

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Fig. 8.27  A Molinietum litter meadow on wet, fairly base-rich soil in southern Bavaria (Murnauer Moos) at the foot of the Limestone Alps with Cirsium tuberosum, Allium suaveolens, Laserpitium prutenicum and other endangered species of the calcareous Molinietum

8.3.3.2.2  Juncetum and Scirpetum Communities of Grassland Springs Areas surrounding springs that are constantly wet are often visibly darker green than the surrounding wet meadow. These are dominated by rush species, which in calcareous areas is mainly Juncus subnodulosus and in acid areas Juncus acutiflorus. The Juncetum subnodulosi and the Crepido-Juncetum acutiflori used to be assigned to the small-sedge communities and considered to be natural grassland. However, Oberdorfer (1957 and later) noted that these are usually mown grasslands that are floristically more similar to the Molinietalia, and accordingly assigned the Juncetum subnodulosi to the Calthion alliance. This community is still quite common on marl soils in the foothills of the Alps but rare elsewhere. Whilst it only rarely contains character species of the Molinietalia, it always contains many species of the Caricetum davallianae (see Sect. 3.3.5). Wet meadows with Juncus acutiflorus (Crepido-Juncetum acutiflori) also belong to the Calthion (Burkart et al. 2004; Chytrý 2007, but cf. Oberdorfer 1957). They are widespread in the suboceanic low mountains on base-poor bedrock and into the northern German lowlands, although usually only in small areas. Both Juncetum communities degenerate very quickly if no longer regularly cut, in that the Juncus species spread and suppress the other meadow plants. The same is true for the often very species-poor Scirpetum sylvatici of spring areas. Their appearance is quite like that of the Magnocaricion, but their flora is more similar

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Fig. 8.28  A somewhat schematic cross-section through a Filipendula ulmaria riverbank community on the Swiss Plateau (From M. Mayer 1939) From left to right: Carex acuta, Phalaris arundinacea (top cut off), Carex acutiformis, Geranium palustre, Equisetum palustre, Filipendula ulmaria with Calystegia sepium (liana), Caltha palustris, Galium mollugo, Colchicum autumnale, G. m., Car. ac. Shown to 15 cm below the soil surface, roots largely omitted

to abandoned Calthion communities. They indicate lime-poor but relatively nutrient-­rich water (Yerli 1970; Balátová-Tulácková et al. 1977; Amani 1980; see Fig. 8.25).

8.3.4  F  ilipendula Riverbank and Similar Tall Forb Communities Some species of Molinion and Calthion meadows, namely large-leaved and tall-­ growing forbs such as Filipendula ulmaria, Lysimachia vulgaris, Lythrum salicaria or Thalictrum flavum, resemble the nitrophilic species of riverbanks, scrub fringes and the herb layers of forest clearings, and grow particularly well when nutrient supply is high. They are therefore more common in fertilised wet Calthion meadows than in Molinion meadows (see Fig.  8.28 and Sect. 8.6.2.2). These stands, characterised by the presence of Filipendula ulmaria (meadowsweet), are usually classed as their own alliance (Filipendulion) or even their own class, separate from the Calthion, Molinion and Cnidion (Preising et al. 1997; Burkart et al. 2004). Other authors (e.g. Chytrý 2007) place them in the Calthion.

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The edges of small streams or ditches are particularly nutrient-rich, as they are repeatedly (unintentionally) ‘fertilised’ by the mud deposited from dredging. Filipendula dominates here, as long as it is only rarely mown. The species composition of these communities (either the Filipendulo-Geranietum or the Valeriano-­ Filipenduletum) is similar to that of wet meadows, but these tall forb stands are better assigned to their own alliance, namely the Filipendulion (see Meisel 1969b), especially as they are not real meadows and lose their character if mown too frequently. Structurally, their tall growth resembles that of the mountain tall forb communities (see Fig. 5.41). Balátová-Tulácková (1979a) distinguished several Filipendula communities for the western Czech Republic depending on elevation and other habitat factors (see also Dierschke 1996b). The current range of the Filipendulo-Geranietum is the result of the meadow management of earlier centuries. However, it is possible that Filipendula streambank communities existed even before the influence of humans in Central Europe, for example along streams in alder-ash forests in the low mountains (e.g. as the Geranio sylvatici-Chaerophylletum hirsuti). Similar, sometimes almost natural communities have also been observed in southern Fennoscandia. In the east of Central Europe, the Filipendulo-Geranietum is replaced by other communities, in which alongside Filipendula e.g. Veronica longifolia and Scutellaria hastifolia play a role (e.g. Chytrý 2007). They are restricted to river valleys, but colonise stream and ditch edges as well as riverbanks. From here they can spread into the meadows of the Cnidion alliance, similarly to the way the species of the Filipendulo-­ Geranietum spread into Molinion and Calthion meadows that are no longer regularly mown. The Chaerophyllo-Ranunculetum aconitifolii is a montane to upper montane stream bank community of the Alps and low mountains rich in tall forbs. Abandoned wet meadows in many regions develop into large Filipendula ulmaria stands, as mentioned in the previous section (see Dierschke and Briemle 2002). These never develop the characteristic species combinations of the Filipendulion communities, but are rather transitional phases in the succession towards the natural forest vegetation (Falinska 1995). It often takes a very long time for the first willows, alders or other pioneer trees to establish in the shady Filipendula community, but a fire, a rootling wild boar or some other type of disturbance will sooner or later create an open area of soil where woody plants can establish. Filipendula communities and other tall forb stands are best maintained along small-scale structures such as drainage ditches in agricultural landscapes, as studied e.g. by Ruthsatz (1983) in the region around Ingolstadt (Bavaria).

8.3.5  Grass Verges and Traditional Orchards 8.3.5.1  The Development and Environmental Conditions of Grass Verges Habitat Types  Some species of nutrient-poor and low-intensity agricultural grasslands have found refuge in the grass verges of roads and motorways over the course of the intensification of agricultural landscapes. Arrhenatherum elatius, for exam-

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ple, is more commonly found in these habitats now than in agricultural grassland. It has even used them to spread to parts of Central Europe, such as the sandy lowlands of northwestern Germany, in which it was rare or absent until around 50 years ago. Numerous new plant communities have developed along roads and railways over the last 200 years (Brandes 1983, 1988a). Grass verges in the modern sense did not exist until the eighteenth century. Usually roads were unpaved and relatively wide, and also served to drive animals, and at most a narrow strip would have been cobbled. Under the influence of Napoleon, from 1813 onwards the edges of important connecting roads were planted with fast-growing poplars, of which almost all have since disappeared. The Prussian rulers also ordered treed avenues to be built to provide shade, with lime or pedunculate oak, and depending on the soil type, also apple trees or, rarely, damson and sweet cherry trees (see Wagner 1968). Nutrient-poor sands were planted with silver birch and, in cool upland areas, sycamore or rowan. Ash was often used in stormy coastal marsh areas. Apart from the fruit trees, these were therefore mostly native tree species typical of the natural landscape. There was very little grassland accompanying these roads, apart from a few trampled areas or the vegetation of the drainage ditches. Some venerable remnants of these avenues remain in eastern Germany, whilst they have largely been cleared in the west in deference to the demands of modern traffic. Roads for motor vehicles developed only in the 1930s. The verges were first sown not with commercial seed mixtures, but with material collected from meadows, or the sweepings from barns where hay was stored. This allowed many widespread wildflowers from the surrounding landscape to establish from the start in these verges. After 1950, the landscaping of road edges became increasingly mechanised, using standard seed mixtures and woody plants raised in nurseries. This caused not only a reduction in species richness and a homogenisation of verge areas, but also the introduction of non-native or not regionally adapted species. The widening of old motorways and major roads has meant that the once semi-natural verges have mostly been lost. The modern grass verge or motorway embankment communities are rarely more than 30 to 40 years old. Usually they are divided into three zones: 1. The shoulder directly adjacent to the road is strongly influenced by the traffic and most affected by pollutants and salt. It experiences periods of wet as well as periods of very dry conditions, and is generally mown two to four times a year. 2. The ditch (not always present) is mown once to twice a year with removal of the cuttings to avoid it getting clogged. The soil moisture is quite variable depending on the relief, so can differ considerably from zones 1 and 3 despite being spatially very close. 3. The embankment buffers the surroundings from the road, and is managed as little as possible, i.e. rarely cut or thinned, if it is planted with trees or shrubs. Depending on its width, the effects of the road may be quite minimal, whilst from the other side it may be influenced by pesticide and fertiliser run-off from neighbouring farmland. The effects of natural environmental factors are often visible here.

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The shoulder is always species-poor and contains species of trampled ground (see Sect. 8.3.7.1), as well as those of highly disturbed habitats (see Chap. 11), particularly if they are short-lived. The two other zones have more diverse vegetation, but still mainly with species of anthropogenic grasslands, heaths and/or tall forb and scrub communities. As large amounts of nitrogen compounds are emitted by the traffic, nutrient-demanding plants are frequent despite the lack of other sources of nutrient input. Species Composition  Grass-rich verge communities share many character and differential species with the Arrhenatheretum, but always also contain species that are much rarer in ‘true’ Arrhenatheretum communities. These are mainly character species of nutrient-rich forb communities (Schaffers 2000a), in particular of tall forb fringes of scrub patches, namely: Artemisia vulgaris Cirsium vulgare

Torilis japonica Lapsana communis

They also may contain arable weeds, e.g. the also highly nitrogen-demanding or indifferent species Elymus repens Cirsium arvense

Myosotis arvensis Vicia hirsuta

The verge communities are also usually scattered with shrub species that spread vegetatively, particularly Prunus spinosa and Rubus species. Stottele (1994) found that the largest difference between the grass verge Arrhenatheretum and the traditional Arrhenatheretum meadows was the general lack of meadow species, such as: Alchemilla vulgaris agg. Helictotrichon pubescens Crepis biennis Rhinanthus minor

Tragopogon pratensis Pimpinella major Carum carvi

The roadside vegetation also shares some species with dwarf shrub heaths and nutrient-poor grasslands. Scrub and clusters of trees are often present in the verge communities, so that these are often richer in tall forbs and woody plants of woodland habitats, and the N emissions allow a larger number of nitrogen-demanding species to persist. On the other hand, the verge communities lack the low-growing herbs of nutrient-poor grasslands that are promoted by grazing, and particularly those that used to be dispersed by sheep, i.e. species that are specific to the management type. Berg (1993), Stottele (1994), Szwed and Sykora (1996) and Schaffers and Sykora (2002) state that the roadside flora contains a considerable proportion of the local flora of the adjacent habitats. In structurally diverse areas of Germany, i.e. with a high heterogeneity of use types and habitat conditions, this was 39–53 % of the local flora. In intensively managed and homogenous landscapes with acid soils, this

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was even higher (46–62 %) (Stottele 1994). However, the proportion of threatened species here is extremely small. Verge communities are therefore only of limited use as refuges for rare species of the rapidly declining Calluna heaths, Nardus grasslands, Meso- and Xerobrometum and other communities of low-intensity farming practices (Tikka et al. 2000). The spread of diaspores by cars has a large influence on the composition of the roadside vegetation. Schmidt (1989) collected germinable seeds of 124 species in the dirt on the wheels and chassis of a single car over the course of 20 months. The most frequent of these were Poa annua, Plantago major, Epilobium roseum, Stellaria media and Poa trivialis. With the exception of Epilobium roseum, all of these species are common in grass verges. Halophyte species are particularly well adapted to the roadside conditions created by salting roads in cold weather, e.g. Puccinellia distans, Armeria maritima and Spergularia salina (see Breckle and Scheck-Hoffmann 1985 and Heinrich 1984). Some neophytes also profit from the roadside conditions (see Fukarek 1987), although they are dispersed more via the railway network than via roads in Central Europe (Brandes 1991b; Brandes and Griese 1991). Particularly noticeable in summer along motorways are the yellow flowers of the South African Senecio inaequidens. The phytosociological classification of roadside vegetation is reasonably difficult, as it often contains none of the diagnostically important character species of other vegetation units. Many fertilised meadow-like stands can be related to the ruderal Artemisia vulgaris-Arrhenatherum community. Some authors have developed their own solution to the problem, for example Brandes (1988a), Ullmann et al. (1988, 1990), Rattay-Prade (1988), Kopecky and Hejny (1978), Sykora et al. (1993) and Szwed and Sykora (1996), or describe the communities with a simple German or English name (Stottele 1994; cf. also Ross 1986). As the majority of stands are still relatively young, repeated surveys over long time periods are particularly interesting. In the surroundings of Halle/Saale (eastern Germany), Mahn (1957) and Berg (1990, 1993) surveyed Arrhenatheretum-like communities on largely the same areas with a gap of 30 years. Changes in species composition were mainly caused by increased nutrient deposition, particularly of nitrogen. Less nutrient-­demanding species such as Festuca rupicola (N 2), Picris hieracioides (N 4), Cichorium intybus (N 5) and Potentilla reptans (N 5) were largely replaced by ruderal plants like Urtica dioica (N 9), Artemisia vulgaris (N 8–7) and Cirsium arvense (N 7). Elymus repens (N 7) and Convolvulus arvensis (Nx) also increased in frequency, thus a ‘ruderalisation’ of the community occurred. Mederake (1991) gives an overview of roadside vegetation types under different management. The vegetation along railway networks and around stations is much less meadow-like than that of roadside verges, as it is almost never mown but instead treated with herbicides (Brandes 1983; Wittig 2002). They have some similarities to ruderal communities (see Chap. 11) and arable weed communities (see Chap. 12), but the railway flora is rarely important for plant conservation.

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8.3.5.2  Grassland in Traditional Orchards Village orchards were and are still a characterstic element of cultural landscapes in many regions of Central Europe (Weller 1992). They were used for both fruit production and hay cutting, and more rarely for grazing. The vegetation under the semi-­ shade of the trees is usually a relatively species-rich damp Arrhenatheretalia grassland, and in some places also semi-dry grassland or Nardus grassland that is mown two to four times a year or continuously grazed. Because parts of the grassland are shaded or profit from the water dripping from the edges of the fruit tree crown, small-scale mosaics of different grassland communities often form. Knapp (1971) found that Anthriscus sylvestris (cow parsley) stands are characteristic of the damp area of drip-water, while stands with Aegopodium podagraria (ground elder) and geophytes such as Ranunculus ficaria and Gagea lutea are generally found in the shadier areas. Studies of the vegetation in orchards have been carried out e.g. by Huck and Fischer (1988), Langensiepen and Otte (1994), Wolf and Hemm (1994), Bünger (1996) and Hofbauer (1998). After the Second World War, many standard tree orchards were ripped out and replaced by more easily manageable low-growing forms that can be managed more intensively. These orchards contain species-poor, short-cut grasslands suitable for machinery to drive around between the trees, and are mown 5 to 15 times per year. Rösler (2003) compared the vegetation of traditional, standard orchards with that of intensive ones. In many cases, the last remnants of species-rich mesic meadows can now be found in traditional orchards (Kornprobst 1994). Their stuctural heterogeneity means that they are particularly valuable not only in terms of their flora, but also in their diversity of animal species (Langensiepen and Otte 1994; Weller 1994). The trunks of old fruit trees also support a wide range of epiphytic lichen communities (Kupfer-Wesely and Türk 1986).

8.3.6  Pastures and Frequently Mown Grasslands Lowland Pastures  Only some of the grazed grasslands, namely the fertilised pastures, belong to the agricultural grassland class of the Molinio-Arrhenatheretea. The majority of the unfertilised, nutrient-poor pastures with low grazing pressure is instead placed within the order of the Nardetalia (see Chap. 6) and the class of the Festuco-Brometea (Chap. 7). Intensively used and fertilised pastures and frequently mown lawns or parks are floristically so similar that they are combined within the Cynosurion. This alliance is mainly negatively characterised by the lack of character species of mesic meadows, and even the grazing-specific species (particularly the pasture weeds) are often absent. The average number of species per plot is around 30, and in intensive pas-

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tures even less than 20, which is similar to intensive meadows. Dierschke and Briemle (2002) consider at most Agrostis stolonifera, Bellis perennis, Lolium perenne, Plantago major agg., Poa annua, Ranunculus repens, Rumex crispus and R. obtusifolius to be characteristic, as they are slightly more common in pastures than in other vegetation types. A single plant community dominates these pastures across all soil types, the Lolio-Cynosuretum. In the Netherlands and northwestern Germany, which is the optimum zone of this community on the European mainland, various sub-­associations, varieties and forms have been distinguished (see Table  8.10). The most common were the typical (a) and the damp (e) sub-community, whilst the habitats of the dry type (c) are mostly planted with crops. The saline type (g) is restricted to a narrow transitional zone between pasture and salt marsh communities on the North Sea and Baltic coasts. The soft soils of damp pastures often mean that the sward is damaged by trampling, promoting indicators of disturbance such as Juncus effusus and J. inflexus (Fig.  8.29). Nutrient-poor pastures with similarities to the Nardetalia and Festuco-Brometea communities used to be widespread mainly on acidic or dry soils in mountain areas, but have now become rare in many parts of Central Europe. Remnants of this relatively species-rich community sometimes persist under fences, where the fertilisation and grazing pressure is lower. These ‘pasture fence communities’ have been studied by Vollrath (1970) and Husicka and Vogel (1999). Management intensification has caused this differentiation of nutrient-rich pastures to become blurred or even disappear. The more intensive the use, the fewer species the pasture contains, until only a few species with high regenerative ability that are strongly promoted by fertilisation remain, such as Lolium and Phleum (see Fig. 8.30). These are joined by creeping plants such as Trifolium repens and species of trampled areas (e.g. Poa annua, Plantago major, Polygonum arenastrum), pasture weeds (particularly species of Rumex) and also arable plants (e.g. Stellaria media) on patches of faeces. Slightly species-richer relicts with a few indicators of nutrient-poor conditions are now often found, if at all, along ditches (e.g. Konrad and Ruthsatz 1993). The now widespread practice of rotational grazing with a first silage cut has also blurred the floristic boundary between the Lolio-Cynosuretum and Arrhenatheretum, so that modern intensive grassland cannot clearly be assigned to either (see Fig. 8.31; Vollrath 1970; Schwaar 1973; Dietl 1995). A revision of these communities thus seems advisable. Montane and Alpine Pastures  With decreasing length of the growing season, high fertiliser and management costs become less and less worthwhile. The intensity of grazing, like of mowing, thus generally decreases with increasing elevation (and latitude). The Lolio-Cynosuretum becomes less common, being replaced by the nutrient-poor Festuco-Cynosuretum wherever grazing takes place. This community is dominated by Festuca nigrescens and other undemanding and low-­ growing grasses (see Table  8.10). Other widespread names are the Luzulo-Cynosuretum (after Luzula campestris, Meisel 1966b) and Anthoxantho-­ Agrostietum (Jurko 1974). In any case, their differentiation from better fertilised

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Table 8.10  Habitat classification of white clover meadows in Central Europe a

b

c

d

e

Species occurring in nearly all white clover-pastures (Cynosurion in a wider sense): G: Cynosurus cristatus L: Trifolium repens Leontodon hispidus Festuca rubra T. pratense Odontites rubra Phleum pratense H: Achillea millefolium Plantago lanceolata Poa pratensis Cerastium fontanum Prunella vulgaris holosteoides Bromus hordeaceus Geranium molle Ranunculus acris Hordeum secalinum Leontodon autumnalis Taraxacum (coast) officinale In ryegrass-white clover-pastures (Lolio-Cynosuretum), intensively fertilised and mainly in the lowlands: G: Lolium perenne H: Bellis perennis Plantago major L. hybridum Cirsium arvense Potentilla anserina Poa annua C. vulgare Ranunculus repens P. trivialis Glechoma hederacea Sagina procumbens Elymus repens Lysimachia nummularia Veronica serpyllifolia In red fescue-white clover-pastures (Festuco-Cynosuretum), poorly manured and mainly at higher elevations; soil relatively dry: G: Agrostis capillaris H: Alchemilla vulgaris Plantago media Festuca nigrescens Campanula rotundifolia Polygala vulgaris Anthoxanthum Euphrasia rostkowiana Potentilla erecta odoratum Briza media Hieracium pilosella Ranunculus bulbosus Danthonia Hypochoeris radicata Stellaria graminea decumbens Luzula campestris Leucanthemum vulgare Thymus serpyllum L: Lotus corniculatus Pimpinella saxifraga etc. In matgrass-rich pastures (F.-C. nardetosum) extensively managed and very acidic, mostly at higher elevations: G: Nardus stricta H: Antennaria dioica Calluna vulgaris Deschampsia Hypericum maculatum Vaccinium myrtillus flexuosa Carex pilulifera Veronica officinalis etc. Indicators of moisture: f Indicators of sand (coast): g Indicators of soil salinity (coast): (.lotetosum uliginosi) (.armerietosum) (.juncetosum gerardii) G: Carex leporina G: Festuca rubra subsp. G: Juncus gerardii arenaria L: Lotus uliginosus H: Armeria maritima Festuca rubra subsp. litoralis Juncus effusus Galium verum L: Trifolium fragiferum H: Cirsium palustre Viola canina H: Glaux maritima Lynchis flos-cuculi etc. etc.

In the style of Tüxen and Preising (1951), Meisel (1966b), Passarge (1969b), Jurko (1974) and ­others G = grasses and graminoids, L = legumes, H = herbs and dwarf shrubs

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Fig. 8.29  An extensively grazed Lolio-Cynosuretum on the island of Neuwerk in the German Wadden Sea. Undergrazing favours Cirsium and Juncus species; trampling has created patches of bare soil

Fig. 8.30  Intensively used and highly trampled Lolio-Cynosuretum sward on fertile soil with Lolium perenne, Poa annua, Achillea millefolium and Plantago major

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Fig. 8.31  The effects of management intensification on the vegetation types in Central European dry to wet grasslands (Modified from Dierschke and Briemle 2002; with permission of Ulmer Verlag, Stuttgart)

Lolio-Cynosuretum communities is difficult, because none of the pasture communities has good character species and many transitional forms exist. Table 8.10 thus limits itself to habitat-based groupings and lists several sub-communities, of which the Nardus-rich type is common at higher elevations. In the upper montane belt, the Lolio-Cynosuretum is replaced by communities rich in Leontodon and Crepis (see below), with increasing numbers of montane and subalpine species, or resembles Festuco-Brometea grasslands (see Table 8.14). These mountain pastures are much species-richer, but lower yielding. In contrast to meadows on the one hand and nutrient-poor pastures on the other, fertilised pastures are only negatively characterised even at high elevations, namely by the absence of tall species sensitive to grazing, and few light-demanding species. However, they do often contain Leontodon hispidus and Crepis aurea (see Fig. 8.32), and one of the most widespread of these communities is the Crepido-Festucetum commutatae (Mucina et al. 1993). In contrast to the lowland pastures, these grasslands contain a lower proportion of grasses and a higher proportion of forbs (Marschall 1958). They are the most valuable communities of the small-scale dairy farmers in the subalpine and lower alpine belt, and form there mainly on flat, deep soils (see Fig. 8.33). Similarly to the Festuco-Cynosuretum, they often occur along-

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Fig. 8.32  Pasture rich in Leontodon hispidus and Crepis aurea below the tree line on acid sandstone in the Salzburg Alps (central Austria), which transforms into a Geo montani-Nardetum on the less well-fertilised slopes. Cattle trampling has partly destroyed the sward. In the depression in the background, several moisture indicators are present

side Nardus grasslands, from which they can form as a result of fertilisation (see Sect. 5.4.3), as well as alongside alpine grassland communities. The latter provide the differential species that allow the Crepido-Festucetum to be distinguished from the Festuco-Cynosuretum. Garden lawns contain even fewer species than pastures, as they are mown 20 to 30 times a year with the application of mineral fertiliser (see Fig. 8.30). They are clearly most closely related to the Lolio-Cynosuretum. The communities identified by Kienast (1978), N. Müller (1989), Müller and Sukopp (1993) and others (Trifolio-­ Veronicetum filiformis and Festuco-Crepidetum capillaris) are strictly speaking not associations, as they do not have any of their own character species. The neophyte Veronica filiformis is now characteristic of many garden lawns, spreads vegetatively and is promoted by frequent mowing but not by grazing (N. Müller 1989).

8.3.7  Vegetation of Trampled Ground and Flooded Grassland 8.3.7.1  The Vegetation of Trampled Paths and Areas Intensive pastures are so species-poor partly because only a few species tolerate the mechanical damage of trampling. The effects of trampling are greatest on oft-­ frequented paths and poorly maintained sports pitches with loamy soils. These areas

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Fig. 8.33  Cattle pasture in the subalpine belt (Austrian Limestone Alps, near Innsbruck) with a livestock resting area, where a productive Rumicetum alpini with Rumex alpinus, Senecio alpinus, Rumex alpestris and other nutrient-demanding forbs has established

support one of the most uniform plant communities in the world, the Lolio-­ Plantaginetum majoris. These always contain the following species in various proportions (see Fig. 8.40): Plantago major Matricaria discoidea Lolium perenne

Poa annua Polygonum arenastrum

Other species that often also occur here but are still mainly found in agricultural grassland or arable communities are e.g.: Agrostis capillaris Capsella bursa-pastoris Leontodon autumnalis

Taraxacum officinale agg. Trifolium repens

This community resembles that of the Lolio-Cynosuretum, and will develop into this if trampling disturbance ceases. Foerster (1983) distinguished several forms based on soil moisture. Passarge (1979b) studied the montane forms of this community, in which the frost-sensitive Lolium is replaced by Poa pratensis, which itself is replaced by Poa supina in the subalpine belt. Other communities of trampled ground rich in perennials are the Juncetum tenuis on shady forest paths with

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the neophyte Juncus tenuis, and the Prunello-Plantaginetum majoris at the edge of wheel tracks. The Lolio-Plantaginetum majoris is unable to develop in its typical form on pure sands, the surface of which quickly dries out. Long drought periods are also the main reason why perennial-rich communities are also rare in trampled areas in the Mediterranean, whilst the cool and damp oceanic climate in northwestern Europe promotes their growth. Therophytes dominate in areas that are both dry and disturbed. Between paving stones and on cinder or gravel tracks, i.e. in relatively nutrient-­ poor places, bryophytes are also able to develop in the trampled communities, as they exploit the small spaces between the stones where they are protected from mechanical damage. This is mainly the low-growing cushion moss Bryum argenteum, which tolerates occasional drought. The Sagino-Bryetum community of paving cracks is only well developed in oceanic climates, particularly in the Netherlands and in northwestern Germany. In drier regions such as the northern Upper Rhine Plain and eastern Austria, the green Lolio-Plantaginetum is replaced by the more drought-tolerant Sclerochloo-Polygonetum avicularis (Korneck 1969; Chytrý 2009) which turns brown in summer. Other communities are characterised by the presence of Poa annua, Spergularia rubra, Malva neglecta or Eragrostis minor, and occasionally also Coronopus squamatus, Herniaria glabra, Myosurus minimus or (in warmer regions) Cynodon dactylon (see e.g. Oberdorfer 1971; Mucina et al. 1993; Pott 1995; Rennwald 2000; Schubert et al. 2001; Wittig 2001, 2002; Chytrý 2009). Some of these sparse stands superficially resemble ruderal communities, although they lack the ruderal character species (Falinski 1963; Berset 1969; Oberdorfer 1971; Mucina et al. 1993). More recent phytosociological overviews often distinguish between perennial and annual communities of trampled areas. The former is assigned to the Molinio-Arrhenatheretea (i.e. agricultural grassland), and the latter to its own class Polygono-Poetea annuae, although this has only few character species. Other authors prefer placing all trampled communities in a single class, the Plantaginetea majoris (e.g. Schaminée et al. 1996; Wittig 2001, 2002). If the vegetation were allowed to develop naturally, all trampled communities would disappear from the Central European landscape as quickly as most grassland, ruderal and arable weed communities. The mechanical disturbance caused by humans or their domestic animals is the only factor protecting this vegetation from the dense forest shade. 8.3.7.2  C  ommunities of Creeping Plants and Flooded Grasslands (Potentillo-Polygonetalia) On river banks and lake shores, as well as sea shore areas rich in seaweed, flooding, wave action and stagnant water promote the formation of natural plant communities that resemble those of trampled areas. Such flooded communities are often also used by ducks and geese, as well as grazing animals and humans, making them

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more similar to the trampled communities than would otherwise be the case (see  Fig.  8.34). The flooded grasslands are classified as the order PotentilloPolygonetalia with the single alliance of the Potentillion anserinae, usually within the class of agricultural grassland Molinio-Arrhenatheretea (Tüxen 1970, but cf. Sykora 1982 and Oberdorfer 1983b; see Chapter 14: 5.4.3). Typical coastal Potentillo-Polygonetalia communities are particularly frequent in Northern Europe and were first described here by Nordhagen (1940) as flooded grasslands influenced by high salinity, the Agropyro-Rumicion. There have since been several studies on these communities (e.g. Tüxen 1970; Müller 1961; Sykora 1982; Oberdorfer 1983b; Verbücheln 1987). These frequently disturbed and species-poor stands are characterised by the presence of Agrostis stolonifera (creeping bentgrass), which is particularly tolerant of wet soils (Davies and Singh 1983). Many species of flooded grasslands are able to form large networks of stolons that can quickly put down roots when they encounter damp, bare soil, e.g.: Potentilla anserina Potentilla reptans Ranunculus repens

Mentha longifolia Mentha pulegium Mentha suaveolens

Other species achieve the same effect with their below-ground rhizomes: Elymus repens Carex hirta

Rorippa sylvestris

Some of these species also occur in trampled, ruderal and weed communities, or fertilised grassland or river banks and muddy areas, but they are mainly found in the natural flooded grasslands. The same is true for some tussock hemicryptophytes and several Rumex species: Festuca arundinacea Juncus inflexus J. compressus

Rumex conglomeratus R. crispus R. stenophyllus

Some of these species can be used as character species of particular associations, e.g. Rorippa sylvestris (for the Rorippo-Agrostidetum on gravelly riverbeds, see Fig.  8.34), Festuca arundinacea (for the Dactyli-Festucetum arundinaceae of higher-lying but still flooded areas), Mentha longifolia (for the Mentho-Juncetum inflexi on temporarily flooded marls) and Mentha pulegium (for the thermophilic Potentillo-Menthetum suaveolentis). These coastal communities are ecologically quite similar to the inland flooded grasslands, and are difficult to separate from the more nitrophilic drift lines and the foredunes (Salsolion and Agropyro-Honckenion, see Sect. 2.3.1). Flooded grasslands also form mosaics with halotolerant communities near to inland saline areas (cf. Fig. 8.35). Fragments of flooded grassland can also occur within the meadows and pastures of river valleys dominated by loamy soils, where they are clearly visible as dense,

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Fig. 8.34  The Rorippo-Agrostietum of Czech and Slovakian river floodplains are usually dominated by grasses or species of Potentilla. Wetter depressions contain almost only Polygonum amphibium (water knotweed) (Modified from Krippelová 1967)

Fig. 8.35  Cross-section through a weakly saline flooded grassland (Triglochino-Agrostietum stoloniferae) by the Schwielow lake in the Potsdam Havelland near Berlin (Modified from Konczak 1968) Apart from Leontodon autumnalis and Triglochin palustris, all species spread quickly vegetatively, with the Trifolium species, Potentilla anserina and Ranunculus repens via above-ground stolons

dark green patches. They often form in shallow depressions that collect rainwater and are flooded for longer periods than the surrounding areas, so that the true meadow plants are hindered by the hypoxia in the soil. Sometimes all of the higher plants die here, leaving space for species of the Potentillion anserinae to colo-

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671

nise. Their proportion in the grassland can vary depending on the weather and the length of flooding. Similar population fluctuations in Cnidion dubii meadows caused by the annual variation in flooding patterns were documented by Balátová-­ Tulácková (1979b). Tüxen (1979c) compared the temporal fluctuations in Potentillion anserinae communities with the movement of an accordion. Particularly the Ranunculo-Alopecuretum geniculati association is involved in this acyclical succession, which is characteristic for river valley grasslands, but does not occur further south than southern Germany. Alopecurus geniculatus, after which the community is named, is characterised by vegetative reproduction as well as the production of large numbers of floating seeds. This allows it to easily spread to and colonise suitable habitats. The dispersal of propagules with floodwater or wave action and their rapid establishment on the wet surface of the soil is probably also decisive in the ability of the other communities of the Potentillo-Polygonetalia to rapidly colonise small and isolated patches of ground.

8.4  Adaptations to the Environment 8.4.1  Tolerance of Mowing and Trampling Effects of Mowing  Mowing, grazing and trampling are decisive habitat factors of agricultural grassland, and have led to some typical adapatations of the species that occur there. Mowing results in a sudden loss of all but the lowest leaves and stems of the plant, drastically reducing the photosynthetic capacity and the amount of nitrogen stored in the plant, and temporarily decreasing nutrient uptake (Richards 1993). The sudden increase in light levels at the base of the plant can also lead to photoinhibition and damage from drought stress. Frequent cutting therefore considerably reduces the productivity of meadow plants. Klapp (1971) measured a reduction in average hay yield to 60 % at 6 to 10 cuts per year, and 51 % at 15 to 20 cuts, compared to the yield at 3 cuts per year (see Fig. 8.52). Frequently cut lawns are thus highly disturbed ecosystems with a much reduced productivity. The survival and growth of mown grassland plants is primarily determined by their ability to rapidly replace their leaves, i.e. to develop allocation patterns adapted to the disturbance regime (Briske and Richards 1995). This promotes the growth of clonal plants with high plasticity in their shoot formation (de Kroon and Hutchings 1995). Meadows cut once to twice a year are dominated by tall grasses and forbs, and the more frequent the cutting, the more numerous the low-growing species. Well-tended lawns are formed exclusively of low-growing grasses, white clover and some rosette forbs. Mowing-tolerant species have many prostrate leaves (e.g. Lolium and Plantago) and have rhizomes or other storage organs in or directly above the soil, in which soluble carbohydrates (particularly fructans) and nitrogen compounds are stored. The degree of translocation of nutrients into these storage organs and their remobili-

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8  Agricultural Grassland on Mesic to Wet Soils

sation determines the growth rate after cutting or grazing (Dhont et  al. 2002). Meadow plants such as Alopecurus, Dactylis, Trisetum and Festuca pratensis, which can quickly move assimilates into the basal parts of the shoots and below-­ ground storage organs, are more tolerant of mowing than those that need longer (e.g. Molinia). The amount of stored substances in the stubble of mown plants is correspondingly variable (Dactylis 19.0 % fructosan content, Phleum 16.4 %, Lolium 11 %, Arrhenatherum 4.6 %, Simon and Daniel 1977). Arrhenatherum has only an intermediate tolerance of mowing. Figure 8.36 shows how quickly mowing or grazing can promote the best-adapted species from sown mixtures using the examples of Dactylis and Lolium. A further successful growth strategy of productive meadow plants is the production of leaves with minimum use of resources, i.e. a high specific leaf area (de Visser et al. 1997). This allows rapid leaf development after leaf loss and a quick return of the C investment (Poorter 1994, Westoby et al. 2002). Increasing frequency of cutting reduces the plant’s root mass and its stored reserves, as well as the average shoot size, thus reducing its regenerative potential (Matthew et al. 2000). High levels of N fertilisation can reduce the ability to resprout, because faster growth is linked to lower investment in below-ground storage organs (Klapp 1971). Interestingly, the ability of a grass to resprout after grazing is also largely determined by the identity of its direct neighbours (Crawley 1983), which can increase or decrease competition intensity compared to growth in monoculture. Dierschke and Briemle (2002) evaluated the mowing and grazing tolerance of 396 taxa of Central European agricultural grassland on a 9-step scale. According to this, extremely mowing-tolerant species include Bellis perennis, Agrostis stolonifera, Festuca rubra, Prunella vulgaris and Poa pratensis (Mo score 9). Dactylis

Fig. 8.36  The influence of mowing and grazing on the proportions of major grasses in stands starting with a similar seed mixture on dry, moderately fertilised sandy soils (From data in Könekamp and König (1929) and Schwarz (1933) in Ellenberg (1963)) Left: Cut several times a year for 3 years and then grazed by sheep for 3 years. Right: The reverse. Dactylis glomerata grows poorly under grazing but spreads under mowing management, Lolium perenne follows the opposite pattern. Poa pratensis spreads under both treatments, whilst Festuca pratensis and a variety of Phleum pratense decline. Y axis = percent fresh weight. The ‘other species’ are in some cases pasture weeds

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673

glomerata, Lolium perenne, Poa trivialis, Phleum pratense and Trifolium repens are also very tolerant, but slightly less so (8). Dierschke and Briemle (2002) found that a cutting level of 7 cm was particularly good for the regeneration of many meadow species. The important fertile meadow grasses Alopecurus pratensis, Arrhenatherum elatius and Festuca pratensis were assessed as being moderately to highly tolerant (6 and 7). In contrast, sensitive species (Mo score 3 and 2) include Nardus stricta, and some hygromorphic to mesomorphic grassland herbs (e.g. Cirsium palustre, Epilobium hirsutum, Filipendula ulmaria, Galium boreale, Iris sibirica, Lysimachia vulgaris, Lythrum salicaria, Molinia caerulea, Selinum carvifolia, all with (3). Highly sensitive species are the taller woody species. Regular mowing produces an alternation between ‘high’ and ‘low’ levels of vegetation (see Fig. 8.1). The different species fit into this sequence of shade and light phases depending on their growth patterns. Spring plants such as Primula elatior, Leucojum vernum, Anemone nemorosa, Crocus albiflorus or Narcissus species (see Fig. 8.20) only make use of the first light phase. They decorate the meadows particularly in areas where a thick snow layer has pressed the grasses to the ground in winter and remains until spring. The hemicryptophytes do not start growing immediately, and Crocus and other geophytes use this pause to develop unhindered. They rely on a phenological head start, which also promotes the secondary spread of Narcissus pseudonarcissus (wild daffodil; Duhme and Kaule 1970). In the cold lowland winters with little snow, they lack this competitive advantage and therefore these areas do not have the attractive spring aspect of the mountain meadows. Taraxacum, Bellis perennis and many low-growing legumes flower during all light phases, in fact their flowering increases with the frequency of cutting. Colchicum autumnale (autumn crocus) produces only a flower in the last light phase of the year, and only unfurls its leaves and fruiting stem in early summer the following year. Euphrasia rostkoviana (eyebright) is a well-known example of seasonal dimorphism. Its tall unbranched form flowers and produces seed in the first shade phase of the meadow, whilst the genetically determined low-growing branched form develops in the later light phase. Tall meadow grasses and forbs form and use either all the shade phases or just one of them. Anthriscus sylvestris (cow parsley), for example, usually flowers before the first cut of meadows with two cuts a year, whilst Heracleum sphondylium, Cirsium oleraceum and other large-leaved tall forbs only flower before the second (see Figs. 8.21 and 8.22). Effects of Grazing  Grazing influences grassland communities through both herbivory and trampling with the resulting soil compaction. Different types of livestock and grazing systems can cause pasture communities with different structures and species compositions to form. Sheep graze very selectively and nibble down to close to the ground, so that the basal buds may be damaged. High densities of sheep can therefore cause gaps to appear in the sward (see Fig. 8.37). Goats have a broader range of preferred plants than sheep, and will even consume the leaves and small branches of woody plants. Horses may also damage the sward by grazing close to the soil surface. However, the soil compaction is much greater from horse grazing (down to 10–15 cm) than from sheep or goat grazing (down to around 5 cm), so that

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Fig. 8.37  The influence of livestock on grassland under appropriate grazing pressure (From von Korn (1987, in Dierschke and Briemle 2002))

water can be trapped on the soil surface in places. Cattle grazing is less selective, as they use their tongue more to rip plants from the ground rather than cutting them with their teeth, which also means that they cannot graze so close to the ground. Livestock tend to prefer the earlier developmental stages of the plants, which have high nutrient but low lignin contents. They also often prefer mixed-species stands over monocultures. Klapp (1971) lists the following species according to decreasing palatability: Trifolium repens, Plantago lanceolata, Taraxacum officinale (young) > Phleum pratense, Festuca pratensis, Dactylis glomerata > Lolium perenne > Festuca rubra, Holcus lanatus, Festuca arundinacea > Nardus stricta, Molinia caerulea + many pasture weeds. Effects of Trampling  Trampling influences the species composition of pastures and trampled communities via several mechanisms (see Fig. 8.38). The soil of pastures is more compacted and has a lower pore volume than that of meadows or crop fields (Lieth 1954; see Fig. 8.39). Klapp (1965) found pasture soils to be on average 13 % denser than meadow soils. The only more compacted soils are those at the edges of paths or in sports pitches. The pasture communities are also often replaced by communities of trampled areas close to gates, drinking troughs and feeding areas (see Fig. 8.30 and Table 8.11). However, beneficial side effects of the mechanical damage are the high light levels and the reduction in competition. Generally, trampling is also linked to eutrophication, either through the dung of the livestock or the proximity to settlements in the case of many trampled communities. Without the high nutrient levels, the damaged plants would not regenerate so well and the seedlings would not be able to develop so quickly. Only trampling-tolerant species and those that can quickly regenerate can survive in intensively managed pastures and trampled communities. The resistance of the perennial or annual/biennial plants to frequent trampling is mainly linked to

8.4  Adaptations to the Environment

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Fig. 8.38  Livestock and human trampling has both direct and indirect effects on the vegetation. The trampling-adapted plants are less severely damaged than the other species and therefore have a relative advantage in colonisation and competition (Modified from Lieth 1954)

Fig. 8.39  The relationship between the pore volume of the soil and the management of permanent grassland. 70–198 measurements in the topsoil down to 10  cm (Modified from Lieth 1954). Trampling compacts the soil, causing the pore volume to be smaller in areas often supporting livestock

their small size, their branching close to the soil surface, the elasticity and stability of their tissues (Soerkarjo 1992), their rapid regeneration and their high seed ­production. On the most intensively trampled areas, at most Poa annua or Polygonum aviculare manage to produce ripe seed. Bothmer (1953) and Vollrath (1970) determined the trampling tolerance of different species by measuring the minimum distance of certain species and vegetation types from the pasture entrance that had been trampled bare (see Table  8.11). Haessler (1954) produced similar results using an experimental approach. The high-

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Table 8.11  The influence of livestock trampling on the species composition of a well-fertilised ryegrass pasture Community no.: Distance to paddock gate (up to m) A Echinochloa crus-galli R Glyceria plicata A Atriplex patula T Juncus tenuis A Chenopodium glaucum T Matricaria discoidea T Polygonum aviculare T Poa annua T Plantago major Agrostis stolonifera Poa pratensis Trifolium repens A Rumex obtusifolius A Capsella bursa-pastoris Lolium perenne Taraxacum officinale A Stellaria media Ranunculus repens Dactylis glomerata Festuca pratensis Potentilla anserina Elymus repens Poa trivialis Phleum pratense Leontodon autumnalis Bellis perennis Deschampsia cespitosa Achillea millefolium Plantago lanceolata Trifolium pratense Bromus hordeaceus Alopecurus pratensis Ranunculus acris Glechoma hederacea Cerastium fontanum   holosteoides

1

2

3

4

6 4 3 3 2 3 5 5 5 5 5 5 5 4 4 4 3 3 3 2 2

18

63

360

1 4 4 5 5 4 5 5 5 5 5 5 4 1 4 2 4 3 2 1 1 1

4 5 4 5 5 4 5 5 5 5 3 5 5 4 4 2 4 4 2 2

1 2 5 5 5 5 4 5 5 5 5 5 5 4 2 5 5 2 2 5 2 3 3 5 4 3 3 3 2

M

N

5 10 5 6 6 5 4 6 5 × 5 5 6 5 5 5 × 7 5 6 6 × 7 5 5 5 7 4 × × × 6 6 6 5

8 8 7 5 9 8 6 8 6 5 6 6 9 6 7 7 8 × 6 6 7 7 7 6 5 6 3 5 × × × 7 × 7 5

(continued)

8.4  Adaptations to the Environment

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Table 8.11 (continued) Community no.: Distance to paddock gate (up to m) Heracleum sphondylium Carum carvi F Crepis biennis Mean coverage (%) Mean species richness of the sward

1

2

3

4

6

18

63

360 2 2 2 97 28

46 15

91 16

98 17

M

N

5 5 5

8 6 5

From surveys by Vollrath (1970) on the rotational pastures of the Veitshof (Upper Bavaria) at increasing distance from the pasture entrance The numbers refer to the constancy over several relevés1 Ecological indicator values are from Ellenberg et al. (1992) T = trampled area plant (Plantaginetalia), F = fertilised meadow plant (Arrhenatherion). Some species with low constancy were omitted Ecological indicator values: No. 1: mM: 5.4, mN: 6.9, ∑T: 22, ∑A: 21. Close to the pasture entrance, the most trampled, over 50 % bare earth; wet in places (Glyceria!) No. 2: mM: 5.3, mN: 6.6, ∑T: 18, ∑A: 15. Strongly trampled, sward short with gaps. The large-­ leaved Rumex obtusifolius, a nitrogen-demanding pasture weed, is common in places No. 3: mM: 5.5, mN: 6.5, ∑T: 9, ∑A: 14. Normal pasture, dense sward dominated by low-­ growing grasses; relatively species-poor like no. 1 and 2 No. 4: mM: 5.3, mN: 6.4, ∑T: 8, ∑A: 14. Low grazing pressure, therefore meadow-like, less fertilised with dung (mN relatively low) and much more species rich mN = average nitrogen IV (less nutrient-demanding species become more frequent with distance from the pasture entrance, but species with high EIV-N occur throughout the pasture) ∑T = sum of the constancy values of the species of trampled areas ∑A = sum of the constancy values of the arable weeds and ruderal plants (which can also be considered indicators of mechanical disturbance) 1 A = arable weed or ruderal plant (Stellarietea, Artemisietea), R = stream reedbed plant (Glycerio-Sparganion) 2 mM = average moisture IV (the soil moisture is largely independent of grazing)

est trampling pressure is tolerated by therophytes such as Matricaria discoidea and Polygonum aviculare, which can survive highest trampling intensity only in small patches. As the trampling pressure decreases, Trifolium repens spreads to join these species, forming a transitional community to the Trifolio-Plantaginetum majoris, which itself contains increasing amounts of Lolium perenne and other species further away from the main trampled area (see Fig.  8.40). This community often merges with the Lolio-Plantaginetum majoris (see Sect. 8.3.7.1). Effects of Under- and Overgrazing  The stocking density determines the effects of grazing and trampling. Low densities lead to selective undergrazing, because more fodder is available than is needed. This leads to the spread of unpalatable, hard-leaved, poisonous or thorny pasture weeds such as Deschampsia, Carex, Juncus, Euphorbia, Artemisia, Rumex, Gentiana, Colchicum and Ononis, as well as shrub species that encroach into the pasture from the edges (see Fig. 8.29). Selective overgrazing, in contrast, damages the palatable plants and promotes the spread of a few particularly trampling-tolerant and low-growing species, including Plantago, Prunella, Bellis and Taraxacum.

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Fig. 8.40  The normal zonation of communities of trampled ground on loam soil between a path and a fertilised grassland (Modified from Oberdorfer 1971) The highest trampling pressure is tolerated by therophytes such as Matricaria discoidea (pineapple weed) and Polygonum aviculare (knotgrass), although even these only occur in patches. Trifolium repens joins these species in the transitional community to the Trifolio-Plantaginetum majoris, whereby the latter already contains Lolium perenne and other species F. Runge (1981b) notes that these communities are not evergreen, as was previously thought, even in the oceanic climate of the Westphalian Basin. With the exception of the Lolio-Cynosuretum, they all belong to the alliance of the Polygonion arenastri

The effect of trampling also plays a large role outside of the typical trampled pastures. For example, trampling by large numbers of holidaying visitors to the Upper Bavarian Ostersee lakes has led to a decline in the total species number in the semi-dry grasslands there from 59 to 30, and in the dry grassland plants from 27 to 7 (Seibert 1974). These have been replaced by species of agricultural grassland and trampled areas.

8.4.2  Some Ecophysiological Properties of Grassland Species 8.4.2.1  Light Absorption and Self-Shading The vast majority of grassland plants are adapted to high light levels: of 500 Central European grassland plants, 420 have an EIV for light of 7 or above. As in other ecosystems, there is a strong relationship between the productivity of grasslands and the absorbed light, whereby both generally increase with leaf area. However, self-shading in dense unmown stands reduces the productivity of meadows via several mechanisms (Matthew et al. 2000). A large part of the leaf area must reduce its

8.4  Adaptations to the Environment

679

photosynthetic activity due to lack of light, because the middle and upper layer is formed from light-demanding grasses and forbs. It is only in the lower layers that shade-tolerant, low-growing creeping and rosette plants dominate. On the other hand, the newly-formed leaves in the increasingly dense stand have a lower photosynthetic capacity than those that were formed earlier in the year under higher light conditions (Parsons and Robson 1981). Thus, in a growing stand, the leaf area index increases whilst the photosynthetic capacity of the newly formed leaves drops. Increased mortality also leads to an increase in the dead leaf mass in the stand interior and therefore a further reduction in productivity. The photosynthetic capacity of the leaves reduces naturally with age anyway (Woledge and Leafe 1976). The productivity of a growing sward therefore reaches a maximum before sinking again, despite the growing leaf area, due to the increasing mortality rate of the shaded leaves (Marshall 1987). 8.4.2.2  The Influence of Drought Leaf area development in Central European meadows and pastures is to a large extent determined by the water supply. Almost all of the meadow grasses grow best when the soil water content is at around 85 % of its field capacity (Kautner 1933 and see Ellenberg 1963). Even in the relatively humid climate of Central Europe, meadows and pastures frequently suffer from drought stress, particularly in mid and late summer, and particularly on sandy and clay soils and in regions with high evaporative demand. Lack of water reduces cell elongation and therefore the production of new leaves in the grass plants often from May onwards (Turner and Begg 1978; Jones et al. 1980). Many grasses also roll their leaves during dry periods, reducing both the transpiring surface and the photosynthetic capacity of the leaf (Gastal and Durand 2000). The consequences are visible in the characteristic mid-summer depression in growth in many Central European meadows and to a lesser extent also in pastures (Anslow and Green 1967; see Fig. 8.41). In addition to drought-­ stress, the endogenous developmental rhythms also play a role, with generative development in spring and side shoot formation in summer. The changes in precipitation from year to year can cause considerable fluctuations in the coverage of individual species as well as the total number of species in meadows. On a plot in a Cirsium rivulare meadow in the Black Forest, the number of species varied over eight consecutive years between 52 and 60, whereby Carex panicea was particularly sensitive to the weather conditions (Nowak and Schulz 2002). Species-specific differences in the reaction to summer drought stress are important factors causing the different species compositions of mesic meadows and dry grasslands. The competitive ability of Arrhenatherum, Phleum, Alopecurus, Lolium and other species of mesic meadows and pastures decreases with decreasing moisture, because they respond to drought with a stronger decrease in their dry mass and often also in seed production than e.g. Bromus erectus, Brachypodium pinnatum or Koeleria pyramidata (Salinger and Strehlow 1987, Ködderitzsch and

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Fig. 8.41  Changes in the daily above-ground growth rate in pastures over the course of a year at different elevations in the Alps (average over 2 years) (From Caputa and Schlechtner 1970). In the lowland and lower mountain areas there is a mid-summer depression in growth caused mainly by drought

Leuschner, unpublished; see Fig. 8.42). Lolium, which has a highly oceanic range, suffered strongly from long periods of drought in soil moisture experiments. On average, it had a higher transpiration rate than Arrhenatherum and did not reduce water loss as much when under drought stress. Under long periods of drought, Lolium therefore reached irreversible wilting earlier than Arrhenatherum (Didden-­ Zopfy 1981). Arrhenatherum and Alopecurus, in turn, were more affected by dry soils or low groundwater levels and had a much more reduced yield than Bromus erectus (Ellenberg 1953; see Fig. 8.42). The mesic meadow grasses are therefore competitively inferior to the dry grassland species during periods of drought, although they would dominate under ample water supply (Sharifi 1983). This may also explain why Arrhenatherum does not spread as far into the dry continental climate of Eastern Europe as Bromus.

8.4  Adaptations to the Environment

681

Fig. 8.42  Dry matter yield (in % of the maximum) of eight grasses of mesic meadows and semi-­ dry grasslands in monoculture on sandy soils under different groundwater levels that are maintained constant within each treatment (a; from Sharifi 1983) and at different but constant soil moisture (b and c; from Zielke, Ködderitzsch and Leuschner, unpublished). Aa = Alopecurus aequalis, Ae = Arrhenatherum elatius, Ap Alopecurus pratensis, Be Bromus erectus, Bp Brachypodium pinnatum, Dg Dactylis glomerata, Kp Koeleria pyramidata, Pp Phleum pratense. The damp meadow grass Alopecurus pratensis is more sensitive to the deep groundwater level than Arrhenatherum or Bromus. Low soil moisture leads to greater reductions in yield in the mesic meadow grasses Arrhenatherum and Phleum than in the dry grassland species Brachypodium, Bromus and Koeleria. High soil moisture also reduces the yield in most species (apart from Alopecurus pratensis)

In contrast, Arrhenatherum and Dactylis are more successful on soils with high groundwater levels when rainfall is low and the evaporative demand is high, because the competition from damp meadow grasses such as Deschampsia cespitosa, Alopecurus pratensis or Poa palustris is reduced. Arrhenatherum and its accompa-

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8  Agricultural Grassland on Mesic to Wet Soils

nying species have their widest amplitude in soil moisture levels in the moderately warm and dry climate of the foothills of the Swabian Jura and Black Forest region in southern Germany (location 1 a in Fig. 8.15). Nevertheless, it would be wrong to assume from such an ecological sequence that grasses such as Bromus erectus, Avenochloa pratensis and Phleum phleoides are xerophilic with their physiological optimum in dry habitats, without first testing this experimentally. Similarly, it is not clear whether Alopecurus pratensis, Poa palustris or other indicators of damp or wet conditions cannot also thrive on much drier soils. As discussed in Sect. 7.5.2, their amplitude in grassland communities is highly dependent on the presence of stronger competitors. 8.4.2.3  Flooding Tolerance Similarly to the plants of reedbeds, sedge communities and mires, the species of flooded grasslands and many plants of damp and wet meadows must endure oxygen deficiency around their roots when the groundwater is high or if they are flooded. Partial or total lack of oxygen (i.e. hypoxia or anoxia) in the root zone causes not only changes in metabolism (E.  Beck 2002), but also promotes the formation of aerenchyma in shoots and roots as well as elongation of internodes (see Sect. 8.4.1). These two morphological adaptations improve the oxygen supply to the plant and allow root growth to continue as well as the formation of lateral roots in wet soils. The degree of tolerance to anoxia in the root zone determines the lower limit of a species in the flooding gradient in a floodplain (Laan et al. 1989, Silvertown et al. 1999 and Lenssen and de Kroon 2005), whereby the upper limit is determined mainly by drought. Of six Rumex species, only the riverbank species R. maritimus always had a well-developed aerenchyma, which is constitutional and is formed in this species through division of existing tissue and cell divergence. It was also ­frequently formed in the damp meadow species R. conglomeratus and R. crispus. Aerenchyma was only produced on wet soils by R. obtusifolius, and never by R. acetosa and R. thyrsiflorus, which are limited to drier soils (Laan et al. 1989). Deschampsia cespitosa similarly tolerates water saturation and anoxia as well as the related high concentrations of iron and manganese ions due to its ability to oxidise the rhizosphere through a well-developed aerenchyma (Davy and Taylor 1975; Davy 1980). The riverbank species Rumex palustris, which also has a well-developed aerenchyma, forms a much more extensive root system in wet soils than R. acetosa and Plantago major, both of which do not form aerenchyma and are intolerant of flooding (Engelaar et al. 1993). However, the well aerated root tissue of R. palustris is so soft that it is less able to penetrate compacted soils (see Fig. 8.43). The robust roots of Plantago major, which generally grows in trampled areas, are much better in this respect, making it competitive on compacted soil but not under constantly wet conditions. The roots of R. acetosa are not quite as robust as those of the two other

8.4  Adaptations to the Environment

683

Fig. 8.43  A comparison of the root systems of a plant of trampled areas (Plantago major), a marsh plant (Rumex palustris) and a meadow plant (Rumex acetosa), which were each cultivated in pots with glass tubes for observation of the root development. The three treatments were loose, well-­ drained soil (control), evenly compacted but well-drained soil (compacted) and loose but constantly wet, hypoxic soil (wet). The saturation with water began 21 (Plantago) or 14 days (Rumex) after planting. Simplified from Engelaar et al. (1993; with permission of John Wiley and Sons). The numbers on the outer edge of the pots show the percentage of root dry weight at the end of the experiment in horizontal (left) and vertical (below) sections. The values are the average of 7–8 replicates. All three species grow well on damp, loose and nutrient-rich soil and fill the pot with roots. Plantago also grows deep into the compacted soil, but does not tolerate wet soil. R. palustris tolerates the hypoxic conditions in compacted and wet soil, and R. acetosa grows poorly in both

species and it does not form aerenchyma, which is why it is only competitive on mesic and uncompacted soils. Short periods of flooding do not damage most species of flooded grassland. For example, Phleum pratense, Festuca pratensis and Alopecurus pratensis are consid-

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8  Agricultural Grassland on Mesic to Wet Soils

ered tolerant as they quickly recover. The fertile grassland species Arrhenatherum elatius, Lolium perenne, Dactylis glomerata, Poa pratensis and Festuca rubra, in contrast, are sensitive to flooding (Klapp 1971, cf. also Speidel and van Senden 1954; Stählin 1957; Raabe 1960). This was shown in a groundwater experiment, in which Arrhenatherum and Dactylis reached their maximum productivity at a much higher groundwater level in the damp summer of 1953 than in the dry summer of 1952. Groundwater levels only 50 cm below the surface led to a clear yield reduction in these species, but not in the flood-tolerant Alopecurus (see Fig. 8.44). Festuca rubra, which is also widespread in dry grasslands, forms ecotypes in salt and brackish marshes that are highly tolerant of regular flooding (see Sect. 1.4.1). Agrostis stolonifera is also extremely variable in form, and is adapted not only to conditions in flooded grassland (Misra and Tyler 2000), but also occurs in agricultural grasslands, some meadows, and in arable weed and other communities. These are different genotypes that differ in the length of their stolons, the height of the flowering stem, as well as their flooding, drought and salt tolerance (Aston and Bradshaw 1966). 8.4.2.4  Nutrient Demand and Nutrient Limitation Fertilisation experiments have shown that not only nitrogen and phosphorus but also potassium can limit the productivity of meadows and pastures in Central Europe. N limitation occurs particularly in meadows on acid substrates with low mineralisation rates, if the nutrient loss through mowing is not replaced by fertilisation (Künzli 1967; Vermeer 1986; Stöcklin and Gisi 1989). On soils with low cation exchange capacity such as sand or peat, potassium limitation can also occur, particularly if there are high element losses through leaching and harvesting (Oomes and Mooi 1985; Sach 1997; van Duren et al. 1997; Olde Venterink 2000). Phosphorus can be limiting in some calcareous, as well as acid meadows and pastures, if it is bound by high concentrations of Ca or Fe. This also occurs during the drainage of damp meadows, when Fe2+ is oxidised to Fe3+. Some examples of this are the Junco-­ Molinietum and communities of the Caricion davallianae (Egloff 1983; Olde Venterink 2000) as well as weakly fertilised and sometimes flooded Cnidion meadows in floodplains (Franke 2003; Härdtle et al. 2006). In Dutch moist meadows, the dominance of Molinia indicates P limitation (Olde Venterink 2000), whilst the dominance of Holcus lanatus shows K limitation, and that of Glyceria maxima, Phragmites australis or Carex acuta shows N limitation. Schaffers (2000b) states that N:P ratios of less than 10 in the shoot biomass indicate N limitation in grassland plants, whilst values above 14 show P limitation. Low foliar nitrogen concentrations have little effect on the cell size in grass leaves, but reduce the rate of cell division, thus limiting plant leaf area and the leaf area index (Fricke et  al. 1997). High yields in intensive grasslands are therefore linked to high levels of N fertilisation, which guarantees high LAI and maximum light capture. Generally, the root growth of grassland plants is less affected by nutrient and water limitation than shoot growth (Gastal and Durand 2000). Low phos-

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685

Fig. 8.44  Dry matter production in several meadow grasses in monoculture on sandy soil under different groundwater levels that are maintained constant within each treatment, in the dry summer of 1952 and (after reseeding) in the relatively wet summer of 1953 (cf. Fig. 7.30) (From Ellenberg 1954a) After the low rainfall of 1952, the production of all species was highest at relatively high groundwater levels, because the topsoil often dried out and contained little water at the end of the growing season. This kind of situation is relatively common in southwestern Germany, and allows the species of the Arrhenatheretum to also colonise soils with high groundwater levels (Fig. 8.16 bottom). Frequent rainfall, as occurred in 1953, reduces the growth of Arrhenatherum, Dactylis and other deep-rooting grasses, because they tolerate the resulting hypoxia in the soil less well than Alopecurus and other damp meadow species, the root cortex of which contains air-filled aerenchyma. In the comparatively humid climate of northwestern Germany, this is so often the case that the Arrhenatheretum is limited to soils with low water tables (Fig. 8.16 top) The shallower-rooting Bromus erectus reacted surprisingly little to the wet conditions, thus not behaving as a ‘xerophilic’ plant in this experiment (cf. Figs. 7.27 and 7.31)

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8  Agricultural Grassland on Mesic to Wet Soils

phorus levels also may lead to a reduction in leaf area. For example, Kavanova et al. (2006) found a reduction of 39 % in Lolium perenne. P limitation thus leads to particularly low grassland productivity (Olde Venterink 2000). The grass species of Central European meadows and pastures differ considerably in their genetically determined potential growth rate. If different grass species are cultivated under constantly high moisture and favourable nutrient conditions, then e.g. Poa annua and Holcus lanatus grow three times faster than Danthonia decumbens and Nardus stricta (Grime and Hunt 1975; see Fig. 8.45). The relative growth rates of grass species increase with the nitrogen availability in their natural habitat (as expressed in the indicator value for N). Bradshaw et al. (1960, 1964) found a particularly large increase in growth rate upon N and P fertilisation in Lolium perenne, followed by Agrostis stolonifera, A. capillaris, Cynosurus cristatus, Festuca ovina and Nardus stricta. Although moderate N fertilisation promotes the growth of Nardus, very high rates actually reduce it (Bradshaw et al. 1964; see also Sect. 6.4.2). High soil P concentrations do not significantly affect the growth of Nardus (Chadwick 1960).

Fig. 8.45  The relative growth rates of 31 Central European Poaceae species under optimum light, moisture and nutrient conditions in a greenhouse. Indicator values for N are also given from Ellenberg et al. (1992). Generally, grasses of nutrient-rich habitats have a higher growth potential (Unpublished data from Ellenberg)

8.4  Adaptations to the Environment

687

The fast-growing grasses of nutrient-rich grasslands have a higher N demand and synthesise more transport proteins for N uptake in their root membranes than species of nutrient-poor grasslands (Chapin 1980; Larsson 1994). However, the latter have a higher nitrogen use efficiency (NUE), i.e. they use N more economically in growth than the former (Aerts 1989b) and have longer-lived leaves and roots (Schläpfer and Ryser 1996). Based on leaf nutrient concentrations, meadow forbs appear to have a greater ability to take up potassium, phosphorus and calcium than grasses, with on average 50–100 % higher contents in their leaves (Briemle 1998b). Central European grassland plants use not only nitrate and ammonium as an N source, but can, like many plants of acid and cold habitats, also take up organic N in the form of amino acids (e.g. glycine, alanine or glutamine) (Weigelt et al. 2005). The experimental results of Falkengren-Grerup et al. (2000) showed that the proportions of organic N uptake are very low in forbs such as Prunella vulgaris and Galium aparine, but can reach 20–50 % of ammonium uptake in grasses (Poa, Agrostis, Festuca and Elymus species) or even more in the case of Deschampsia flexuosa. Some grassland plants such as Trifolium repens form a range of ecotypes that are physiologically adapted to different soil chemical conditions such as might occur in different patches within a stand (Bradshaw 1976). Snaydon (1962a) and Snaydon and Bradshaw (1962) found that the Trifolium repens clones from phosphate-rich soil grew less well under low P conditions than clones from phosphate-poor soil. This is probably due to the lack of high-affinity P uptake systems in the root membranes that efficiently can extract P from the soil when the concentrations are low (cf. Lambers et  al. 2008). Such adaptations may also be linked to the different growth rates observed in calcareous ecotypes and acid ecotypes on basic and acidic soils. Whilst the former grows poorly on acidic soils, the growth rates of the latter are more or less the same on both acid and calcareous substrates (Snaydon 1962b). Trifolium repens thus shows an intraspecific differentiation of calcicole and calcifuge ecotypes. A transplantation experiment with numerous genotypes of Trifolium pratense, Dactylis glomerata and Plantago lanceolata with different geographical origins showed that each genotype grew best in its original habitat, i.e. it was adapted to the specific conditions here (Joshi et al. 2001). In an experiment with Trifolium repens, local genotypes grew better in the experiment when in combination with the matching Lolium genotype from their original habitat than with a random Lolium genotype (Lüscher et al. 1992). The Trifolium repens genotypes must thus have adapted their genome over generations to their neighbouring plants in the community. 8.4.2.5  Heat and Cold Stress High summer temperatures can reduce growth in Central European grasslands. Maier (1971) found that many grassland plants are damaged in mid-summer even at temperatures below 50  °C, probably caused by the denaturation of proteins. Grassland species such as Galium mollugo (heat damage at 50 °C), Arrhenatherum elatius and Taraxacum officinale (44 °C), Poa pratensis (43 °C) or Bellis perennis

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(42 °C) have a relatively low heat tolerance. Sensitivity to high temperatures is probably linked to the simultaneous effects of soil desiccation and high evaporative demand. Temperate grasses such as Arrhenatherum seem to possess a stomatal regulation with high sensitivity to air humidity, which reduces photosynthetic carbon gain in dry summer periods and makes transpiration cooling ineffective (Stocker 1967; Ruetz 1973). The eastern and southern limit of Arrhenatherum meadows and related communities is therefore probably caused in part by the low heat tolerance of Central European meadow plants. The cold tolerance of many meadow plants is, however, high. For Bellis perennis, for example, it was −26 °C in the winter of 1961 and −20 °C in 1962 (Schnetter 1965). Lolium, Deschampsia cespitosa and Holcus are, in contrast, more sensitive. This can result in the spread of Phleum and Alopecurus, as well as Bromus hordeaceus, Ranunculus repens and others after hard winters and cold springs, because Lolium, Deschampsia and Holcus have been weakened (Klapp 1965). High potassium contents in the soil liquid phase increase the winter hardiness of grassland plants, whilst high N contents reduce it, as they facilitate early spring growth (Voigtländer and Jacob 1987). The photosynthetic apparatus of grassland plants is adapted to the low temperatures that occur in Central Europe not only in spring and autumn, but also often in the early morning hours in summer. The compensation point is exceeded even at 0 °C, i.e. often whilst frosty leaves are just thawing. Species that are more common in mountain meadows than in the lowlands, e.g. Festuca rubra (red fescue) as studied by Ruetz (1973), achieve their maximum light-saturated net photosynthetic rate often at less than 10 °C, whilst in other species this is at most 25 °C (Kalckstein 1974). Cooper and Tainton (1968) found that many grassland plants begin to grow above temperatures of 5 °C, and reach their optimum growth at 17–21 °C. A strong decline in light saturated photosynthesis and growth occurs sometimes already at 20 °C (Ruetz 1973), and usually at 25 °C and above, until it completely stops at 30–35 °C (Rüegg 1976). Low light levels reduce the optimum temperature for photosynthesis in Lolium perenne (Schäfer 1972). In this respect, Central European meadow plants are therefore well adapted to the often cloudy weather they experience.

8.5  Population Biology and Community Ecology 8.5.1  Phenology On deep, mesic and fertile loam soils, Arrhenatheretum meadows remain largely green during normal winters, and resume growth immediately after snowmelt. Before the grasses grow too high, the countless flowerheads of Cardamine pratensis (cuckoo flower) spread a light violet veil over the deep green. In the period between around mid-April and early May, the meadows are dotted with yellow from

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689

Taraxacum officinale (dandelion) and gleaming Ranunculus acris (meadow buttercup). Later, the large golden yellow flowers of Tragopogon pratensis (salsify) reach up into the tall grass stalks. As Leucanthemum vulgare (oxeye daisy) opens its white flowers in May or early June, the umbels of Anthriscus sylvestris (cow parsley) stand tall above their feathery leaves and the white, foamy flowers of Galium beneath, whilst Dactylis glomerata (cock’s foot) and Arrhenatherum elatius (false oat-grass) begin to flower. The grasses then finally dominate the appearance of the meadow with their monotonous green, tinted slightly silver, gold or violet by their flower spikes. Suddenly, on a sunny morning in May or June, the whole resplendent stand disappears under the mowing machine. The tall grasses are reduced to pale stubs, whilst the lower grasses, legumes and low-growing forbs recover from the earlier period of deep shade, until they are again covered by tall grasses and forbs. Before the second cut, Crepis biennis (rough hawksbeard) and Heracleum sphondylium (hogweed), and in places also Pimpinella major (greater burnet saxifrage) and Pastinaca sativa (parsnip), develop. In some regions, Campanula patula (spreading bellflower) and Geranium pratense (meadow cranesbill) with their violet or pure blue are also found in mid-summer. However, before the stems and leaves become straw-like, the meadow is cut a second time in mid-summer, and it is only in warmer climates that the tall-growing species flower for a third time (cf. the phenological spectrum of flowering in Dierschke and Briemle 2002). Typical Cirsium oleraceum communities on moist soil are dominated in springtime by the flowers of Cardamine pratensis and Caltha palustris (marsh marigold). Later these are replaced by the crimson, rust-coloured, pink and yellow flowers of Lychnis flos-cuculi, Ranunculus acris or Polygonum bistorta, as well as the anthocyanin-­rich red buds of Rumex acetosa and Holcus lanatus (see Fig. 8.22). Some Calthion meadows in occasionally inundated floodplains are particularly richly decorated with Fritillaria meleagris (snake’s head fritillary) with its chequered flowers, which according to Horsthuis et al. (1994) is originally a floodplain species. These meadows have almost no pure white or blue flowers, even in the later grass-dominated stages. After the first cut, Cirsium oleraceum (cabbage thistle) extends its yellow-green and soft-leaved stem and begins to flower. In places, Angelica sylvestris finally provides a white tint to the meadow. The second cut often yields as much as the first, because this community begins to develop later in spring than the Arrhenatheretum and because the soil remains moister after cutting. Compared to the deep green and flower-rich fertilised moist meadows, the Molinia litter meadows remain a pale straw colour well into spring. In mountain areas, Anemone nemorosa (wood anemone) or Primula elatior (oxlip), i.e. technically forest plants, often take advantage of this spring pause. It is only long after the first cut of the lowland hay meadows that the blue-budded stems of Molinia (purple moor grass) begin to grow and a few of the other species begin to flower. In late summer, the dark blue bells of gentians, the yellow-white umbels of Silaum silaus (pepper-saxifrage), the crimson heads of Serratula tinctoria (saw-wort) and other colourful flowers finally produce an impressive show. Once the stems and leaves of

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Molinia start to turn golden yellow and copper coloured, the areas of litter meadows shine out from the surrounding ‘evergreen’ hay meadows. The majority of meadow species flower 3–4 weeks later in the montane belt than in the lowlands. Some exceptions to this rule are Alopecurus, Carum, Colchicum, Cynosurus and Lolium (Nowak and Schulz 2002), in which the induction of flowering is apparently not driven by temperature, but rather by photoperiod and is therefore largely independent of elevation. Under traditional management with first cut in mid June, almost all meadow species will come to flower. However, if the meadow is first cut for silage in the second week of May, then this will be before the main flowering period of 85 % of species as observed in the Black Forest by Nowak and Schulz (2002). These include many species that do not, or barely, flower twice. Although early mowing can stimulate a second flowering, many species are unable to produce seed in meadows with early cutting. The flowering phenology of Central European meadows described here has been divided by Dierschke and Briemle (2002) into nine phenological stages, on which the classification of 242 agricultural grassland species into phenologically similar groups is based. From spring to autumn, these are: (1) the flower-poor pre-stage, (2) the Anemone nemorosa-Primula stage, (3) the Cardamine pratensis-Taraxacum officinale stage, (4) the Ajuga reptans-Alopecurus pratensis stage, (5) the Anthriscus sylvestris-Ranunculus acris stage, (6) the Leucanthemum-Lychnis flos-cuculi stage, (7) the Cirsium palustre-Galium album stage, (8) the Centaurea jacea-Filipendula ulmaria stage, and (9) the Colchicum autumnale stage.

8.5.2  Seed Bank, Germination and Dispersal Vegetative reproduction via rhizomes, stolons and suckers plays an important role in many agricultural grassland species. The rootstocks and rhizomes of the common grassland species can often be quite long-lived; Klapp (1971) estimates a maximum age of up to 300 years. Daughter plants remain functionally linked for a long time and exchange assimilates, so that the growth of shaded shoots or those that have lost their leaves is supported by more productive ramets (Ong and Marshall 1979). Diaspore Bank  Although Central European grassland soils often contain high densities of seeds, the importance of the seed bank for the population dynamics of the species involved appears to be relatively low (Milberg 1992). This is partly to do with the fact that intensive cutting and grazing regimes reduce the seed bank, because fewer flowers are able to form and mature (Archer and Rochester 1982). Various authors have studied meadows with between 7000 and 57,000 germinable seeds per m2, mostly in the uppermost 2 cm of soil. Pastures and young meadows have lower values. In some seed bank studies, more than 100 species were found, of which many did not occur in the community (Klapp 1971). In contrast, typical grassland species were often absent from the seed bank. Based on the evaluation of Thompson et al. (1997), Dierschke and Briemle (2002) classified 267 agricultural

8.5  Population Biology and Community Ecology

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grassland species in terms of the length of viability of their seeds. This showed that almost two-thirds of the assessed grassland species have no permanent seed bank, i.e. their seeds remain viable for at most 1 year, and often only a few weeks (transient seed bank). Roughly 20 % of the species produce seeds viable for up to 5 years, and another 20 % longer than this (cf. also Bekker et al. 1998). Flooding appears to reduce the seed bank of floodplain meadows (Hölzel and Otte 2004). Nevertheless, some species of grasslands do not follow this general pattern, as for example the seeds of Taraxacum officinale, Trifolium repens and Rumex crispus have been recorded as remaining viable for over 50 years (Thompson et  al. 1997). In wet meadow fallows, Rosenthal (2001) also found relatively long-lived seed banks. Dispersal In addition to the short-lived seed bank, also the short dispersal distances of seeds can reduce the regeneration of meadows and pastures via sexual reproduction. Fischer (1987) analysed seed traps in Arrhenatheretum meadows, and found that 97.5 % of the seeds originated from within a distance of 1 m. Similarly, Jensen (1998) found that 80 % of the species of wet meadows and abandoned meadows had a dispersal distance of less than 1.5 m. This is partly due to the low height at which the seeds of grassland plants are released, and the resulting aerodynamic disadvantage (Greene and Johnson 1986). Wind dispersal of seeds should therefore be relatively unimportant in meadow species, because there they are unlikely to find a gap suitable for germination in the dense stands (Grubb 1977). In wet meadows, diaspores were once deposited in large numbers by flooding (e.g. Hölzel and Otte 2001), but the dyking of rivers has since much reduced this means of dispersal. Some species such as Bromus and Thymus are often dispersed by ants. However, over two-thirds of the seeds in a meadow will be consumed by mice (Hulme 1993). The most important dispersal vector in Central European agricultural grasslands is probably humans and their livestock, at least when considering long-­ distance dispersal. A detailed study of Bonn and Poschlod (1998) shows that the management of litter and hay meadows as well as pastures must have significantly contributed to the dispersal of grassland species, and probably was largely responsible for their species richness. Seed input from surrounding habitats thus plays a larger role in determining the course of succession in abandoned meadows than the local seed bank (Bekker et al. 2000). The flower-rich meadows with low-intensity management produce high densities of diaspores, e.g. up to 4000 seeds per m2 and cut in Molinia and Cirsium rivulare meadows (Biewer in Bonn and Poschlod 1998), which were once transported around the local area with the litter or hay. Large quantities of diaspores were also spread on meadows and pastures with the manure, in which they could remain germinable for several weeks (Korsmo 1930). Boeker (1959) recorded 90 seedlings in 1 g of stall manure. The practice of diverting water for meadow irrigation, which was once widespread particularly in central and southwestern Germany (see Sect. 8.7.2), must have played a role in the regional transport of grassland diaspores (Bonn and Poschlod 1998). Also important was the long-distance transport via transhumant herds. The livestock carried the diaspores of pasture plants in their fur or digestive tract, depositing them also on meadows in aftermath grazing (i.e. after mowing was complete for the year).

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Germination  The dense swards of most pastures and meadows mean that the germination and successful establishment of young grass and forb plants is only possible in a few gaps and disturbed areas, produced by the activity of mice, rabbits and grazing livestock (Bullock et al. 1995; see also Sect. 7.5.2). Ruderal species such as Arabis hirsuta and Myosotis arvensis often germinate on small mounds of earth created by common voles (Microtus arvalis) in meadows (Gigon and Leutert 1996). Dactylis glomerata and Galium album are particularly promoted by the nutrient-­ rich excrement at the entrance of mouse or vole holes. Successful seedlings may remain in a vegetative state for a very long time and often flower only after 3–5 (and sometimes 10–30) years, because the competition from the established plants is so strong (Rabotnov in Klapp 1971). In contrast, Schreiber (2001) states that trees easily germinate in the dense mat of grass litter if the other conditions are conducive to their growth. Plants of productive grassland germinate faster than those of unproductive grassland (Olff et al. 1994), which are more dependent on winter stratification. The species that germinate in autumn and winter probably thereby avoid the more intensive competition for light in the nutrient-rich meadows, which is less important in the sparser vegetation of nutrient-poor grasslands. Plants of frequently disturbed trampled communities are particularly dependent on colonisation via seeds. These usually arrive in sufficient numbers from neighbouring areas, as plants of trampled areas are mostly epizoochorous or produce such small seeds that they can stick to the mud on shoes, hooves or wheels (see Table 8.12). Trifolium repens and other species are also dispersed by granivorous birds such as rooks (Krach 1959). Car tyres have been playing an increasingly important role in seed transport in recent decades (Clifford 1959; Schmidt 1989). Only 1.5 months after the creation of a new shoulder to a road, Runge (1970) noticed the first plants of Poa annua and Lolium perenne in May, and by August the Lolio-­ Plantaginetum was completely formed. This shows how dynamic the vegetation of trampled areas can be. Even if the seeds of a grassland plant have arrived in a new area, it is far from certain that it will be able to develop there. The establishment of species of trampled areas is particularly affected by the level of moisture at the soil surface, as the germination of some species such as Plantago major and Polygonum aviculare is tightly linked to wet conditions (Ellenberg and Snoy 1957). They thus benefit from the standing water that often forms over compacted soil. Grasses such as Poa annua also germinate during long periods of damp conditions, and Trifolium repens requires damp surface soil for its above-ground stolons to put down roots. Lolium colonises intensively trampled areas mainly via vegetative reproduction (Boeker 1959). The germination and seedling development of various Plantago species were found by Blom (1976, 1977) to differ in their response to experimentally increased trampling intensity. As expected, P. major was the most tolerant, followed by P. media. P. lanceolata suffered considerably more and P. coronopus was the most damaged.

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693

Table 8.12  Seed germination in highly compacted soil near a water trough The following number of seeds germinated from around 3 kg of soil collected in December and kept moist: Plants of trampled ground (sensu Arable weeds (sensu lato)a Meadow plants (sensu lato) lato) 37 Poa annua 10 Agrostis capillaris 5 Stellaria media 31 Plantago major 10 Poa trivialis 4 Juncus bufonius 15 Trifolium repens 6 Poa pratensis 1 Sonchus arvensis 13 Matricaria discoidea 3 Juncus articulatus 1 Sonchus oleraceus 10 Sagina procumbens 2 Cerastium fontanum 1 Urtica urens holosteoides 3 Capsella bursa-pastoris 1 Festuca rubra 2 Polygonum aviculare 1 Stellaria graminea 2 Veronica serpyllifolia 1 Achillea millefolium 1 Taraxacum officinale From data in Boeker (1959) Lolium perenne is absent here and also failed to germinate in parallel tests, even at different times of year, although it was common in the vegetation

a

8.6  Productivity and Cycling of Water and Nutrients 8.6.1  Productivity 8.6.1.1  Above-Ground Productivity The tall grasses and forbs of Central European meadows have a relatively high light-­ saturated net photosynthesis rate (15–20 μmol CO2 m−2 s−1; Gloser 1993). Low photosynthetic capacities were only found in the low-growing plants of the shady understory (e.g. Briza media and Luzula campestris: 3–8 μmol CO2 m−2 s−1). However, species-specific differences in photosynthetic capacity only explain a small proportion of differences in growth rates (Poorter and Remkes 1990). Slow-­ growing grasses such as Nardus stricta have similarly high light-saturated photosynthetic rates (17 μmol CO2 m−2 s−1) to some fast-growing nutrient-rich meadow species (Cooper and Tainton 1968; Gloser 1993). Meadows can have similar or even higher leaf area indices to forests. For example, Leuschner (1986) recorded an LAI of 9  m2 m−2 in a montane Trisetetum in mid-summer (cf. Geyger 1977). The leaf morphology, the spatial arrangement of the leaves in the stand (i.e. their interception of radiation; Sheehy and Cooper 1973) and carbon allocation patterns within the plant all also influence the productivity. Fast-growing grasses and forbs are usually particularly tall, and have very thin leaves with a large specific leaf area, the growth of which is promoted by high N and water availability (Garnier 1992; see Sect. 8.4.2). The low influence of photosynthetic capacity on growth rate is shown by the negative relationship between leaf size and photosynthetic capacity found by Evans (1976) in grasses. Measurements by Gloser (1993) in eastern Czech Republic showed that a Nardus grassland with

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moderate growth rates had a similarly high stand carbon gain to tall Phalaridetum arundinaceae or Glycerietum maximae reedbeds, although the Nardus stand produced only a third of the above-ground biomass of the reedbed. This was because the assimilates were invested in the Nardetum in the root system rather than in the expansion of the leaf area, which would have led to a further increase in carbon gain, illustrating the importance of plant-internal allocation patterns for grassland productivity. Meadows with large volumes of dead litter and little assimilating phytomass also have low growth rates. Most measures of productivity in agricultural grasslands are based on the dry weight of the hay yield, or the fodder consumed by livestock. These numbers vastly underestimate the annual net primary productivity of meadows and pastures, as they ignore the root production and only incompletely record the above-ground growth. Yield measurements can vary depending on the vegetation type, from less than 100 to over 1500 g dry weight m−2 year−1 (i.e. < 1 to >15 Mg ha−1). Dry and nutrient-poor grasslands, Nardus grasslands and some small-sedge communities have very low yields (c. 100–300 g m−2 year−1) as well as Molinia meadows and unfertilised pastures (c. 200–400 g m−2 year−1); dry Arrhenatheretum meadows with little or no fertiliser application produce roughly 200–450 g m−2 year−1. Mesic Arrhenatheretum, Angelico-Cirsietum oleracei and other wet meadows with low levels of fertilisation produce c. 400–600 g m−2 year−1. Yields over 600 g m−2 year−1 are only possible in agricultural grassland through intensive fertilisation (see Table 8.13); modern frequently mown meadows and mown rotational pastures produce 1000–1200 g m−2 year−1. Similarly high above-ground productivities have also been found in unmanaged Magnocaricion and reedbed communities, i.e. grassland communities in a loose sense, with year-round high water availability (see Fig.  8.46). In contrast to some unproductive small-sedge communities that occur on wet soils with little water movement, the high productivity of the Magnocaricion is likely to be due to the nutrient inputs from the slow-moving water. Role of Water  The arrangement of grassland communities along a gradient of soil moisture in Fig. 8.46 highlights the importance of the water supply for the yield: from dry grasslands to intensive pastures and Calthion meadows, water limitation decreases and productivity increases. Correspondingly, Rychnovská (1993) found a linear relationship between the productivity of Central European meadows and their stand transpiration, which demonstrates the importance of water supply for productivity. Damp and relatively warm summers with optimal water suppy thus produce the highest hay yields. Role of Nutrient Availability  The base-saturation and pH of the soil have a significant influence on the species composition of grassland communities (Hundt 1966), whereby the species richness generally increases with pH (see also Sect. 7.1.1). The influence of this factor on the productivity of pastures and meadows is, however, much weaker than that of water. Nevertheless, the highest hay yields in unfertilised grassland in Upper Bavaria and the Harz Mountains are found at pH 6 and above (Hundt 1964; Spatz and Voigtländer 1969). Speidel (1962) found an

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Table 8.13  Net primary production of a Festuca rubra-Trisetum flavescens meadow at different nutrient levels in the Solling, 5 years after an application of fertiliser providing good nutrient levels Fertilisation: Net primary production (energy fixation in 105 kcal ha−1 year−1)   Green biomass (used):  Litter + stubble (above-ground, unused)   Roots without finest rootsa  Totala  Green biomass in relation to total biomass (%) Effectiveness of net primary production  In % of annual global radiationb  As above, in the growing season (%)

0

PK

NPK

98 194 140 432 22.7

204 220 134 558 36.6

306 198 110 614 49.8

0.54 0.77

0.62 0.88

0.82 1.14

From data in Runge (1973) The finest roots are not detected in normal root measurement procedures, but make up a much larger proportion of the total production under low than under high nutrient conditions. The sum of the net primary production is therefore probably roughly the same at all nutrient levels (approximately 700–800 105 kcal ha−1 year−1, cf. Fig. 4.83 in Vol. I) b When photosynthetically active radiation is considered, the percentage is roughly twice as high a

increase in the hay yield of unfertilised mesic meadows in Hesse of 70–80 g m−2, and in wet meadows of even 150 g m−2, after increasing the base-saturation of the soil by 10 %. However, Schmitt and Brauer (1979) state that higher hay yields on clay soils compared to sandy soils in Hesse are mainly caused by their hydrological properties and less by their chemistry. If high levels of fertiliser are applied, then the influence of soil pH and base richness on the yield disappears (Speidel 1962). Drought can induce nutrient limitation and thereby reduce the growth of meadow plants in mid-summer (Onillon et al. 1995; see Sect. 8.4.2). Garwood and Williams (1967) added N fertiliser in small areas in the damp subsoil of a meadow with dry topsoil, showing that the low growth rates were less caused by lack of water and more linked to drought-induced N limitation. The clear relationship between productivity and soil moisture in Fig. 8.46 could therefore also be a result of differences in nutrient supply, as the nitrogen mineralisation increases from dry to damp soils due to the higher microbial activity (see Fig. 8.50). In support of this, many ­semi-­dry grasslands can be turned into Arrhenatheretum meadows by fertiliser appliation without irrigation. Role of Species Richness  Experiments with sown grass and forb mixtures have shown that species-rich stands have a higher productivity than species-poor stands (e.g. Tilman et al. 1996; Hector et al. 1999; Roscher et al. 2005; but see Kenkel et al. 2000). According to these findings, diversity positively influences the productivity of grasslands, at least at low levels of species richness (i.e. 100 years of habitat continuity (Krause and Culmsee 2013). In the southern Black Forest, nutrient-poor meadows needed at least 50 years to show saturation in species number (Nowak and Schulz 2002). These findings demonstrate the essential role of historically old grasslands for conservation.

8.7.2  Succession in Abandoned Meadows If the biomass removal through mowing or grazing ceases due to grassland abandonment, within a few years the structure and species composition of the vegetation will start to change, accompanied by changes in the soil and the stand climate. Wet meadows of the Calthion usually quickly transfom into species-poor stands dominated by tall herbs such as Filipendula ulmaria, Phalaris arundinacea, Glyceria maxima or Scirpus sylvaticus, and in upland areas also by Polygonum bistorta. Damp pastures are often dominated by Deschampsia cespitosa, Juncus effusus and J. conglomeratus. Wet areas with high groundwater table become covered with tall sedge species (Carex acutiformis, C. acuta, C. paniculata, C. vesicaria etc.). These

708

8  Agricultural Grassland on Mesic to Wet Soils

are joined by nitrogen-demanding plants such as Urtica dioica, Cirsium arvense, Galeopsis tetrahit (e.g. Rosenthal 1992; Schwartze 1992; Jensen 1997; Schrautzer and Jensen 1999; Dierschke and Briemle 2002). A long-term study by Falinska (1995) showed that Filipendula first spread into abandoned moist meadow clonally, and after around 12 years the daughter plants began to separate as their rhizomes disintegrated, allowing other meadow species to become competitive. In abandoned Molinia litter meadows in the Allgäu (Bavarian Alps), Abt and Ege (1993) recorded an initial increase in Molinia, but a decrease in older fallows. Arrhenatheretum meadows and Festuca rubra pastures in upland areas tend to be encroached by tall herbs such as Hypericum maculatum or Trifolium medium, which are then soon replaced by tall herbs with large storage organs such as Geranium sylvaticum, Heracleum sphondylium or Galeopsis tetrahit that alternately dominate the stand (e.g. Neuhäusl and Neuhäuslová-Novotná 1985). The colonisation of shrubs and trees is largely driven by the environmental conditions. Under favourable climatic and edaphic conditions, Schreiber (1993) found bushes up to 2 m high in abandoned meadows in southwestern Germany after only 5 years, whilst in damp to wet areas or in meadows with a thick litter layer or in the montane belt, there was almost no growth of woody species (cf. also Kienzle 1979 and Wolf 1980). The pioneer tree species were almost all wind dispersed, whilst the shrubs spread via suckers or their seeds were deposited by birds. Browsing by wild animals and insect herbivory can considerably delay scrub development (Müller and Rosenthal 1998). Chance also plays an important role in determining which of the many possible courses of succession abandoned grassland will take (Schreiber 2001). The decline in species richness over time in abandoned grasslands has been frequently documented in permanent plot studies (e.g. Wolf 1979; Oomes and Mooi 1985; Bakker 1989; Hachmöller 2000). This decline is particularly strong for low-­ growing species, especially if they mainly reproduce by seed. These plants are not only disadvantaged by shading from the taller vegetation, but also by the rapidly accumulating litter layer, which effectively prevents germination and seedling growth (Jensen 1997). This is the case for many grasses and forbs, but apparently not for tree seedlings, which are only occasionally hindered by thick mats of litter (Schreiber 2001). The number of species present in the seed bank also decreased with time after abandonment in a grassland in eastern Poland (Falinska 1999). In contrast, the number of species with clonal reproduction generally increases, particularly of those with rhizomes (Prach and Pysek 1994). The behaviour of numerous species after the abandonment of meadows and pastures in Sweden was analysed by Ekstam and Forshed (1992). The increasingly dense vegetation in abandoned grasslands leads to a cooling of the air close to the soil surface, and of the soil itself, by several degrees compared to a managed meadow, as well as a reduction in the compaction of the soil (Gisi and Oertli 1981a,b). Organic matter accumulates mainly at the soil surface as litter and in the biomass, but not in the soil as generally occurs in abandoned arable fields. In the abandoned montane grasslands in the Swiss Jura, Gisi and Oertli (1981a,b) found a 70-fold increase in litter production. As a result, the C:N ratio of the litter layer increased while the pH of the soil decreased. In most cases, the microbial

8.8  Human Influence

709

activity of the soil also sinks after abandonment, and therefore also the mineralisation rate (Schreiber 2001). The frequently observed establishment of plants with high N demand in abandoned grasslands, such as Urtica or Aegopodium, must therefore be mainly caused by an increase in the internal nutrient cycling of the individual plants rather than higher mineralisation rates, i.e. ‘auto-eutrophication’. In some abandoned nutrient-poor grasslands, the N mineralisation rate can also increase (Hartmann and Oertli 1984), whilst grassland succession on calcareous soils can continue for long periods without any changes to the soil conditions and without strong accumulation of humus (Broll 1996). The abandonment of wet meadows is generally associated with blockage of the drainage system and therefore an increase in soil moisture (Amani 1980), which leads to a decline in N mineralisation and nitrification rates (Broll and Schreiber 1985; Olde Venterink 2000). Consequently, the above-ground net primary production can increase in abandoned damp meadows by up to 100 % whilst the N mineralisation rate sinks (Rosenthal 2001). If the water flooding the soil is lime-poor, then this can lead to higher P availability, which increases the productivity of the unmanaged grassland (Koerselman et al. 1993).

8.8  Human Influence 8.8.1  T  he Effects of Management Intensification and Fertilisation Management Intensification  Since around the 1960s, in much of Central Europe an unprecedented structural change in grassland management has happened, which has led to a significant floristic impoverishment and homogenisation of agricultural grassland communities (see Table 8.17). In most lowland areas, species-rich grassland communities are now no longer found, or only in small fragments (Hundt 2001; Oppermann et al. 2009; Wesche et al. 2009; Krause et al. 2015). This is also the case on the Swiss Plateau, where today only 2–5 % of the area used to produce livestock fodder supports species-rich Arrhenatheretum communities, compared to 89 % in 1939 (Dietl 1995). In 2001, only around 20 % of the managed grassland in Germany was assumed to be species-rich, mostly in the uplands (Schumacher 2005). In the upland and mountain areas of Central Europe, management intensification proceeded more slowly and around 20–40 % of the grassland area still remains under generally extensive use (Hundt 1996; Poschlod and Schumacher 1998; Hachmöller 2000; Schumacher 2005). This fundamental change was driven by many factors closely related to the ­transition from non-mechanised to industrial farming on grassland, including: –– continued application of high levels of NPK fertiliser or manure, –– background eutrophication from atmospheric deposition and/or via the groundwater,

710

8  Agricultural Grassland on Mesic to Wet Soils

Table 8.17  Relative frequency of the major grassland types in Germany between 1950 and 2000 (in %) Type of grassland Arrhenatheretum (lowland, submontane) Trisetetum (montane) Montane pastures, not manured Sum of extensively used mesic sites Nutrient-poor Salvia-Arrhenatheretum Damp, species-rich foxtail meadows Moist meadows, small-sedge swamps Sum of hydrologically more extreme sites Frequently mown grassland Species-poor foxtail meadows Highly fertilised pastures Sum of intensively used mesic sites

1950 35 10 10 55 10 20 5 35 3 2 5 10

2000 5 5 5 15 5 4 1 10 55 10 10 75

From Briemle et al. (1999)

–– –– –– –– –– –– –– ––

earlier first cutting in meadows, multiple grass cuts per year instead of one to two, silage production instead of hay production, rotational grazing with higher stocking rates instead of continuous grazing, drainage of wet areas, ploughing and reseeding of species-poor or single-species seed mixtures, use of herbicides against weeds, land consolidation, i.e. the joining of many smaller fields to make several large ones.

Other rapid changes in recent years have been caused by the general decline in the profitability of livestock grazing systems, which has been widely replaced by indoor systems. Formerly intensively grazed or mown grassland has thus been abandoned, particularly in marshes and moorland, where the maintenance of drainage systems is too expensive. A considerable proportion of formerly managed grassland has been converted to arable fields, whilst the remaining grassland areas have been intensified. Lowland pastures are now so well fertilised and so intensively grazed, that the species composition of the sward is primarily determined by these two factors. The same is true for the highly productive meadows, which have been homogenised in terms of their habitat conditions and flora through fertilisation and management for silage production. Grassland management in Central Europe is tighly linked to livestock farming, and in particular dairy farming. The now almost ubiquitous high-production cows require pastures and meadows with very high fodder values. The phytomass must be easily digestible and have a high energy density and protein content with little crude fibre. Fodder with an energetic value below 5.0 mJ net energy for lactation

8.8  Human Influence

711

(NEL) per kg dry matter is of little use in modern dairy farming (Dierschke and Briemle 2002). Only a Lolium grassland cut four to five times a year with an early cut in mid May has sufficiently high energy and protein density to satisfy the needs of high production cattle. Intensification thus not only allows an increase in yield (although this actually declines at higher cutting frequencies; Korte and Harris 1987), but also the production of silage (see Fig.  8.52). Nevertheless, hay and silage still have a lower energy concentration than fresh grass. Earlier mowing and shorter intervals between cutting also increase the fodder quality of the hay. The high demands of modern livestock farming thus cannot be met by moderately fertilised Arrhenatheretum meadows cut only twice a year. Instead, they need highly productive meadows with frequent cutting, mown pastures or even sown stands of Lolium that have to be intensively fertilised with 150 to over 350 kg N ha−1 (see Fig. 8.53). Feeding with maize and soya has also increased considerably in recent decades. Effects of Fertilisation  The application of nitrogen, phosphorus and potassium fertiliser has profound effects on the species composition of the grassland. Unbalanced N inputs, such as occur with manure spreading, promote tall grasses and forbs such as Heracleum sphondylium, Anthriscus sylvestris, Ranunculus acris and Rumex acetosa and reduce the proportion of legumes. Numerous N fertilisation experiments synthesized in Klapp (1962) show that only a small minority of grassland species are promoted by fertilisation. This included 14 of the 21 grass species present there, but only 1 of 10 legumes and 5 of 46 ‘other’ species (mostly forbs).

Fig. 8.52  Dry matter yield and crude protein content of meadows and pastures in the Rhineland under differing use intensity (From Klapp 1971)

712

8  Agricultural Grassland on Mesic to Wet Soils

Fig. 8.53  Change in quality of meadow fodder with increasing age (From Rieder (1997, in Dierschke and Briemle 2002))

The negative or positive effects on species were largely dependent on the habitat-­ specific species combination and therefore varied (see Fig.  8.54). The higher the nutrient availability, the clearer the difference betweeen the winners and losers of interspecific competition, i.e. the asymmetry of competitive relationships increases from nutrient-poor to -rich grasslands (Grime 1979; Schippers et al. 1999). All valuable fodder species in modern grassland management depend on high N availability. In mesic meadows, mineral NPK inputs particularly promote Arrhenatherum, but also Trisetum, Festuca pratensis, Dactylis, Poa pratensis and P. trivialis, whilst in damp meadows it is particularly Alopecurus pratensis in addition to these species that profits. Slurry spreading creates grass-poor meadows dominated by tall forbs.

8.8  Human Influence

713

Fig. 8.54  The effects of additional applications of nitrogen (left) and potassium (right) on the yields of pure and mixed stands of meadow plants near Zurich under otherwise normal levels of nutrients (Modified from Hofer 1970). Yield is in g dry matter per m2 The monocultures are marked with hatching. Over the 3 years of the experiment (1964 to 1966), all species were promoted by the addition of both N and of K. However, Taraxacum officinale and Ranunculus acris subsp. friesianus responded more to K than to N. Both nutrients promoted the growth of Rumex acetosa, but only moderately. The grasses had a higher dry matter yield with N as with K. The addition of one competing species caused the growth to change, depending on the competitiveness of the species in question. Dactylis glomerata won in most cases, as a tall and quick-growing grass. Only Taraxacum and Ranunculus were able to outcompete it under additional K fertilisation. The low-growing Festuca rubra outcompeted only Rumex under additional K, and under additional N only Rumex and Taraxacum. Rumex acetosa is indeed rare on well-­ fertilised meadows, whilst it is much more common on neglected grasslands with low fertilisation, together with the poor competitor Festuca rubra

Long-term application of high levels of mineral N fertiliser generally has a profound effect on the character of a grassland community. It was recognised very early on that fertilisation can turn a Mesobrometum or Salvia-Arrhenatheretum into a typical Arrhenatheretum (Stebler and Schröter 1887–1892; Aichinger 1933). The application of fertiliser can in these cases compensate for the lack of water. Fertilised Molinia meadows do not, however, develop into species-rich Calthion communities, but rather into species-poor Molinietalia fragmentary communities dominated by Cirsium oleraceum, C. rivulare or Alopecurus (Nowak and Schulz 2002). Moderate N fertilisation often initially leads to a decrease in yield in meadows (Klapp 1971), because the legumes are outcompeted by other species, so that their contribution to the productivity of the stand through N2 fixation is lost. If the stand is only fertilised with P and K, then the legumes usually increase. Under low phosphorus availability, Molinia, Nardus, Festuca ovina and others start to dominate (Klapp 1965). On the other hand, the growth of Molinia responds more positively to phosphorus than that of other grasses (Graf 1996).

714

8  Agricultural Grassland on Mesic to Wet Soils

The situation is similar in pastures: the more intensive the management, the lower the number of species, especially because Lolium and Phleum can form dense stands in the intervals between grazing periods and they shade out other species. Long-term mowing and intensive grazing promoted, alongside Lolium, particularly Poa trivialis, P. pratensis and Dactylis glomerata in the pastures around Göttingen (Ruthsatz 1970). Community Change and Diversity Loss  The effects of the rapid rise in synthetic fertiliser application and increasing mowing intensity after the Second World War were visible in the species composition of Central European meadow and pasture vegetation as early as the end of the 1950s (e.g. Tüxen 1955; Knapp 1969; Meisel and Hübschmann 1976; Rosenthal and Müller 1988; Böger 1991; Kornas and Dubiel 1991; Prach 1993; Hundt 1996, 2001). For example, Meisel (1970) compared vegetation surveys from various parts of the northwest German lowlands in the 1930s to ones carried out in the same areas 20 to 30 years later. Many indicators of low nitrogen concentrations (EIV-N values < 5, see Table 8.18) had become rarer and had a much lower constancy over all of the relevés. In their place, the indicators of moderate to high N (N 5 and above) increased, particularly those that are also very resistant to trampling and cutting. The group of declining species included the legumes, which lost a considerable proportion of their coverage. This was also the case for the highly constant Trifolium repens (N 6), which is clearly not overly sensitive to high nutrient conditions. These trends, shown in Table 8.18, have continued to accelerate, so that the Lolio-Cynosuretum grasslands are now extremely species-­poor and homogeneous. The numerous sub-communities identified by Meisel (1970) are difficult or even impossible to find today. In the uplands of northern and eastern Hesse (central Germany), the area of agricultural grassland declined by 30 % between 1950 and 2000 and lost on average (over 10 community types) 2.6 species per relevé (Raehse 2001). The largest losses were in the Sanguisorba officinalis-Silaum silaus communities and the Trisetetum meadows (losses of 8.3 and 6.3 species per relevé). Species of wet and nutrient-poor agricultural grassland suffered particularly strong declines in Hesse, which were often character or differential species of the grassland communities. Hundt (1996, 2001) documented the changes in the meadow communities of central and eastern Germany between 1964 and 1994. The average number of species in an Arrhenatheretum community declined from 39 to 13, with the winners being primarily Alopecurus pratensis and Elymus repens. Overall, the number of species with continental and subcontinental ranges increased at the expense of oceanic and suboceanic species. Zechmeister et  al. (2003) found a linear decline in species richness of vascular plants and bryophytes with increasing N fertilisation in Austrian meadows. Kornas and Dubiel (1991) followed the changes in Arrhenatheretum meadows in Ojców National Park in Poland over 30 years. The results are shown in Table  8.19, which contrasts the changes in meadows that received increasing amounts of fertiliser with meadows that are mulched by the park authorities for nature conservation purposes, and abandoned meadows (see also Sect. 7.8.3).

Indicators of nutrient shortage Species with formerly higher constancy: N  Grasses and graminoids: 2  Agrostis canina 2  Briza media 2  Nardus stricta 3  Luzula campestris 4  Festuca ovina 3  Cynosurus cristatus   Carex panicea Legume: 3  Lotus corniculatus Other herbs: 2  Centaurea jacea 2  Galium uliginosum 2  Hieracium pilosella 2  Pimpinella saxifraga 2  Succisa pratensis 3  Galium verum 3  Hypochoeris radicata 3  Leucanthemum vulgare 3  Rhinanthus minor 3  Stellaria graminea 4  Daucus carota a N = N indicator values, (−) = sensitive to trampling, (+) = highly tolerant of trampling ×  Vicia cracca (−) 5  Senecio jacobaea 6  Leontodon hispidus (−) 7  Galium mollugo

N    Legumes: ×    Trifolium pratense ×    Trifolium dubium (−) 6  Lathyrus pratensis Other herbs: ×  Prunella vulgaris ×   Veronica chamaedrys

Indifferent species and N indicators

(continued)

Table 8.18  Shifts in the species composition of Lolio-Cynosuretum pastures from 1935/40 to 1960/68 in the northwest German lowlands (from data in Meisel 1970)a

8.8  Human Influence 715

Species with higher constancy in 1960/68:

Indicators of nutrient shortage Species with both decreases and increases: Grasses: 3  Bromus hordeaceus 4  Agrostis capillaris

Table 8.18 (continued)

 

Grasses: 6    Festuca pratensis Phleum pratense 6     6     Poa pratensis 7     Elymus repens (+) 7 Lolium perenne (+) 8 Poa annua Herbs: ×    Ranunculus repens (+) 5 Bellis perennis (+) 6 Plantago major (+) 7 Taraxacum officinale 8    Cirsium vulgare

Grasses: ×    Anthoxanthum doratum ×    Festuca rubra Herbs: 5     Achillea millefolium 5     Leontodon autumnalis

Indifferent species and N indicators

716 8  Agricultural Grassland on Mesic to Wet Soils

8.8  Human Influence

717

Table 8.19  Changes in Arrhenatheretum meadows over 30 years under different management treatments in the Ojców National Park (Poland)

C C C

C

C

A

A A

M × × 4 3 4 6 × 6 5 5 × 5 × 5 4 5 5 6 4 6 4 × × 5 5 6 6 6 3 7 5 – 6 5 4

± Declining species: N Grasses and graminoids 2 • Briza media × • Bromus hordeaceus 3 • Luzula campestris 4 Avenochloa pubescens 6 • Carex muricata 5 • Carex hirta 4 • Agrostis capillaris × • Festuca rubra 4 • Cynosurus cristatus 7 Phleum pratense × • Anthoxanthum odoratum 7 • Lolium perenne 5 Trisetum flavescens 6 • Poa pratensis Legumes 4 • Trifolium dubium 6 • Trifolium repens × • Trifolium pratense × Vicia cracca × • Medicago lupulina 6 Lathyrus pratensis Herbs 3 • Plantago media × • Plantago lanceolata × • Prunella vulgaris 6 • Leontodon hispidus 6 • Bellis perennis 7 • Glechoma hederacea × • Lysimachia nummularia 7 • Primula elatior 2 • Pimpinella saxifraga × • Alchemilla gracilis 5 Cruciata glabra − • Leontodon hispidus glabra 5 • Crepis biennis 5 • Cerastium holosteoides 4 Knautia arvensis

I a 3 3 – 3 2 2 4 5 2 5 4 2 5 4

b 1 1 – 3 1 – 3 3 1 4 3 1 3 3

II a 3 3 3 3 4 5 4 5 4 5 4 2 4 2

b 1 – – 2 – – 1 2 – 1 3 – 4 3

III a 3 4 4 4 2 3 5 5 2 2 5 2 4 5

b − − − – − − − − − 1 − − – 1

5 5 5 5 3 5

– 3 2 2 1 3

4 5 4 3 2 5

– – – 2 – 4

5 5 3 4 4 5

− − − − − 1

5 5 5 5 5 4 3 + + 4 3 3 5 5 5

1 3 – 2 3 2 1 + + – + – 3 3 3

5 5 5 4 4 2 4 4 2 3 3 4 4 5 3

1 1 1 1 – 1 2 2 – 2 2 – – 3 2

5 5 5 5 5 – 4 4 4 2 4 4 5 5 4

− − − − − − 1 − − – – – 1 1 −

(continued)

8  Agricultural Grassland on Mesic to Wet Soils

718 Table 8.19 (continued)

A A

C

A

A

A A

M 4 4 5 4 × 4 5 6 × 8 5 7 5

N ? 5 5 3 3 2 5 6 4 9 5 6 6

M 7 6 5 × 6 7 × 8

N 3 6 6 7 7 7 7 7

5 7 5 5 5 6 T 8 7 5 5 7

8 7 × 7 4 × 6 3 × 6 3 7

± Declining species: Grasses and graminoids Galium mollugo • Achillea millefolium • Campanula patula • Leucanthemum vulgare • Equisetum arvense Centaurea jacea • Leontodon autumnalis • Ajuga reptans • Euphrasia rostkowiana • Geum rivale • Veronica serpyllifolia Rumex crispus Tragopogon orientalis ± Stable and new species: Grasses Deschampsia cespitosa Festuca pratensis Dactylis glomerata Arrhenatherum elatius Alopecurus pratensis • Poa trivialis Elymus repens Phalaris arundinacea Herbs Heracleum sphondylium • Ranunculus repens • Veronica chamaedrys Geranium pratense • Alchemilla monticola Ranunculus acris Rumex acetosa Equisetum palustre Lychnis flos-cuculi Pimpinella major Stellaria graminea Geranium palustre

I a 4 5 4 5 2 4 + + – 2 3 3 3 I a 5 5 5 4 5 5 – + 5 5 5 5 1 5 5 2 3 1 1 +

b 3 5 5 3 4 3 2 –

II a 3 4 2 5 4 – 3 – 2 – – + – II a 5 4 4 4 3 2 + +

5 5 5 5 1 5 5 – 3 2 3 +

4 5 5 2 1 4 5 2 4 – 2 3

b 3 3 2 3 1 2 1 + – 1 – – –

b 5 3 5 4 3 5 + –

III a 4 5 3 4 4 – 2 4 2 2 – – – III a 5 5 5 4 3 2 1 –

b 1 1 2 3 2 4 3 2

1 3 4 1 – 4 5 3 5 – 3 4

5 2 5 4 2 3 5 3 3 2 + 2

1 3 1 3 − − 1 1 1 − + 3

b 2 2 1 4 2 – – – – – – – –

b 1 − − – 1 − − − − 1 − − −

(continued)

8.8  Human Influence

719

Table 8.19 (continued)

7 6 5 7 6 8 7 × 6

8 9 5 5 8 7 7 8 9

± Stable and new species: Anthriscus sylvestris Chaerophyllum aromatic. Rumex obtusifolius Geranium phaeum Cirsium oleraceum Aegopodium podagraria Mentha longifolia Stellaria nemorum Galium aparine Urtica dioica

I – – 2 – 1 – – – – –

3 2 3 2 2 – + – 2 3

II – 1 1 + 2 – – – – –

+ 2 2 – 5 + – + 4 5

III + – – 2 1 – 1 – 1 –

− 3 3 4 5 3 4 3 5 5

Compiled from tables in Kornas and Dubiel (1991) The numbers 1 to 5 in columns I to III refer to constancy classesa a = first survey in 1958/9, b = repeat in 1988. In 1958/9, all stands were well-developed Arrhenatheretum communities with many Arrhenatherion species (A), but also some Cynosurion species (C; the meadows were cut twice, fertilised with manure and grazed after the second cut). Later they were only partially used by farmers: I = strongly fertilised and mown several times a year (7 relevés each) II = mown and mulched once a year by the park authorities, no longer fertilised (5 relevés each) III = abandoned (1958/9 5 relevés, 1988 10 relevés) M = indicator value for moisture, N = for nitrogen, from Ellenberg et al. (1992) • = low-growing, at least temporarily The largest loss of species appears to have been caused by abandonment. Surprisingly, the lack of fertilisation and/or use increased the growth of some particularly nitrogen-demanding forbs, namely species of swamp and floodplain forests such as Urtica dioica and Aegopodium. They profited from being able to grow (almost) without disturbance, and could build up their own internal nutrient cycle. Species that had previously profited from dry conditions such as Avenochloa pubescens and Pimpinella saxifraga declined or completely disappeared a 1 = in 1–20 % of the relevés, 2 = 21–40 %, 3 = 41–60 %, 4 = 61–80 %, 5 = 81–100 %, + only in one relevé, − absent

The rapid effects on the species composition of a typical Lolio-Cynosuretum pasture of high levels of fertilisation and high stocking densities were shown by Dörrie (1958) in repeated surveys of permanent plots. By the second year, many species had disappeared including even persistent weeds such as Cirsium arvense. Only a few new species appeared, e.g. Stellaria media as an indicator of high N supply and Polygonum aviculare as a trampling-resistant plant. Lolium perenne, Phleum pratense and other grasses spread at the expense of the legumes to such an extent that they were almost the only species left by 1954. A single application of large amounts of nutrients can therefore have very long-lasting effects on the character of a meadow or pasture. In addition to the sharp increase in fertilisation, changes in the species composition of agricultural grassland were also caused by atmospheric nitrogen deposition (Stevens et al. 2004). Linusson et al. (1998) compared relevés on acidic, unfertilised hay meadows in southern Sweden from 1963 to 1966 with ones from 1990, under largely constant management. The average species density over all the plots changed little, but the total coverage of the stand increased. 39 of the species present in 1963 to 1966 had disappeared by 1990, and another 31 had joined the communities. The

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8  Agricultural Grassland on Mesic to Wet Soils

species that had been lost were mainly naturally rare grassland species, whilst the new species were widespread generalists. The authors considered these changes to be mainly caused by the doubling of the atmospheric N inputs over the 25-year period.

8.8.2  The Effects of Drainage and Irrigation Drainage of Wet Meadows  In groundwater-influenced areas such as valley bottoms and low-lying coastal plains, large-scale drainage preceded the intensification of grassland use. A large-scale re-sampling study in more than 380 semi-permanent plots in five floodplain grassland regions in the lowlands of northern Germany showed that the mean species number per relevé had decreased by 30 % (from a median of 27 species to 19) between the 1950s and the 2000s, and 23 of 30 characteristic species of wet meadows had significantly declined (Krause et al. 2014, 2015; Wesche et al. 2012). The list of ‘losers’ includes many once widespread grassland species such as Anthoxanthum odoratum, Lychnis flos-cuculi, Cardamine pratensis and Bellis perennis (Fig. 8.55). Only 7 nitrogen-demanding, productive species had increased, including Phleum pratense, Lolium perenne, Rumex obtusifolius and Urtica dioica. While the size of the regional species pool in the five regions remained unchanged, 37 % of the species formerly present had disappeared, mostly former characteristic species of wet grassland. They had been replaced by less specialised taxa with more widespread distribution in the cultural landscapes of Central Europe. Comparison with historical vegetation maps of the five areas showed that the area of species-rich, extensively used mesic and wet grasslands has decreased since around the 1950s by 84 and 85 %, respectively (Krause et al. 2015), mostly due to drainage, management intensification or conversion to arable. Given that plant diversity had also decreased in the remaining extensively used grassland patches, it can be concluded that the population sizes of many characteristic grassland species such as Silene flos-cuculi, Caltha palustris, Leontodon antumnalis or Agrostis capillaris must have decreased by 95 % or more during the last 50–60 years in the lowlands of northern Germany (Wesche et al. 2009; Leuschner et al. 2013a; Krause et al. 2011). Species-rich extensively used floodplain meadows are additionally threatened by increasing isolation. Winter et  al. (2008) studied the remnant populations of five endangered perennial herbs (Lathyrus palustris, Pseudolysimachion longifolium and others) in the floodplains of river Weser and found that seed numbers and germination were negatively affected by small population sizes and strong isolation. In the Holtumer Moor, a typical low-lying area of northwestern Germany with former fens and raised bogs east of Bremen, the grassland communities have been surveyed and mapped at intervals of 25 years (1963/1964 by Dierschke, 1988 by B. Wittig and 2006 by B.Wittig et al., published in Wittig et al. 2007). Based on these 40 permanent plots in meadows and pastures, the changes here can be quantified. Instead of species-rich wet meadows and damp continuously grazed pastures, these areas are now covered with homogenous rotational pastures and frequently cut meadows, and in places also arable fields (Dierschke and Wittig 1991). The changes in the species composition of the meadows were particularly dramatic (see

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721

Fig. 8.55  Change in the frequency of characteristic meadow species in each 385 relevés conducted either in the 1950s/1960s or 2009 in five floodplain meadow regions of northern Germany (Modified after Krause et al. 2014)

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8  Agricultural Grassland on Mesic to Wet Soils

Table 8.20  Changes in constancy and coverage of species from 1963 to 1988 in 40 damp meadow areas (average percent) in the Holtumer Moor east of Bremen, modified from Dierschke and Wittig (1991)a, compared to indicator values from Ellenberg et al. (1992) Lost species: B Calliergonella cuspidata   Lathyrus pratensis   Eleocharis palustris   Luzula campestris   L. multiflora   Carex rostrata   Eriophorum angustifolium   Cirsium vulgare   Briza media   Glyceria maxima   Rhinanthus minor Species with strong decline:   Carex nigra   Galium palustre   Juncus filiformis   Lotus uliginosus   Galium uliginosum   Juncus conglomeratus   Carex leporina   Cynosurus cristatus   Caltha palustris   Carex acuta   Senecio aquaticus   Angelica sylvestris   Filipendula ulmaria   Juncus effusus   Bromus racemosus   Ajuga reptans   Trifolium repens   Carex canescens   Myosotis palustris Species with slight decline:   Ranunculus acris   Vicia cracca   Crepis paludosa   Anthoxanthum odorat.   Plantago lanceolata   Lychnis flos-cuculi   Trifolium pratense

Constancy 1963 1988 43 – 35 – 23 – 20 – 20 – 18 – 15 – 15 – 13 – 13 – 13 –

Coverage 1963 1988 3.0 – 0.9 – 0.3 – 0.2 – 0.1 – 0.7 – 0.4 – 0.1 – 0.1 – 0.8 – 0.8 –

M

N

7 6 10 4 5~ 10 9= 5 × 10~ 4

– 6 ? 3 3 3 2 8 2 9 3

70 38 58 90 40 25 23 55 85 63 73 88 63 40 33 38 78 18 98

5 3 5 10 5 3 3 8 13 10 13 18 13 10 8 10 18 5 30

3.4 1.2 2.9 2.8 0.6 0.8 0.2 2.3 4.5 10.6 2.1 2.2 2.3 2.3 0.4 0.5 3.0 0.6 1.5

no management. However, almost all mowing or mulching experiments have shown that the nutrient availability only gradually declines, and characteristic species of low-nutrient conditions (e.g. orchids) only colonise many years later, if at all. This is particularly difficult on clay soils with high cation adsorption, as shown in studies by e.g. Pfadenhauer et al. (1987) in the Allgäu, Schiefer (1981) in southwestern Germany, Rosenthal (2001) in northwestern Germany and Bakker and ter Heerdt (2005) in the Netherlands. Fix and Poschlod (1993) were unable to reduce the nutrient content of managed grassland, at least in the short term. Bischof (1992), in contrast, found a clear increase in species richness over a 15-year period on lowland meadows (some of which had dry soils) in Switzerland without fertilisation and with one to two cuts per year. On areas with formerly low fertiliser application, this was from 58 to 75, on areas with formerly high fertilisation from 43 to 62, and on intensively grazed sheep pastures from 37 to 58 species (see Fig.  8.57). Poptcheva et  al. (2009) observed the re-establishment of species-rich wet meadow communities after 20 years of mowing twice annually without fertiliser application. The conservation of at least a few, large unfertilised litter meadow complexes with late mowing is, alongside the protection of wet meadows, the most important task for the conservation of agricultural grassland (Klötzli 1979; Briemle 1986, 1998a). Whilst large areas of litter meadows were available for study throughout Central Europe in the early twentieth century (e.g. Koch 1926), soon only a few fragmentary examples of this extensive management form will remain in nature reserves. Long-running experiments on the Swiss Plateau have shown that the characteristic species composition of Molinia grasslands is best maintained by mechanical mowing at least every 2 years in late autumn with removal of the cuttings. Regular mowing and removal of the material for pure conservation benefit is, however, only feasible if this is financed through e.g. subsidy schemes or regional marketing of high-value products (Hampicke 1996, 2013). If mowing is carried out too early, then it disturbs the phenological development (Weber and Pfadenhauer 1987), the germination conditions, flowering and seed-set, in particular of Molinia (see Sect. 8.3.3.2). The high grass:forb ratio typical for Molinia meadows is only maintained if the nitrate and phosphate levels are very low (Pegtel 1983). The species composition of litter meadows can by changed even by being next to a fertilised meadow (Boller-Elmer 1977). Particularly P limitation is very important for high species richness in these stands (Olde Venterink 2000). Jannsens et al. (1998) found higher species densities (>20 species per 100 m2) in Western European grassland communities only when the exchangeable P content in

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729

Fig. 8.57  Changes in the species richness in formerly fertilised meadows cut once to five times a year in the first 10 years after fertilisation ceased (Aulendorf extensivation experiment) (From Dierschke and Briemle 2002)

the soil (tris-acetate-EDTA extraction) was below 1.6 μmol g−1. P reduction and thereby species-rich grassland can most effectively be achieved by topsoil removal. Based on numerous long-term studies, Schreiber (1993, 2001) concluded that mulching twice a year can be an acceptable alternative to the more labour-intensive mowing with hay removal in various meadows in southwestern Germany, although it does cause some decline in differential and character species. The mulching of dry or mesic nutrient-rich meadows can, similarly to hay removal, reduce the nutrient levels in the stand. In contrast, in damp and wet meadows, the mulch decomposes slowly and suppresses low-growing plants (Dierschke and Peppler-Lisbach 1997; Rosenthal 2001). The loss of nutrients from mulched but unfertilised wet meadows is thus very slow. Broll and Schreiber (1993) even observed an initial increase in the P and K concentrations in the topsoil of wet spring communities, montane Arrhenatheretum meadows and formerly grazed calcareous grasslands in the first 6 to 9 years of conservation management of these damp to wet grasslands (ranging from mulching twice a year to no interventions). It was only in the following 10 years that the P and K concentrations returned to their levels before conservation management began. Thus, meadows can become nutrient-poorer, at least in P and K, even without hay harvesting, but this may take long time, in particular on wet soil. The restoration of former species-rich wet meadows is usually just as problematic as that of drained mires (see Sect. 3.7.4). Flooding by blocking drainage channels initially promotes mainly species of reedbed and tall- and small-sedge

730

8  Agricultural Grassland on Mesic to Wet Soils

communities, but not the Calthion and Molinion species (Hellberg 1995). The restoration or creation of communities in these alliances depends on the water table remaining high throughout the growing season. Increasing the soil moisture can also help to reduce the nutrient status of damp meadows if nitrogen, rather than phosphorus, is the most important limiting factor (Rosenthal 2001). Schreiber (1995, 2001) states that grazing of mesic meadows is relatively ineffective as an early restoration measure, as the animals selectively target only certain species and cause trampling damage. However, after a dense sward has been created by several years of mowing, low-intensity grazing can be effective at maintaining certain communities. For example, sheep grazing helped to restore species-rich Czech mountain meadows, although there were problems with the domination of Deschampsia cespitosa in some areas (Krahulec et al. 2001). Sheep and horse grazing combined with regular mowing helped to maintain species-rich Salvia-­ Arrhenatheretum communities in southern Germany (Wagner and Luick 2005). In general, the shorter the grass is grazed, the more similar the effects of grazing to mowing (Spatz 1994). Longer grazing periods promote pasture species and finally lead to the formation of selectively grazed pastures, as were once widespread in Central Europe. Controlled burning is generally not a good tool to conserve Central European grassland communities, because it is detrimental to various plant species and may harm the invertebrate fauna, but can be useful to control woody species (Schreiber 1997, 2001). In any case, the restoration of intensive grassland is much more difficult than the conservation of recently abandoned traditional meadows and pastures, which still contain a seed bank (Poschlod and Binder 1991). In large areas of intensive grassland, even small remaining patches of species-rich mesic or wet grassland deserve protection because they can function as diaspore sources for re-colonisation (Krause et al. 2015). Gugerli (1993) found that the soil of nutrient-rich mesic meadows in Switzerland recently taken out of intensive use contained no seeds of e.g. Helictotrichon pubescens, Bromus erectus, Crepis biennis, Centaurea scabiosa, Scabiosa columbaria or Tragopogon orientalis.There were also only tiny numbers of seeds of Arrhenatherum elatius, Trisetum flavescens, Festuca pratensis, Centaurea jacea, Salvia pratensis and other species present in the vegetation. Generalists such as Poa trivialis, Cerastium holosteoides, Cardamine hirsuta and Taraxacum officinale were, in contrast, common in the seed bank, and arable weeds such as Chenopodium polyspermum and Capsella bursa-pastoris as well as other non-meadow species were also present in large numbers. Gugerli (1993) therefore recommends removing sods from intact grassland communities and transplanting them here and there in the restored area, so that the major species can spread. The introduction of seeds through hay spreading (Poschlod and Schumacher 1998; Hölzel and Otte 2003) is also an effective measure. Rosenthal (2001) states that abandoned wet meadows can be restored more easily and in shorter time frames than intensive grasslands or arable fields, because they contain a reasonably complete seed bank even after 30 years of abandonment (but cf. Dierschke and Briemle 2002 and Sect. 8.5.2). However, the seeds of target species such as Silaum silaus

8.8  Human Influence

731

and Serratula tinctoria rarely disperse more than 2 m from the mother plant, making their reintroduction difficult (Bischoff 2002). Organic farming practices can increase the number of species in meadows and pastures. In a comparison of 275 plots in organic and conventional grassland, Mahn (1993) found that organic grassland contained on average 5 species per plot more. Schwabe and Kratochwil (1994/95) found a 30 % higher species richness. However, very few Red List species were found in Mahn’s plots (cf. Bakker and ter Heerdt 2005).

Chapter 9

Communities on Heavy Metal-Rich Soils

The scattered fragments of metallophyte vegetation, i.e. plants growing on soils with high concentrations of zinc, copper, lead, manganese, iron, nickel, chrome or cobalt, form one of the ecologically most intriguing groups of plant communities in the world (Ernst 1974). In the dense vegetation of Central Europe, these islands of low-growing and sparse vegetation are usually found in the vicinity of calcareous and sandy nutrient-poor grasslands. Rock with high concentrations of heavy metals, defined here as elements with a density greater than 5 g cm−3, are found at the surface in several Central European mountain areas (e.g. in southwestern Poland and western Czech Republic, the Ore Mountains, the Harz and the Hunsrück Mountains and the Alps) and form natural communities of heavy metal-tolerant vegetation. In contrast, such outcrops of heavy metal-rich rock are rare in the lowlands, as younger layers of sediment cover the older bedrock. Metalliferous veins have been mined since the early Bronze Age (from around 1800 BC), destroying many natural heavy metal habitats. However, the mining activities also created numerous spoil heaps, considerably increasing the area of heavy metal-rich substrate e.g. in the Eifel and the Siegerland in western Germany. Miners initially deposited the excess rock on small, scattered heaps, but later created large, contiguous dumps. The early, primitive smelting techniques left a large amount of metal in the slag, so that even their slag heaps were rich in heavy metals. In the valleys of the Harz Mountains, such slag heaps can be found long distances from the next source of ore, because of the difficulty in hauling the increasingly limited firewood e.g. to Goslar from the central and southern Harz valleys. When returning to the forest, the pack animals were therefore loaded with ore to smelt it there. Secondary habitats also include the floodplains where heavy metal-­ rich tailings left over from processing the ore were washed down and deposited, and e.g. in the Harz reach well into the foothills (e.g. Dierschke and Becker 2008). A special class of naturally heavy metal-rich substrates that can be found scattered across the globe are the serpentine soils, which are not only rich in the heavy metals nickel, chrome and cobalt, but also have very high magnesium concentrations (also known as ultramafic soils). They are therefore quite exceptional in ­several © Springer International Publishing Switzerland 2017 C. Leuschner, H. Ellenberg, Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats, DOI 10.1007/978-3-319-43048-5_9

733

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9  Communities on Heavy Metal-Rich Soils

aspects of their soil chemistry. In Central Europe, serpentine soils only occur in small areas, namely in the Harz (near Clausthal), in northeastern Bavaria (near Grötschenreuth and Wurlitz), in several areas in northern Styria (e.g. near Kraubath in the Mur valley), in Burgenland and in Lower Austria, in western Czech Republic (Mariánské Lázně) and south central Czech Republic (e.g. near Mohelno) and in southern Poland (Szklana Góra near Zabkowic, Eggler 1955; Sarosiek 1964; Sulej et al. 1970; Sasse 1979a, b; Müller-Stoll and Toman 1984; Brooks 1987). In recent decades, air pollution has provided heavy metal-tolerant species with new opportunities to spread, for example close to smelting works and on roadsides with high emissions of heavy metals from industry and automobile exhaust fumes. Waste water, sewage sludge and rubbish can also contain heavy metals (see Sect. 3.7.5 in Vol. I).

9.1  The Origins and Development of the Heavy Metal Flora Central Europe is home to only a few plant taxa adapted to heavy metal-rich soils that can be considered separate species. The perhaps best known of these metallophyte species is Viola calaminaria (zinc violet), which is morphologically and cytotaxonomically clearly distinguishable from the related Viola lutea (Heimans 1961, 1966, but cf. Nauenburg 1986). There are also specialised serpentine species that occur on the serpentine soils rich in magnesium, nickel and chrome in the Alps, the Czech Republic and Bavaria, for example Asplenium cuneifolium, Sempervivum pittonii, Myosotis stenophylla, Festuca eggleri, Armeria serpentini and Cerastium alsinifolium, which are almost entirely restricted to these habitats (Zlatnik 1928; Gauckler 1954; Mucina and Kolbek 1993a,b). However, there are numerous other species, particularly of nutrient-poor grasslands and scree communities, as well as of fertile meadows, that have ecotypes, varieties or subspecies adapted to one or more heavy metals. Many of these will be discussed in Sect. 9.3 (see also Brown 2001). Silene vulgaris (bladder campion) forms a particularly large number of ecotypes in Central Europe, many of which are resistant to Cd, Cu, Ni, Pb and Zn (Ernst 2001). These heavy metal forms have developed from the ‘normal’ ecotype independently in many areas of Central Europe (Baumbach 2005). The same is true for the heavy metal ecotypes of Armeria maritima, and many of its regional forms have been wrongly considered subspecies or even species in the past. Many metallophyte taxa are phenotypically very variable over small spatial scales, and even a single slope can contain several different ecotypes of a species (Ernst 1974). Tolerance of heavy metals is mainly genetically determined, but can be adaptively modified, as genetic analyses and crossing experiments have shown (Bröker 1963; Ernst 2001). The required characteristics appear to be able to develop in grasses within several generations, so are likely to be linked to only a few genes. Ernst (1976) recorded the selection of heavy metal tolerant ecotypes within only 3 years in populations of Agrostis capillaris under the influence of zinc and

9.2  Heavy Metal Soils

735

c­ admium deposition. This process promotes the development of metallophyte neoendemics at very small scales, because the newly formed taxa are spatially and genetically isolated from the parent populations due to their extreme habitat. This is especially the case if they have a strong tendency to self pollination (cf. Duvigneaud and Denaeyer-de Smet 1963). Ernst (2001) found that dicot herbs, particularly taxa of the Brassicaceae and Caryophyllaceae families that do not have mycorrhizae, need much longer to develop heavy metal resistance than grasses. Presumably the now strongly isolated ranges of heavy metal ecotypes such as the Caryophyllaceae Minuartia verna subsp. hercynica are therefore more likely to be a case of palaeoendemism. Their genome varies relatively little but is nonetheless geographically highly differentiated (Baumbach 2005). This may also be the case for the genus Thlaspi with the mostly Alpine species and subspecies T. alpestre, T. alpinum subsp. sylvium, T. calaminare, T. goesingense and T. rotundifolium, which are highly nickel-resistant (Reeves and Brooks 1983). The predecessor of these taxa probably used to be more widespread, and sought refuge in the open vegetation of heavy metal soils during the postglacial forest expansion (Ernst 1974). As Ernst showed in the case of Wales, heavy metal plants have been present in northwestern Europe since at least the eleventh century AD (and probably even longer). Today, more competitive grasses, forbs and woody plants prevent the spread of these specialists into less heavy metal-­rich habitats. Particularly the decline in mining now means that some of the Central European heavy metal communities are threatened with extinction. For a discussion of metallophyte conservation see e.g. Pardey (1999).

9.2  Heavy Metal Soils Heavy metal soils occur on both base-rich and acidic bedrock and may contain a mixture of 0.1–10 % zinc, 0.1–3.2 % lead and 0.05–10 % copper or up to 10 % nickel, cobalt or cadmium (Ernst 1974). The wide variety of different bedrocks means that the qualities and quantities of heavy metals in both natural and anthropogenic metalliferous soils can be very variable. However, ecologically more important than the total soil content is the exchangeably-bound fraction, and particularly the concentrations in the soil liquid phase. As Table 9.1 shows, heavy metal soils can have 10–10,000 times the concentrations of Zn, Pb and Cu, and sometimes also of Ni, Cr and Cd, in their liquid phase than heavy metal-poor ‘normal’ soils (e.g. the Rendzina- and Ranker-Cambisols in Table 9.1). The amounts of heavy metal bound to the cation exchangers are also several times higher, and can occupy 50–80 % of the exchange sites (Ernst 1982). Brown (2001) states that the absolute heavy metal concentrations are physiologically less relevant than the relative proportions of the ions. Typical heavy metal grasslands in the Eifel are characterised by Pb/Ca ratios of > 1 and Zn/Ca ratios > 0.2 in the soil. High calcium and phosphorus concentrations lead to the precipitation of heavy metals and reduce their availability and thus toxicity (Proctor and Woodell 1975;

9  Communities on Heavy Metal-Rich Soils

736

Table 9.1 Heavy metal concentrations (total, exchangeable with ammonium‑acetate and extractable in water) in soils from four areas contaminated with heavy metals in comparison to two ‘normal’ soils with low heavy metal contents (Rendzina and Ranker Cambisols) (in μg g−1) Locality or soil type Blankenrodeb Eislebenc Erbendorfd ‘normal’ non-heavy metal soil Blankenrode Eisleben Erbendorf Kraubathe Calcaric Leptosol-­ Cambisol Dystric Leptosol-­ Cambisol Blankenrode Eisleben Erbendorf Kraubath Calcaric Leptosol-­ Cambisol Dystric Leptosol-­ Cambisol

Bedrock typea Fe Mn Total element content Zn/Pb-rich 27,000 730 rock Zn/Cu/ 37,500 2000 Pb-rich rock serpentine 28,000 930 2000– 40– 50,000 1000

Zn

Pb

Cu

Co

Ni

Cr

Cd

90,200 2800 45

14

63

4

58

100

10

49

16,500 6000 16,000 110 120 0 25 10–80 2–80 2–40

100 700 330 0 1–40 3–50 5–100 0.1– 0.6

Salt-exchangeable element content (1n ammonium acetate) Zn/Pb-rich 0 100 1340 366 4 0 0 rock Zn/Cu/ 0 23 740 378 435 0 1 Pb-rich rock serpentine 10 47 3 100 1 3 8 serpentine 10 204 8 160 6 13 16 limestone 0 300 10 12 2 0 0 (Ca-rich) silicate rock 15

10

5

10

Element concentration in water extract Zn/Pb-rich – 4 180 0.2 rock Zn/Cu/ – 1 21 0 Pb-rich rock serpentine – 0 1 0 serpentine – 22 0.7 0 limestone – 0 0.6 0 (Ca-rich) silicate rock –

–

0.01

0

0

26

0

11

10 0 0

0 0 1

2

0

0

0

0

1.0

–

0

0

0.20

19.6

–

0

–

0.02

0.3 0.01 0.004

– – –

6.6 2 10.4 0 0 0

0 0 0

0.008

–

0

0

0

Soil chemical characteristics Sauerland, western Germany c Eastern Harz, eastern Germany d Oberpfalz, southern Germany e Styria, Austria From Ernst (1974) with additional data from Scheffer and Schachtschabel (2002, 2010) a

b

9.2  Heavy Metal Soils

737

Table 9.2  Soil chemical properties and nutrient availablity in soils from four areas contaminated with heavy metals compared to two ‘normal’ soils with low heavy metal contents (Rendzina and Ranker Cambisols)

Locality/soil type Blankenrode Eisleben Erbendorf Kraubath Calcaric Leptosol-­ Cambisol Dystric Leptosol-­ Cambisol

Bedrock type Zn/Pb-rich rock Zn/Cu/Pb-rich rock serpentine serpentine limestone (Ca-rich)

pH (H2O) N (%) 6.8 0.12

P2O5 (mg kg−1) 167

7.8

0.09

–

247

2.8

12.5

6.0 6.6 6.8

– – 0.27

14 15 185

67 65 394

6.2 0.9 5.8

39.5 197.5 49.3

silicate rock

5.4

0.12

115

22

2.2

3.0

Ca (μmolc g−1) 205

K (μmolc g−1) 3.1

Mg (μmolc g−1) 3.5

From Ernst (1974), with additional data from Scheffer and Schachtschabel (2002, 2010). Ca, K, Mg: exchangeable concentrations

Garland and Wilkins 1981). Roots and mycorrhizae influence the heavy metal availability in the soil by release of H+ and organic acids. The latter lead to the formation of water-soluble organometallic complexes (particularly of Cu, Ni and Pb). Depending on the components of the complex, these can either increase or decrease the mobility of the heavy metals in the soil (Kabata-Pendias and Pendias 1992). Organically-bound heavy metals, e.g. in association to humic acids, are taken up in smaller quantities than free heavy metal ions. Humus-poor metalliferous Leptosols are therefore particularly toxic and difficult for plants to colonise. Table 9.1 clearly shows the difference between the zinc, lead and copper-rich soils of Blankenrode (Sauerland) and Eisleben (Harz foothills) on the one hand, and the nickel, chrome and cobalt-rich serpentine soils of Erbendorf (northwestern Bavaria) and Kraubath (Styria) on the other. Krause (1958a) states that substrates formed from serpentine rock (Mg3Si2O5(OH)4) consist on average of 37.4 % MgO, 10.9 % iron oxide, 0.3 % NiO, 0.5 % Cr2O3 and 0.02 % CoO. Serpentine soils are therefore not only rich in many heavy metals, but also have very high Mg and Fe concentrations whilst being poor in Ca. For comparison, sandstone consists of 80 % SiO2 and only around 3.6 % Mg and Fe oxides. The serpentine soils with high Mg, Ni, Cr and Cd concentrations are relatively poor not only in calcium, but also in phosphorus, which adsorbs to the serpentine (see Table 9.1 and Table 9.2; Sasse 1979a, b; Brooks 1987). The very low Ca/Mg ratio (< 1) indicates an excess of Mg, which has a large physiological impact. Usually, the cation exchangers in these soils are completely occupied by magnesium ions. Zinc-, lead- and copper-rich metalliferous soils differ from ‘normal’ soils only in their heavy metal content and much lower microbial activity, but not in any other

738

9  Communities on Heavy Metal-Rich Soils

soil chemical properties (Ernst 1974). However, Ernst also found 30–50 % lower nitrogen mineralisation rates in the heavy metal soils of Blankenrode near Paderborn and Marsberg in Sauerland (central-western Germany) than in a Rendzina-Cambisol. Low mineralisation rates lead to the accumulation of mor humus in many heavy metal soils. Their relatively uncompacted structure also means that many soils on metalliferous spoil heaps have a low water storage capacity, a characteristic they share with dry nutrient-poor grasslands.

9.3  Vegetation 9.3.1  Vascular Plant Communities Similarly to the plant communities on saline soils or in other extreme habitats, the vegetation of true heavy metal habitats is species-poor. More diverse transitional stages occur to neighbouring grassland communities, e.g. to the Xerobrometum or Mesobrometum. This is why the heavy metal vegetation was at one time assigned to the order Brometalia erecti or at least the class of the Festuco-Brometea (see Sect. 7.3.2). However, Braun-Blanquet and Tüxen (1943) place it in its own class for Europe and western Siberia, namely the Violetea calaminariae. As emphasised by Ernst (1974), Viola calaminaria, after which it is named, tolerates heavy metals, but is not a good class character species as it does not occur in many regions. The European metallophyte vegetation is instead best characterised by Minuartia verna subsp. hercynica (see Fig. 9.3) and Armeria maritima subsp. halleri as well as ecotypes of Silene vulgaris var. humilis. Some ecotypes of Rumex acetosa and R. acetosella can also be used, but these are morphologically difficult to distinguish from the normal form. All Western and Eastern European phanerogam communities on heavy metal soils furthermore belong to the order of the Violetalia calaminariae, which is best characterised by the lack of Southern European species. At most, the ecotypes of Festuca ovina and Agrostis capillaris can be used as order character species. Ernst (1974) distinguishes three alliances: 1. The Galio anisophyllo-Minuartion vernae of alpine habitats with ecotypes of Galium anisophyllum, Dianthus sylvestris and Poa alpina; 2. The Thlaspion calaminariae of Western Europe (to western Central Europe) with Thlaspi calaminare (= T. caerulescens subsp. calaminare); 3. The Armerion halleri scattered across Central Europe with ecotypes of Armeria maritima subsp. halleri and without Thlaspi calaminare (although its range stretches to Poland). Not all phytosociologists accept this classification. Mucina and Kolbek (1993a), for example, assign the serpentine communities characterised by high magnesium and heavy metal concentrations (the alliance of the Avenulo-Festucion pallen-

9.3 Vegetation

739

tis) to the class of the Festuco-Brometea, avoiding the use of the Violetea calaminariae. The other heavy metal communities in the eastern Alps not on serpentine are also assigned by Punz and Mucina (1997) to the class of the Thlaspietea rotundifolii rather than to the Violetea calaminariae. Schubert (1954) and Becker et al. (2007) found that some grassland-like heavy metal communities of central Germany are best related to the Festuco-Brometea. Although this may make sense from a synsystematic point of view, it is ecologically unsatisfactory, because the habitats and physiology of the species in these communities are clearly different from those on non-metalliferous soils (cf. Dierschke and Becker 2008). The best-studied associations in this group include the Violetum calaminariae (in the Thlaspion calaminariae alliance) and the Armerietum halleri (also known as the Silene vulgaris-Armeria maritima community; in the Armerion halleri) of the Harz Mountains, its northern and eastern foothills, and the Eifel (western Germany; see e.g. Schubert 1954; Brown 2001; Becker et al. 2007; Dierschke and Becker 2008). Brown (2001) found that the main selective pressure in the Violetum of the Eifel was exerted by zinc, whilst in the Armerietum it was lead or copper (cf. Becker and Brändel 2007). Subassociations can be distinguished in both communities based on differences in soil depth and moisture. In the case of soil moisture, e.g. Cardaminopsis halleri (which probably forms a different ecotype on non-­ metalliferous soils, e.g. in montane meadows) can only be found on soils with a good water supply, whilst numerous lichen species are found mainly on shallow soils, and Plantago media, Galium album and Achillea millefolium are characteristic of the transition to meadow communities. The character species that have been mentioned so far make up a large proportion of the vascular flora of Central European metalliferous habitats. Some of them may not occur in certain areas, e.g. because no diaspores have yet reached it. For example, an expansion in the range of Thlaspi calaminare has recently been observed, showing that this species has not yet colonised all available habitats. The yellow-­ flowering Viola calaminaria is replaced in North Rhine-Westphalia by the blue Viola guestphalica, both of which are endemic to northwestern Central Europe but over different ranges. Isolation and random dispersal events thus played, and continue to play, an important role for the species composition of heavy metal communities.

9.3.2  Lichen Vegetation Naturally acidic metalliferous soils and slag heaps support a highly characteristic lichen vegetation. The lichen communities of iron-rich substrates, which belong to the alliance of the Acarosporion sinopicae (Wirth 1972), clearly differ from those of copper-rich substrates (the Lecideion inopis; Purvis and Halls 1996). The often species-rich stands of the Acarosporion sinopicae are typically characterised by, apart from Acarospora sinopica itself, Lecanora gisleriana, Rhizocarpon furfurosum and R. oederi. Communities on rock overhangs sheltered from the rain are

740

9  Communities on Heavy Metal-Rich Soils

classified as the Lecanoretum epanorae (with Lecanora epanora and L. handelii), whilst the vegetation of exposed habitats is placed in the Acarosporetum sinopicae. Only one community, the Lecideetum inopis, has been described from copper-rich habitats (Purvis 1996; Purvis and Halls 1996). In Central Europe, it is found mainly on copper-rich spoil heaps in the southern foothills of the Harz with species such as Lecidea inopis and Psilolechia leprosa (Huneck 2006). Lichens grow on both bare rock and soil in heavy metal habitats, often with a high species richness and high coverage. In contrast to the saxicolous (rock) lichens, the terricolous (soil) lichens are not heavy metal specialists, but widespread species of sparsely vegetated, acid and nutrient-poor grasslands, such as the fruticose lichens in the genera Cetraria, Cladonia and Stereocaulon as well as foliose lichens in the genera Peltigera. Base-rich metalliferous habitats generally have few lichens, as the productivity of the higher plants is greater here and the concentrations of free heavy metal ions lower.

9.4  Adaptations to the Environment 9.4.1  Heavy Metal Soils with Low Magnesium Concentrations Heavy Metal Elements  The heavy metals iron, manganese, zinc, copper, molybdenum and cobalt are essential for plant growth. However, in contrast to Fe and Mn, the other five elements are only needed in trace amounts. These, and the elements not needed by most plants (Pb, Cr, Cd, As, Se and Hg) can damage or even kill plants at concentrations in the soil as low as 1 mg g−1 (0.1 %) (Ernst 1974). Plants without any special adaptations cannot grow on metalliferous soils, or sustain considerable damage. However, particularly species of the Poaceae, Caryophyllaceae, Brassicaceae and Asteraceae families, as well as some ferns, bryophytes and lichens have evolved adaptive mechanisms to deal with these conditions, and are termed heavy metal plants or metallophytes. Hyper-accumulation  Some of these metallophytes accumulate large amounts of certain heavy metals in their biomass. Nickel, for example, can reach concentrations of up to 4 % dry mass (or up to 2 % in the case of Zn, Cu, Cr and Pb) (see Table 9.3). These plants are termed hyper-accumulators if they contain more than 0.1 % metal in their dry mass. Forbs generally accumulate much larger amounts of heavy metals than grasses, as is also the case for NaCl (see Sect. 1.4.1). Some examples from Central Europe are Thlaspi alpestre subsp. calaminare and Minuartia verna subsp. hercynica with 10 mg Zn g−1 and Armeria maritima subsp. halleri with 7 mg Zn g−1 in a Violetum calaminariae near Aachen (Ernst 1982). These plants were used by prospecting miners as indicators of heavy metals, and can in some cases even be used to extract metals (phytomining) or decontaminate soil (phytoremediation).

9.4  Adaptations to the Environment

741

Physiological Response  Free heavy metal ions in a reduced form (e.g. Fe2+, Mn2+, Cu+) are toxic, because they damage the electron transport systems in the cell (Shaw et al. 2004). They also form sulphides with the sulfhydryl groups of proteins, which can reduce enzyme activity at even low heavy metal concentrations. Some metal ions also replace more important elements with a similar ionic radius from their complexes (e.g. Mg2+ or Fe2+ replaced by Zn2+), which can disturb the functioning of essential enzymes (e.g. in the light and dark reactions of photosynthesis), as well as of membranes (Mysliwa-Kurdziel et  al. 2004; Rama Devi and Prasad 2004). Heavy metals can also hinder nitrogen uptake by plants. Metallophytes and non-metallophytes react similarly to heavy metal stress, but differ in their tolerance limits. Both take up higher proportions of metals from heavy metal-rich soils than from normal soils (see Fig. 9.1), and at high concentrations, the same elements damage both the non-metallophytes and the adapted metallophytes. However, the latter tolerate higher concentrations over the long-term better. A Silene vulgaris individual taken from zinc-rich soil, for example, grew well under experimental conditions with 25 mg Zn per litre nutrient solution, and much better than at low concentrations of 10 mg l−1 (see Fig. 9.2). At 50 and 100 mg l−1 there was a clear decrease in its productivity caused by the symptoms of toxicity. Under conditions that could also occur in natural heavy metal soils, the metallophyte was thus close to its limits of existence. Plants not adapted to Zn cannot survive under these conditions, and suffer even at low concentrations. The species that would suppress the light-demanding, low-growing Silene on less extreme soils are therefore no challenge to it on zinc-rich soils. The zinc form of Silene vulgaris reached its highest levels of productivity at around 30 mg Zn l−1, and therefore has a very high zinc demand. However, even under these conditions its productivity is still lower than that of the less tolerant normal form with a low zinc demand. At 50 mg, the normal form is killed whilst the zinc form continues to grow well, so that if these two forms compete under natural conditions, the Zn content of the soil liquid phase determines the outcome. In principal, the heavy metal plants therefore behave similarly to the halophytes (see Sect. 1.6.1), that require small amounts, or no, Na and Cl, but tolerate it at higher concentrations better than the glycophytes. It is physiologically inaccurate to talk of heavy metal plants in general, as the tolerance is highly variable depending on the element and the taxon. Metallophytes are usually only tolerant of the metals that are common in their habitat, and multiple tolerance only occurs in a few cases, for example, zinc-resistant plants are often also copper-resistant. Some metallophytes tolerate heavy metals by reducing their uptake (e.g. of Cu2+). These were termed excluders by Baker (1981), because they chelate the ions with organic acids exuded from the roots so that the resulting complexes cannot be transported across the cell membranes. Similar processes have been observed in the case of aluminium in many calcifuge plants (see Sect. 4.5.2.6 in Volume I; Ma et al. 2001). Mycorrhizae can also contribute to the reduction in heavy metal uptake by plants (Hüttermann et al. 2004). The heavy metal excluders contrast with the hyper-­ accumulators such as the Thlaspi species and Cardaminopsis halleri, which have leaf concentrations several times higher than in the soil (see Table 9.3).

9  Communities on Heavy Metal-Rich Soils

742

Table 9.3  Heavy metal concentrations in Central European soils and plants Plant species or soil Locality Content of ‘normal’ arable soil Non-metallophytes: vascular plants Metallophytes: bryophytes Weissia viridula Blankenrode Homalothecium Elpetal sericeum Metallophytes: vascular plants Viola calaminaria Blankenrode Silene vulgaris Langelsheim Minuartia verna Langelsheim Thlaspi alpinum Alps subsp. sylvium Thlaspi goesingense Austria Metallophytes: maximum contents

Zn 10–80

Cu 2–40

Pb 2–80

Cd Ni 0.1–0.6 3–50

Cr 5–100

50

10

1

0.05

1.5

1.5

9254 9745

19 30

1886 593

47 26

– –

– –

1210 2302 5291 –

25 45 573 –

19 195 100 –

– – – –

– – – 31,000

– – – –

– 25,000

– 13,700

– 11,400

– 560

12,000 38,100

– 20,000

From Ernst (1974), Reeves and Brooks (1983), Scheffer and Schachtschabel (2002, 2010) and Frey and Lösch (2004) (in μg g-1 dry weight). The locations in Blankenrode, Elpetal and Langelsheim (western Germany) are rich in Zn, Pb, Cu, the locations in the Alps have serpentine soils

If heavy metals are allowed to pass the root membranes and reach the shoot, metallophytes may use various mechanisms to immobilise and compartmentalise them, thus reducing their concentration in the cytosol and in particularly sensitive organelles. Heat shock proteins, which maintain the functioning of the cells under stress, are formed not only at high temperatures, but also under high heavy metal concentrations (Frey and Lösch 2004). Heavy metals can also be turned into complexes by organic chelating substances in the cell wall. Within the cell, heavy metal-­binding polypeptides (metallothioneines and phytochelatins) take over this function (Steffens 1990; E.  Beck 2002). The resulting organometallic complexes are probably sequestered into the vacuole via membrane-bound transporters, where they accumulate. Organic acids (particularly citric, malic and oxalic acid) and amino acids also form heavy metal complexes. Some metallophytes rid themselves of the large amounts of heavy metals by early leaf abscission, as is also observed in the case of NaCl in halophytes. The maintenance of heavy metal resistance is so energy-expensive that the productivity of the plants suffers, and many adopt a dwarf growth form. Metallophytes also typically have a more scleromorphic structure, higher anthocyanin concentrations, an unusually extensive root network (see Fig.  9.3), occasionally succulent leaves and often some developmental anomalies, e.g. deformed petals. In contrast to the physiological resistance, these characteristics disappear when the plant is grown on normal soil, and are therefore examples of phenotypic plasticity.

9.4  Adaptations to the Environment

743

Fig. 9.1  Almost all plant species growing in very zinc-rich soil take up more Zn (relative to the dry weight of their leaves) than when growing in ‘normal’, Zn-poor soil (Modified from Denayer-de Smet (1970)). With the exception of Thlaspi alpestre subsp. calaminare the heavy metal plants (Silene, Viola and Armeria) do not behave differently in this respect to the other species, including several tree species. However, they tolerate the high Zn concentrations better over the long term

Response of Lichens  Lichens growing on bare rock and slag heaps are exposed to much higher heavy metal concentrations than the higher plant specialists. The heavy metal concentrations in iron ore slag in the Harz Mountains and the lichens colonising them were studied by Lange and Ziegler (1963) and Noeske et al. (1970). They measured concentrations in the lichen thalli of up to 65 mg g−1 (6.5 %) copper and lead, although these elements were present at much lower concentrations in the substrate (0.5–2 %). In vascular hyper-accumulating species, concentrations up to 40 mg Zn g−1 are tolerated (Lambers et al. 2008). Heavy metals are accumulated

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9  Communities on Heavy Metal-Rich Soils

Fig. 9.2  The influence of increasing zinc concentrations in the culture medium on the biomass production of zinc-tolerant and zinc-sensitive populations of Silene vulgaris (From Baumeister (1954) in Kinzel (1982))

Fig. 9.3  The roots of Minuartia verna subsp. hercynica on a copper slag heap (right) and in soil with low heavy metal concentrations (left) (From Schubert (1954)). The heavy metal form is much more xeromorphic, and its root system is correspondingly larger and denser

particularly in the cell walls of the lichens, due to their high affinity to the cation exchange sites in the walls (Nieboer and Richardson 1980). Lichens achieve tolerance to heavy metals through various mechanisms, e.g. by immobilising them through the formation of complexes with lichen substances produced as secondary metabolites (depsidones) (Purvis et al. 1987, 1990). A. Beck (1999, 2002) found that the lichens in copper-rich habitats contain only certain copper-­tolerant green algae. Lichens of iron-rich rock and slag heaps generally lack

9.4  Adaptations to the Environment

745

iron-complexing lichen substances, which are otherwise widespread in lichens and it is assumed that they help to increase iron uptake (Hauck et al. 2007). In addition to high heavy metal concentrations, low pH values at the surface of ore slag and siliceous, ore-rich rock also exert a selective pressure on the lichen vegetation (Wirth 1972).

9.4.2  Heavy Metal Soils with High Magnesium Concentrations Serpentine soils are an extreme habitat for plants due to their high concentrations of the heavy metals nickel and chrome, as well as sometimes cobalt (see Table 9.1; Sasse 1979a, b; Proctor and Nagy 1992). However, their very low Ca/Mg ratio in the soil liquid phase (30 °C) and profit from warm summers (Schmitz 2002). The aggressively spreading neophyte Artemisia annua reacts with phenotypic plasticity to increasing interspecific competition as stands develop, rather than with increased mortality, allowing it to dominate in dense stands (Otte 1996; Brandes and Müller 2004). This feature may also explain why the banks of large rivers have particularly large proportions of introduced species (Planty-Tabacchi et al. 1996). In contrast, the seed rain from the handling of cargo in harbours plays a less important

10.2 Nitrophilic Bidentetea Communities of Still and Flowing Water Bodies

763

role in the spread of neophytes along river valleys, as the study of Brandes and Sander (1995) along the Elbe showed. Influence of Man  The Chenopodion rubri communities are undergoing a period of rapid expansion and change, due in large part to the new, empty habitats created by river engineering measures, as well as the transport of diaspores via the river traffic. The combinations of native species (e.g. Polygonum lapathifolium subsp. danubiale; syn. P. brittingeri) and archaeophytes that spread to the rivers of Central Europe with humans in prehistoric times (e.g. Conium maculatum or Portulaca oleracea), have been joined by recently introduced river bank neophytes (e.g. multiple Bidens species, the Amaranthus species and Artemisia annua), spreading via the ‘pathways’ of the riverbanks (see Sect. 9.5.2 in Vol. I). The character species Xanthium albinum evolved in the nineteenth century along the Elbe from the introduced X. saccharatum and, like Eragrostis albensis, is considered a neoendemic species (Wagenitz 1979; cf. also Wisskirchen 1995). Xanthium strumarium, in contrast, is an archaeophyte in southeastern Central Europe (Krippelova 1974). The Chenopodion rubri (like the Bidention tripartitae of standing water) have therefore gradually increased in species richness as a result of human activity, without the original native species as yet being completely suppressed by the neophytes (Brandes 1999a; Krumbiegel et  al. 2002; Schmitz 2002). As the nutrient supply in these habitats has also been artificially increased and the area suitable for colonisation expanded by the creation of groynes and through the greater fluctuations in water level in the narrowed river bed, the Chenopodion rubri has become increasingly anthropogenic. In contrast, highly modified rivers such as the Weser and the Rhine, with managed embankments and weirs, have a lower diversity of habitats and relatively few typical riparian plants, although they may support large populations of weed species and neophytes (Wisskirchen 1995; Hachtel et al. 1999; Schmitz 2002). Studies of the vegetation of the Loire and Allier in France give an impression of the variety of habitats for the therophyte flora along the larger rivers before they were straightened and modified by engineers (Wisskirchen 1995). Species Origin  Most riverbank neophytes that colonise the higher-lying and less frequently flooded riparian habitats are ornamental plants that have spread from gardens upstream and have been transported by the water (Adolphi 1996). They are typically tall and light- and nutrient-demanding, such as the perennial, 1.5–2 m high species of the genera Aster, Solidago, Helianthus and Rudbeckia. These species originate from North America and in some cases were planted in European gardens up to 400 years ago. They first grew spontaneously on rubbish heaps and similar habitats, but soon then spread to floodplains, particularly in areas where they were no longer flooded for prolonged periods. They initially colonised the drift lines and gaps in the willow or grey alder floodplain scrub, outcompeting the indigenous tall herb species there such as Urtica dioica and Senecio fluviatilis. Extended stands have developed particularly of Fallopia japonica (syn. Reynoutria japonica), Impatiens glandulifera and in places also Heracleum mantegazzianum, Solidago gigantea, S. graminifolia and Helianthus tuberosus, especially along rivers in the

764

10  Banks, Shorelines and Muddy Habitats Influenced by Man

lower parts of the uplands (e.g. Ludewig 1996, Nezadal and Bauer 1996). These species can sometimes compete fiercely with one another, and often spread vegetatively to quickly dominate any habitat that they happen to colonise first. Chance obviously plays a large role in their distribution. However, they are increasingly becoming a permanent part of the Central European vegetation and have allowed us to observe the formation of new plant communities within the space of only a few decades. These communities mostly belong to the order of the Convolvuletalia sepium (nitrophytic tall bank vegetation; see Chap. 14: no. 3.6.1). Moor (1958) named one association of this order ranging from Switzerland to Slovakia after the Himalayan balsam (Impatienti-Solidaginetum, see also Kopecky 1967a). Some introduced and some native tall herbs of river valleys were once seen as character species of floodplain willow scrub or forests, as they often occur in open areas in these habitats. However, they have since been recognised as independent, although very labile and light-demanding communities that are now assigned to the alliance of the Senecion fluviatilis (Tüxen 1955). The Cuscuto-­ Convolvuletum is particularly notable for its habit of winding around and climbing up tall forbs and low shrubs, covering them like a veil in late summer. Echinocystis lobata (wild cucumber), a Cucurbitaceae from northeastern North America, has spread in these communities in almost all river valleys of southeastern Central Europe. The spread of this species has been closely monitored as a carrier of the cucumber mosaic virus (see Slavik and Lhotska 1967). Chytrý (2009) lists for the Czech Republic four associations of the Senecion fluviatilis.

Chapter 11

Ruderal Communities on Drier Soils

The ruderal vegetation of rubble, rubbish heaps, eutrophic roadsides or other disturbed habitats on mesic to dry soil is even more strongly linked to human activity than the nitrophilic herb communities of temporarily wet habitats discussed in the previous chapter. The word ruderal is derived from the Latin rudus, which means rubble or broken stone. In contrast, the term segetal flora is used when referring to the wild plants (or weeds) of crop fields or gardens, where disturbance occurs more regularly and is often more intense than in rubble or roadside habitats. Whilst Krause (1958b) advised using ruderal to describe all ‘weed’ communities, including the Potentillo-­Polygonetalia of flooded grasslands (see Sect. 8.3.7.2) and the Atropetalia of clearings (see Sect. 11.3.2 in Volume I), we use the term here in its original and stricter sense.

11.1  Flora and Development Nitrogen-rich habitats were present in the natural landscapes of Central Europe only in very small areas, such as the drift lines of river banks and seashores, the surroundings of bird colonies and animal dens, in stands of nitrogen-fixing plants (e.g. black and green alder scrub) and in forests disturbed by windthrow or fire with accelerated humus decomposition. These were presumably the original habitats of the native ruderal species that made up part of the Central European flora before the influence of man. These are mostly species of perennial or winter annual communities, as well as the annual Bidentetalia communities, including Urtica dioica, Rumex obtusifolius, Chenopodium glaucum and C. album, Stellaria media, Silene latifolia subsp. alba, Chelidonium majus, Cynoglossum officinale, Galeopsis tetrahit, Verbascum thapsus, Artemisia vulgaris, Cirsium arvense and Lapsana communis. The Central European ruderal archaeophytes established here as a result of human activity, but before the beginning of the modern period. These are also mainly species of long-lived ruderal communities, even though they are mostly © Springer International Publishing Switzerland 2017 C. Leuschner, H. Ellenberg, Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats, DOI 10.1007/978-3-319-43048-5_11

765

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11  Ruderal Communities on Drier Soils

­therophytes, e.g. Bromus sterilis and B. tectorum, Hordeum murinum, Urtica urens, Fallopia convolvulus, Chenopodium polyspermum, Echium vulgare, Sisymbrium officinale, Descurainia sophia, Melilotus albus and officinalis, Malva neglecta, Verbena officinalis, Ballota nigra, Onopordum acanthium, Tanacetum vulgare, Carduus nutans and Arctium lappa. The majority of these archaeophytes probably reached Central Europe from the Fertile Crescent in the Middle East and from the Mediterranean region (Brandes 1999c). Generally, ruderal strategies among plants should be more widespread in regions with warm and dry summers than in Central Europe, because the vegetation is naturally sparser. This may also explain why the number of ruderal nitrophilic communities and their species richness increases from atlantic Western Europe to continental Eastern Europe (Braun-Blanquet 1964). Ruderal Neophytes  colonised Central Europe in the course of the increase in global travel in the modern period. These are mostly species found in short-lived ruderal communities (see Table 11.1). Scholz (1960) identified three groups of neophytes for the Berlin area, which demonstrate the constant influx of new species: (a) 1500–1787: e.g. Atriplex hortensis, Armoracia rusticana, Oenothera biennis; (b) 1787–1884: e.g. Atriplex tatarica, Chenopodium opulifolium, Sisymbrium altissimum, S. irio and S. loeselii, Matricaria discoidea; (c) from 1884: e.g. Atriplex sagittata, Chenopodium botrys, Amaranthus albus and A. blitoides, Lepidium densiflorum, Solidago canadensis. As shown by only these few examples, the Polygonaceae, Chenopodiaceae, Brassicaceae, Lamiaceae and Asteraceae are found in almost all groups. The Malvaceae, Onagraceae and Scrophulariaceae also play a notable role, whilst some families that are otherwise widespread in the native flora of Central Europe are surprisingly rare or never found as neophytes. For example, Scholz (1960) lists e.g. no Cyperaceae (apart from Carex hirta, which does not occur in true ruderal communities), Juncaceae, Liliaceae, Orchidaceae, Ranunculaceae or Gentianaceae, because these families contain only a few or no fast-growing nitrophilic species. Families particularly adapted to low-nutrient conditions such as the Ericaceae and Pyrolaceae or slow-growing succulents such as the Crassulaceae (apart from Sedum telephium), are also (unsurprisingly) completely absent from the ruderal neophytes. The neophyte species in Central Europe have very different origins but, like the archaeophytes, mainly come from landscapes that are warmer and drier than Central Europe. Many of the native ruderal species, in contrast, have ranges reaching far to the north and high into the Alps (see Table 11.1). On the Scandinavian coast, for example, there are large areas of drift line that were considered by Nordhagen (1940) as a natural habitat of many ruderal plants and arable weeds.

11.1  Flora and Development

767

Table 11.1  Annual and perennial ruderal communities of nitrogen-poor to nitrogen-rich soils in relatively dry areas of Central Europe, including their character species Weakly nitrophilic Fairly nitrophilic A Short-lived pioneer communities 1 Goosefoot pioneer community  Chenopodietum ruderale (mineral soils, in towns) (the first colonisers always 7 Amaranthus albus find sufficient nutrients on 7 A. hybridus rubble and rubbish tips as 6 Chenopodium strictum well as on roadsides and road 6 Ch. opulifolium and railway embankments) 7 Nicandra physalodes 6 Xanthium strumarium B More permanent annual communities 3 Prickly lettuce 5 Rocket-orache community community  Conyzo-Lactucetum (often  Sisymbrio-Atriplicetum succeeding 1) (continental, otherwise as 3) 5? Diplotaxis tenuifolia 7 Atriplex sagittata 4 Lactuca serriola (opt.) 6 A. oblongifolia etc. 6 Lepidium densiflorum 4 Wall barley community 6 Submediterranean rocket community  Bromo-Hordeetum  Descurainietum (building sites, sandy (thermophilic) fallow land) 5 Bromus sterilis 6 Descurainia sophia 5 Sisymbrium loeselii 6 Cnicus benedictus 4? S. altissimum 5 Hordeum murinum 5 Lepidium graminifolium C Perennial hemicryptophyte communities 9 Woolly thistle community 10 Scotch thistle community    Cirsietum eriophori (road  Onopordetum (in verges on limestone) villages, ± warm-continental) 5 Cirsium eriophorum 11 Grey cress community 5 Anchusa officinalis  Berteroetum incanae (road 7 Carduus acanthoides verges, embankments etc.) 7 Echinops sphaerocephalus 4 Berteroa incana

Strongly nitrophilic 2 Goosefoot-mallow community  Urtico-Malvetum (mineral soils, in villages) 9 Chenopodium murale 9 Ch. vulvaria 9 Malva neglecta 8 Urtica urens (?) (often in transition to 14)

7 Rocket-madwort community  Sisymbrio-Asperuginetum (at the entrances to limestone caves) 9 Asperugo procumbens 7 Sisymbrium austriacum 8 Lappula-madwort community  Lappulo-Asperuginetum (as 7, subalpine) 9 local: Asperugo procumbens 8 Lappula deflexa (7 and 8 are very rare)

15 Alpine dock community Rumicetum alpini (cattle resting places, subalpine)

9 Rumex alpinus (opt.) 9 Senecio alpinus

(continued)

768

11  Ruderal Communities on Drier Soils

Table 11.1 (continued) Weakly nitrophilic 12 Viper’s bugloss community  Echio-Melilotetum (on limestone gravel etc.) 4 Echium vulgare (opt.) 4 Melilotus alba 3 M. officinalis 4 Oenothera biennis 3 Oe. parviflora

Fairly nitrophilic 9? Hyoscyamus niger 8 Onopordum acanthium 5 Verbascum densiflorum

Strongly nitrophilic 16 Subalpine goosefoot community  Chenopodietum subalpinum 9 Chenopodium bonus-henricus

13 Tansy-mugwort community  Tanaceto-Artemisietum (embankments, roadside verges etc.) 7? Artemisia vulgaris (opt.) 5? Linaria vulgaris (?) 5 Tanacetum vulgare

14 Markery community  Balloto-Chenopodietum (sides of village streets) 8 Ballota nigra (opt.) 9 Chenopodium bonus-henricus

Based on data from Lohmeyer, Tüxen, Oberdorfer, F. Runge and others. The names of common communities are in bold. An overview of the alliances (and orders) and their character species is given in Table 11.2. The numbers in bold refer to the communities, and the numbers in normal print in front of the species names are the indicator value for nitrogen (N, from Ellenberg et al. 1992)

11.2  Environmental Conditions and Habitat Classification Three main factors shape the habitats of ruderal plants: high nutrient levels (particularly of nitrogen), high light levels, and high disturbance levels, which can mean repeated damage to or even destruction of the plant community and soil movement. If the stand is disturbed every year, then the ruderal vegetation is dominated by annuals, whereas if the disturbance is less frequent then it can contain biennial to perennial tall herbs or even shrubs. Brandes (2007) defines ruderal vegetation as the mainly herbaceous communities of strongly altered and/or disturbed habitats that are not cultivated farmland. In contrast, the annual or otherwise regular disturbance from tillage and herbicide application that is typical for arable habitats cause segetal communities to develop (see Chap. 12), which are dominated by annual species. These can therefore be seen as ruderal communities adapted to extreme disturbance intensities. Although the ruderal vegetation is naturally restricted to mesotrophic to eutrophic habitats, its communities can still be clearly differentiated according to their nutrient supply. They can also be divided into those adapted to more or less dry conditions (discussed in this chapter), and those adapted to damp to wet conditions (discussed in Chap. 10).

11.3 Vegetation

769

11.3  Vegetation Ruderal communities in the strict sense (i.e. excluding segetal vegetation) can be divided into two major groups. These are the short-lived communities in the class Sisymbrietea (previously placed in the class Chenopodietea; see Table 11.2, top) and largely perennial communities in the class Artemisietea (see Table 11.1, bottom). Both contain numerous communities linked to specific climatic and edaphic conditions, in a similar manner to the range of forest, heathland or dry grassland communities that are not mainly determined by the presence of groundwater. This link can be seen in the works of Tüxen (1937, 1950a, b), Lohmeyer (1950, 1970a), Oberdorfer (1957), Grosse-Brauckmann (1953a, b), Düll and Werner (1956), Krause (1956, 1958b), Ubrizsy (1956), Oberdorfer (1992–1998), Tillich (1969), Brandes (1980, 1989, 1991, 1999), Otte and Ludwig (1990), Pysek (1991), Chytrý (2009) and many others. Table 11.1 shows the Central European ruderal communities arranged according to life span and soil nutrient levels. Table  11.2 gives an overview over the corresponding alliances and orders, their character species and common accompanying species.

11.3.1  Ruderal Communities of Summer and Winter Annuals 11.3.1.1  Short-Lived Ruderal Communities on Dry Rubbish Dumps One of the first colonisers of dry rubbish dumps in the towns of the Upper Rhine region is the Chenopodietum ruderale (Oberdorfer 1957), with character species such as the thermophilic Amaranthus and Chenopodium species (see Table 11.1). The summer annual species are later replaced by winter annuals, i.e. species that germinate in autumn and remain green over the winter, such as Lactuca serriola and other character species of the Conyzo-Lactucetum. In Stuttgart, the Descurainetum sophiae with more submediterranean species also occurred alongside this. In Berlin, Düll and Werner (1956) described a similar Sisymbrium community as the most common pioneer community of dry rubble heaps, possibly because the lack of nutrients prevents the Chenopodium species from developing. In Halle/Saale (eastern Germany), Atriplex acuminata played a particularly large role in ruderal habitats, leading Oberdorfer (1957) to identify this species as a character species of the Sisymbrio-Atriplicetum (see Table 11.1) with a range strongly limited to the east of Central Europe. Other studies of the vegetation of rubbish and waste dumps can be found e.g. for central Germany in Klotz (1982) and Gutte (1986), for the Czech Republic in Pysek (1974, 1977), Kopecky (1980–1984) and Chytrý (2009), and for Austria in Mucina et  al. (1993) (see also the overview in Oberdorfer 1983a, b; Wittig 2002).

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Table 11.2  Alliances and orders of major ruderal communities on mesic to dry soils, with their character species and common accompanying speciesa A and B, 1 to 8 Rocket communities sensu lato: Sisymbrion, Sisymbrietalia (class Sisymbrietea) 5 Anthemis austriaca 4 Bromus tectorum 5 Lepidium virginicum 5 A. cotula 4 Cardaria draba 5 Plantago indica 9 Atriplex hastata 5 Conyza canadensis 5 Sisymbrium irio 6 A. heterosperma 6 Crepis tectorum 7 S. officinale 6 Barbarea verna 8 Datura stramonium 7 Solanum luteum 5 Bromus lepidus 6 Lappula squarrosa etc. C 9 and 10 Scotch thistle communities sensu lato: Onopordion, Onopordetalia (class Artemisietea) 8 Artemisia absinthium 7 Malva alcea 6 Reseda luteola 6 Carduus nutans 8 M. sylvestris 5 Stachys germanica 6 Crepis pulchra 8 Marrubium vulgare 6 Verbascum blattaria 7 Cynoglossum officinale etc. C 11 and 12 Melilot communities: Dauco-Melilotion (Onopordetalia) 5 Cichorium intybus 5 Pastinaca sativa 4 Picris hieracioides 4 Daucus carota 4 Potentilla intermedia 4 Rumex thyrsiflorus C 13 and 14 Burdock communities: Arction, Artemisietalia (class Artemisietea) 9 Arctium lappa 8 Chelidonium majus 9 Leonurus cardiaca 8 A. minus 8 Conium maculatum 7 Parietaria officinalis 9 A. tomentosum 7 Dipsacus fullonum 6 Solidago canadensis 9 Armoracia rusticana 8 Geranium pyrenaicum 7 S. gigantea 9 Carduus crispus 9 Lamium album etc. C 15 and 16 Alpine rock communities sensu lato: Rumicion alpini, Glechometalia (class Galio-Urticetea) 9 Cerinthe glabra 8 Cirsium spinosissimum Based on data from Lohmeyer, Tüxen, Oberdorfer, F. Runge and others, cf. Table 11.1 a The assignment to classes has not yet been completely clarified; therefore, their character species are not given here. In addition to their character species, each of the associations 1 to 16  in Table 11.1 also contains several of the species listed here, as well as numerous generalist species that are found in various different communities, e.g. Poa species, Dactylis glomerata, Urtica dioica or Chenopodium album

11.3.1.2  Short-Lived Ruderal Communities on Rubble Short-lived ruderal plants grow particularly well on the debris of demolished buildings, and in Central Europe have never had such a large area of habitat available as in the years 1943–1945. They were astonishingly quick to colonise newly created piles of rubble, and despite the initial complete lack of seeds and plants on the humus-free heaps of loose tiles, plaster and burnt remains of wood, the first seedlings of higher plants appeared within only a few months. These were mostly species of Asteraceae and other anemochorous plants from the Sisymbrion alliance, or plants of forest clearings such as Epilobium angustifolium, which grew well particularly on wood ash. The habitat conditions provided by

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the rubble depended on the amount of fine earth, the water permeability of the ­rubble, its slope and aspect, as well as the general climate. The plant communities spontaneously growing on rubble therefore did not always develop in the same way. Brandes (1985) found that the succession on debris was mainly driven by the changes in soil moisture. Düll and Werner (1956) found that ruderal plant community succession in Berlin involved a large number of continental species. The rubble was often dominated by vegetation types of open ground for several years, whilst mesophilic meadow vascular plants and bryophytes established slowly and only in areas with damp microclimates. More nutrient-demanding ornamental plants also colonised slowly, or not at all, as did the garden and arable weeds that had been rare in these urban areas for decades. These communities did not have time to develop woody vegetation before the succession was interrupted by the clearing of the rubble, probably because the moisture levels in the substrate and air were still too low to allow the growth of trees. In contrast, Engel (1949) found many more trees and shrubs establishing on rubble in the much more oceanic climate of Münster in North Rhine-Westphalia. These communities were dominated by the numerous ruderal species of damper soils that were so rare in Berlin. However, these were soon replaced by species of agricultural grassland (Arrhenatheretalia). Bryophytes covered even the areas exposed to full sunlight. The woody species included many plants of early forest succession and ornamental plants, particularly berry-producing ornithochorous and anemochorous species such as Betula pubescens and B. pendula, Populus tremula and Alnus glutinosa. The ruderal rubble vegetation of Brunswick, Dresden and Stuttgart was intermediate between these two extremes, as shown by Düll and Werner (1956, and literature cited within). However, the colonisation always began with the short-lived ruderal communities described here. 11.3.1.3  S  hort-Lived Ruderal Communities in Villages and in Natural Habitats All of the ruderal communities mentioned so far consist of numerous neophytes (cf. Table  11.1), a few archaeophytes, and practically no species native to Central Europe. The colonisers of loose, stony rubble from buildings and roads thus appear to be a relatively new element in the Central European flora and vegetation, in the same way that settlements constructed largely of stone have only been present since the beginning of the modern period in this region. Similar habitats in villages, in contrast, are colonised by archaeophytes. The species of the Urtico-Malvetum neglectae of villages (see Table 11.1) have not only been in Central Europe for longer, but are also more nitrophilic than the summer and winter annual ruderal plants of cities. This is probably the main reason why they are not found in larger settlements. The waste dumps, corners of farmyards, bases of walls and similar habitats in villages were (and in some cases still

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are) heavily fertilised with excrement and other sources of nutrients largely due to the free-ranging domestic animals (see Fig. 13.8). Kopecky (1986) also considers the scratching and selection by hens to be responsible for the promotion of annuals in the Malvetum and the slowing of the succession to the Arction lappae. The same is true for the now very rare Marrubium vulgare stands (Raabe and Brandes 1988). The rubble of older village buildings is also richer in fine earth, because these sites were generally plastered with loam instead of with mortar, and usually built without stones or bricks (Ellenberg 1990). Their rubble thus provided a fertile and damp substrate. In the last 30 years, Malva neglecta has started to reappear in city centre green areas due to the increase in dog faeces deposited there, and forms communities along the edges of many grassy patches that are intermediate between the ‘classical’ Urtico-Malvetum and trampled or grassland communities (e.g. Brandes 1985; Wittig 2002). The unique but unfortunately increasingly rare Asperugo procumbens (madwort) communities at the entrances of limestone caves and under rocky overhangs in southwestern Germany and the Alps are even older than the Urtico-Malvetum. Oberdorfer (1957) distinguished a colline-montane Sisymbrio-Asperuginetum and a ‘natural’ subalpine Lappulo-Asperuginetum association. The Asperuginetum, Urtico-Malvetum, Chenopodietum and Sisymbrietum communities all require open ground that is (almost) completely free of vegetation. In places where the vegetation is not always destroyed, perennial species will take up increasing amounts of space, so that the short-lived stand eventually disappears and is replaced by a longer-lived ruderal, grassland or scrub community. The Bromo-Hordeetum stands (see Table  11.1) are somewhat intermediate between the short- and long-lived ruderal communities. They can be found in both villages and cities, but always on flat or slightly sloping, generally sandy areas, e.g. on arable fields that have been left uncultivated for building developments, or at the edges of roads or tracks that are rarely disturbed. Directly after World War II, they dominated the rubble fields of urban areas decimated by bombing. This is one of the few grass-rich ruderal communities, and is clearly visible in late summer from its straw-coloured vegetation, which stands out from the usually still deep green weed stands. In particular, Bromus tectorum, Bromus sterilis and Hordeum murinum can form fairly dense stands here. This community can persist for several years in its relatively nutrient-poor habitat, although it develops each year anew from seed. However, like the other communities of the Sisymbrion, it is eventually replaced by grassland or trampled communities, or by ruderal communities dominated by biennial or perennial species. Wittig and Ou (1993) studied the Hordeetum murini in seven cities across a climatic gradient from Brussels to Warsaw. They found that the average indicator values for temperature, nitrogen and soil pH increased, corresponding to the increase in continentality, and the moisture value decreased. Similar gradations are probably also found in other widespread ruderal communities, e.g. in the communities discussed in the next section.

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11.3.2  Communities of Perennial Ruderal Plants 11.3.2.1  T  hermophilic Thistle and Bugloss Communities (Onopordetalia) The largely perennial ruderal communities are even more variable than the short-­ lived communities. They can be divided into four alliances based on the climate of their habitats: 1. the thermophilic and relatively drought-tolerant Onopordion acanthii (no. 9 and 10 in Table 11.1), 2. the less nitrogen-demanding Dauco-Melilotion (no. 11 and 12), 3. the more nitrophilic Arction lappae of intermediate moisture and temperature (no. 13 and 14), and 4. the Rumicion alpini (no. 15 and 16) of sometimes highly nitrophilic alpine ruderal areas, i.e. adapted to cool climates and related to the tall forb communities. All of these alliances belong to the class Artemisietea vulgaris (i.e. perennial herbaceous nitrogen-rich communities), but within this to three different orders, namely the Onopordetalia (no. 9–12), the Artemisietalia (no. 13 and 14) and

Fig. 11.1  An Onopordetum community along a garden hedge in Kurdějov (eastern Czech Republic). Onopordum acanthium (right), Leonurus cardiaca (left), Arctium minus (with broad leaves) can be seen, with Sambucus nigra in the background (Photo: Iltis)

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the Glechometalia (no. 15 and 16, see Table 11.1). However, the classification of the ruderal communities is still actively debated and e.g. treated differently by Pott (1995) compared to Oberdorfer (1992–1998), Mucina et  al. (1993) or Chytrý (2009). The Sisymbrion communities discussed in the previous section are the most closely related to the Onopordion. They also colonise free-draining and periodically dry substrates, where they are free from competition from forest plants for longest. Tillich (1969) even found them on rubbish dampened by groundwater near Potsdam. The Onopordetalia communities often develop from pioneer stages of the Sisymbrion. Following the Sisymbrion, Düll and Werner (1956) found the Echio-Melilotetum developing on large areas of rubble in Berlin. Until around 40 years ago, this was the most common association within the Central European Onopordetalia. These communities often occupy gravelly railway embankments and storage yards, as long as they are calcareous and have not been treated with herbicides (Brandes 1991b). They can also be found in limestone quarries if loose, dry debris accumulates. Their natural habitat is probably the carbonate gravel plains of Alpine rivers that are particularly high above the water level. They can also develop temporarily in the scree below carbonate cliff faces, before this is colonised by grassy and scrub communities. Many species of the Echio-Melilotetum are also not native to Central Europe, so that this association should not really be considered part of the natural landscape. Of all the ruderal communities, the Echio-Melilotetum is one of the most colourful and structurally rich. The tall, delicate flower-heads of the white and yellow Melilotus rise above the intensive violet of the Echium. These are dotted with loose groups of green-gold Reseda, crimson Carduus nutans and light yellow Oenothera biennis (evening primrose), the fragrant flowers of which open in the evening with an audible pop. The small and light-green leaves of the plants and the presence of legumes, which is otherwise unusual for ruderal communities, indicates a lack of soil nitrogen in these communities, as shown by Kronisch (1975) (see also the N values in Table 11.1, no. 11 and 12, which rarely exceed 4). The Onopordetum (see Fig. 11.1), the eponymous association of the Onopordion alliance, grows somewhat more vigorously. It can be found in its best-developed form on deep, fertile Chernozems in Eastern and Southeastern Europe, and similar habitats in the drier areas of central and eastern Central Europe. Tall and stable thistles like Onopordum itself, as well as Carduus and Cirsium species or even the rarer Echinops sphaerocephalus dominate this community, which was characteristic for farm yards and greens of continental villages on loess, but is in decline in many regions due to the modernisation of the villages (Chytrý 2009). Once they have established, they are able to persist for a relatively long time, especially as many of the species are very adaptable and can survive periods of low nutrient levels. Nevertheless, they do not colonise new areas easily, as according to sowing experiments by Krause (1950), most of the species germinate poorly, or die soon after germination. Krause explains this as the effect of fungal infections of the young plants. However, the risk of such a fungal attack is possibly lower in the dry and

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warm habitats of the Onopordetum than in the temperate climate of the botanical garden in which the germination experiments took place. All of the species of the Onopordetum are more nitrophilic than those of the Echio-Melilotetum (N 5–8 in Table 11.1). The Cirsietum eriophori (see Table 11.1, no. 9) also belongs to the Onopordion and is a community of calcareous, sheep-grazed areas that until recently was relatively understudied. With its feathery leaves and long-haired seed heads, even Cirsium eriophorum (woolly thistle) looks more like an ornamental plant than a weed. Like almost all species of Onopordion, and in fact most ruderal plants, it is avoided by grazing livestock. This is also one of the main reasons why villages with low-intensity livestock farming are (or were) often particularly rich in Onopordion communities. This community declined faster than many other ruderal communities, although it doubtlessly profited from the general eutrophication of the landscape. It suffered instead mainly from the increasing urbanisation of villages, with the greater covering of concrete and brick and the loss of free-roaming livestock. In relation to the Onopordetalia acanthii, we should briefly mention a group of semi-ruderal, semi-open pioneer and semi-dry grasslands often dominated by Elymus repens (couch grass, syn. Elytrigia or Agropyron repens). Despite a general lack of good character species, these stands were often assigned to their own class, especially as Braun-Blanquet (1961) had emphasised their uniqueness, later supported by Passarge (1989). Today, most authors classify these stands as an order (e.g. Agropyretalia intermedio-repentis) alongside the ruderal communities. However, this vegetation type is somewhere between the ruderal and segetal communities on the one hand, and the dry grasslands on the other, and is therefore not only ecologically, but also phytosociologically intermediate. Even phytogeographically, its range is somewhere between continental and oceanic. The above description is valid for the most common association of this vegetation type, the Convolvulo-Agropyretum repentis, which Pott (1995) described as an ‘initial community that is difficult to describe but found almost everywhere’. It is common in large parts of Central Europe on roadsides, waste places, along fences and abandoned fields (e.g. Chytrý 2009). Like Convolvulus and Elymus, most of the species in this community are able to spread quickly via vegetative reproduction, e.g. Poa angustifolia and P. compressa. This is also the case for the Poo-­Tussilaginetum, which spreads on calcareous and periodically damp, loamy Leptosols, as well as for the more continental Falcario vulgaris-Elymetum repentis of calcareous, loamy field edges. These, as well as associations dominated by Cardaria draba, Diplotaxis tenuifolia or Chondrilla juncea, make up the alliance of the Convolvulo-Agropyrion repentis. 11.3.2.2  Wormwood and Burdock Communities (Artemisietalia) In damper and cooler climates or on less well-draining soils, the Onopordion is replaced by the Arction lappae (see Fig. 11.2). Only Arctium minus (lesser burdock) thrives in warm and dry habitats, whilst its large-leaved relatives (Arctium

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lappa and A. tomentosum) are mainly found in wetter areas. The other species of the Arction communities are also more mesic, such as the tender-leaved Chelidonium majus, the shade-tolerant Lapsana communis, Conium maculatum, Armoracia rusticana with its long leaves, Urtica dioica and Lamium species. The most common association within the Arction is the Tanaceto-Artemisietum, which can persist for periods of several years on rubbish and rubble heaps, as well as on road embankments (see Fig. 11.2). It usually develops after Sisymbrion communities, particularly the Bromo-Hordeetum (see Table 11.1, no. 4). In areas with higher nutrient concentrations along village roads and the bases of walls, it is replaced by the Balloto-Chenopodietum (no. 14), which succeeds the Urtico-­ Malvetum neglectae (no. 2). The Balloto-Chenopodietum is mainly found in western and northern Central Europe. It is replaced in areas with more continental climates by the Leonuro-Arctietum tomentosi. The character species for this community include Leonurus cardiaca, Conium maculatum and Arctium tomentosum. In the drier lowlands of southeastern and eastern Central Europe, Artemisia absinthium (common wormwood) with its grey, absinth-scented leaves forms perennial ruderal communities; this species escaped from cultivation and naturalised in these communities rich of thermophilic and drought-tolerant species (Chytrý 2009).

Fig. 11.2  Small remaining patch of ruderal vegetation near a parking lot at the margin of the built-­up city center zone of Hanover with Galinsoga parviflora, Artemisia vulgaris, Lapsana communis and other nitrophilic species

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11.3.2.3  Subalpine-Alpine Ruderal Communities (Glechometalia) The Balloto-Chenopodietum community mentioned in the previous section is, or was, found throughout Central Europe as well as in all elevational zones up to the highest settlements. With increasing elevation, the accompanying species of Chenopodium bonus-henricus change, so that Oberdorfer (1992–1998) distinguishes not only western and eastern types of this community, but also planar-­ colline, montane and oreal-subalpine forms of the Chenopodietum boni-henrici in the Arction order (which includes the Balloto-Chenopodietum). The subalpine form of this community is, however, more similar to the alpine ruderal vegetation discussed in Sect. 5.3.11.4 than the Arction of lower elevations. It can therefore be assigned its own association, the Chenopodietum subalpinum. Together with the Rumicetum alpini (see Fig. 8.33), it belongs to the subalpine-alpine alliance of the Rumicion alpini, (see Table 11.2). These subalpine communities are even more mesophilic than the Arction communities. However, there have been no comparative studies of either the water or the nutrient supply of these communities, with the exception of the N supply of the Rumicetum alpini mentioned in Sect. 5.3.11.4.

11.4  Adaptations to the Environment Walter (1963) described typical ruderal plants as ‘nitrophilic’, as nitrogen promotes their growth more than it does in plants of other habitats. Indicators of extremely high nitrogen supply are in fact found almost exclusively in the ruderal classes of the Artemisietea and Bidentetea (30 of the 41 Central European species with EIV-N = 9, Brandes 1999b). The higher the N value in Ellenberg et al. (1992), generally the higher the nitrogen concentration in the biomass and nitrate reductase activity in the leaves or roots, as shown by Gebauer et  al. (1988) in 48 Central European plant species. Nitrate promotes not only the growth, but also the germination of many nitrophytes (Gassner 1915, Bischoff 1996). Most ruderal habitats are, however, not only rich in nitrogen, but also in phosphorus, potassium and other nutrients released by faeces, rubbish or other waste accumulated there. For Urtica dioica (stinging nettle), P even seems to be the growth limiting nutrient, rather than N. Experiments by Mayser (1954, in Walter 1963) and others showed that the form of nitrogen does not affect the growth of ruderal plants, i.e. ruderal and segetal plants grow equally well with nitrate as they do with the same amount of N as ammonium (see Fig. 12.8). A high indicator value for N should not therefore be seen as synonymous to being a nitrate-demanding species, as is occasionally stated. The close relationship between ruderal plants and nitrogen-rich habitats is mainly caused by the following factors. Firstly, these species are able to exploit high ­concentrations of nitrate (and probably also ammonium) in the soil liquid phase through the formation of roots with high maximum uptake rates (maximum rate of nutrient inflow, Imax). Secondly, experiments by Janiesch (1973a, b) showed that the nitrophytes Aegopodium podagraria and Anthriscus sylvestris not only absorbed

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Fig. 11.3  The leaves and roots of several ruderal species at the edge of a village road. Modified from Grosse-Brauckmann (1953a). Most species have mesomorphic to hygromorphic leaves. Left (south-facing): Verbena officinalis, Chenopodium bonus-henricus, Artemisia vulgaris, Ballota nigra (Balloto-Chenopodietum). Right (shadier): Rumex obtusifolius, Urtica dioica and Lamium album

large quantities of nitrate, but also stored this in the vacuole and gradually metabolised it during the growing season. These species also have a high protoplasmic tolerance of nitrate. Many nitrophytes (e.g. Solidago canadensis and Elymus repens) additionally have nitrogen-­storing rhizomes or stolons that can supply N for growth during periods of low availability via efficient internal cycling (Werner 1983). Finally, nitrophytes achieve fast growth rates through an allocation strategy in which N is principally used to form thin and short-lived leaves with high photosynthetic activity per unit mass (van Arendonk et al. 1997), which is a strategy seen at its most extreme in annuals (Garnier 1992). For this reason, the majority of ruderal plants have delicate, meso- or hygromorphic leaves even when growing in sunny and relatively dry habitats. This is true in particular for Artemisia vulgaris, Ballota nigra, Chenopodium bonus-henricus, Lamium album, Malva neglecta, Urtica urens and Verbena officinalis, which belong to the Sisymbrion alliance. Grosse-Brauckmann (1953a) characterised their habitats in settlements as xerophytic, especially as the air humidity is often low due to the relatively high emission of long-wave radiation from heated house walls (see Fig. 11.3). Despite their mesomorphic leaves, he found these ruderal plants to have conservative and xerophytelike transpiration patterns. Ruderal plants also typically produce a large number of seeds that ripen relatively early, and can produce these seeds even under poor growing conditions (Boot et al. 1986). The creation of a long-lived diaspore bank and the ability to germinate under a wide range of conditions further contribute to the exceptional ability of ruderal plants to persist in disturbed habitats (Grime 2001). Many ruderal plants regenerate quickly after mechanical damage and can also spread via root fragments (Dietz et al. 1999).

Chapter 12

Vegetation of Arable Fields, Gardens and Vineyards

12.1  Flora and Development 12.1.1  Flora Cropland and gardens now cover large areas of Central Europe (they make up e.g. 30 % of the land area in Germany). As described in Chap. 3 in Volume I, these areas have in some cases been cultivated since the Neolithic, i.e. when humans first became sedentary in Central Europe some 7500–6000 years ago. The arable weed communities that developed among the crop plants as the bane of farmers, and in some cases still persist despite modern farming techniques, are similarly ancient. Seeds of numerous plant species common in crop fields today have been found in excavations of Neolithic and Bronze Age settlements (Willerding 1983, 1986a, b, 1988). A comprehensive list of species that have long accompanied crop plants in Switzerland can be found in Jacomet et al. (1991). Then, as now, the crop plants sown by humans were not able to completely fill the cultivated area and suppress the other species. As a result, around 300–350 plant taxa have used the highly anthropogenic crop fields, gardens and vineyards as their habitat (Ellenberg 1950; Arlt et al. 1991). These arable weeds are also referred to as segetal species (from the Latin segetalis = belonging to the seed). Only around 20 of the 300–350 segetal species, i.e. less than 10 %, cause significant yield reductions in agriculture and can be considered ‘problematic’ weeds today (Eggers 1984). Rademacher (1948) defined arable weeds as ‘plants forming communities with crop plants that tolerate, are promoted by, or even depend on cultivation’. Their tolerance of regular human disturbance in the course of crop management distinguishes the arable weeds from the ruderal plants (see Chap. 11). The latter occur predominantly in anthropogenic habitats, but only dominate when human disturbance does not occur every year, and are therefore restricted to habitats with little or no human management.

© Springer International Publishing Switzerland 2017 C. Leuschner, H. Ellenberg, Ecology of Central European Non-Forest Vegetation: Coastal to Alpine, Natural to Man-Made Habitats, DOI 10.1007/978-3-319-43048-5_12

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Of the approximately 300 taxa occurring in crop fields in southern Germany, only 140 can be considered true segetal plants (Ellenberg 1950). The other 50 % occur mainly in other habitats, such as roadside verges, in ruderal communities, dry grasslands, stream banks or muddy patches. Only three forest species (Equisetum sylvaticum, Holcus mollis and Rubus caesius) and around 20 meadow plants regularly occur in arable fields, emphasising the fundamentally different disturbance regimes in arable and grassland habitats.

12.1.2  T  he Origins of Arable Weeds and Changes in the Segetal Vegetation Since the Neolithic Many Central European arable weeds are native species, naturally occurring in the short-lived herbaceous communities of river beds, ponds and drift lines that dry out in summer, as well as foredunes and forest gaps created by storms or fire. They also exploit temporary bare soil created on eroding banks of rivers, larger lakes and coastal areas, as well as on patches of fine earth on scree slopes and mudflows that reach down into the lower montane belt of mountains. Krause (1956) states that the burrows and tracks of wild animals would have also provided an opportunity for the evolution, persistence and spread of plant species that cannot survive in forests, marshland and other permanently dense vegetation types. Schneider et al. (1994) list around 100 segetal species that naturally occur in the Central European vegetation and should therefore be considered as apophytes (i.e. species of native origin that benefit from association with humans). Nevertheless, none of the natural habitats of Central Europe would have supported species assemblages similar to those of modern arable fields. Like the meadows and fertilised pastures, the segetal communities are therefore new formations created unintentionally in the course of human activity. These are even more artificial than the communities of agricultural grassland, as they contain numerous species that would not occur in Central Europe without human intervention (Weinert 1973; Arlt et al. 1991; Pötsch 1991). Like the wild forms of the cereals themselves and the practice of cereal cultivation, a number of weeds of cereal crops originate from the steppe regions and semi-­ deserts of the Middle East (Weinert 1973). In keeping with their continental origins, these species grow best in arable fields on base-rich and dry soils in Central Europe. However, the majority come from the Mediterranean region and would similarly not be able to survive in Central Europe without the creation of habitats through farming activities. These species were brought to Central Europe in several waves, for example with the Corded Ware farmers in the Neolithic, with the metalworking cultures of the Bronze and Iron Ages and particularly with the spread of Roman agriculture north of the Alps (Jacomet et  al. 1991; Schneider et  al. 1994). Some cereal weeds (e.g. Adonis aestivalis and Scandix pecten-veneris) only spread significantly in Central Europe after the beginning of the modern era (Küster 1985).

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781

Some of the weeds of highly acidic soils (such as Spergula arvensis, Arnoseris minima and Hypochaeris glabra) are phytogeographically quite different, as they probably spread to Central Europe from the coastal dunes of the western Atlantic and western Mediterranean (Manthey 2003). The majority of arable weeds in Central Europe are thus at the northwestern edge of their ranges. The number of arable weed species in Central Europe was still relatively low at the beginning of Neolithic agriculture. Willerding (1986a) states that this rose during the Neolithic from 8 to 120, and in the course of the Bronze Age increased almost linearly to 157 species. 212 species have been found from the pre-Roman Iron Age, 253 during the Roman Empire, and at least 305 from around the year 1500. Centaurea cyanus (cornflower), for example, first occurred in large numbers in the Netherlands and northwestern Germany around 1000 AD (Behre 1993) (see Fig. 12.1). One major factor in the development of the modern arable weed communities was the transportation of their diaspores with crop seed, as well as with transhumant livestock that were driven onto the harvested crop fields to graze the stubbles. Recent increases in long-distance global transportation have also brought new species from other continents, particularly from Northern and Central America. The migration of neophytes into crop and garden habitats continues today. For example, Galinsoga ciliata was introduced around the 1850s, but only began to spread in western Central Europe 100 years later and is now well established in many regions. Further neophytes among the crop and garden weeds include Veronica persica, Amaranthus retroflexus, Galinsoga parviflora, Matricaria recutita, M. discoidea, Epilobium ciliatum, Conyza canadensis, Chenopodium ficifolium, Oxalis fontana and O. corniculatus, Alopecurus myosuroides, Cyperus esculentus and various Sisymbrium species (Ries 1992; Belde et al. 2000; Rumpf et al. 2004). The establishment of these novel species is facilitated in cropland and gardens by the continuous creation of open ground, as well as the protection from competition from most of the native perennial species. The arable weeds, like the crop plants themselves, have undergone greater changes over previous centuries than is now generally recognised. As already mentioned in Sect. 3.4.1  in Volume I, the segetal communities must have been more similar to those of grassland during the early years of arable cultivation and in the three-field system of the Middle Ages. In the Neolithic and particularly from the tenth century onwards, there were cereal weeds such as Agrostemma githago, Anthemis arvensis, Centaurea cyanus and Neslia paniculata, as well as some ‘root crop weeds’ such as Aethusa cynapium, Euphorbia helioscopia and Solanum nigrum (Willerding 1973). However, among the numerous weed species that have been identified e.g. by Maier (1988) in the Neolithic settlement of Ödenahlen am Federsee (southwestern Germany), these cereal weeds were very rare, and instead there were many indicators of mesic to relatively dry nutrient-rich soils and ruderal plants. This was probably due to the prevailing summer cropping, particularly of rivet wheat. In early arable systems, part of the crop field was left fallow and grazed, so that arable and grassland vegetation were more closely interlinked than they are today (see Fig. 3.24 in Vol. I). In addition, primitive ploughs such as the wooden scratch

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12  Vegetation of Arable Fields, Gardens and Vineyards

Fig. 12.1  Edge of a cereal field in eastern Poland with species-rich weed community; Agrostemma githago, Centaurea cyanus and Tripleurospermum inodorum are visible (May 2000)

plough only loosened the soil but did not turn it, allowing more perennial grassland species to persist. It is only in rotational cultivation of arable and grass leys, e.g. in the southern Black Forest or in Schleswig-Holstein, as well as in clover leys that meadow plants were until recently still regularly found as arable weeds. However, most of these do not tolerate frequent tilling and harrowing. The introduction of more intensive and effective ploughing thus led to an increase in the proportion of annual species in segetal communities, particularly those that can germinate at any time of the year, such as Thlaspi arvense and Capsella bursa-pastoris (Salzmann 1939). The modern segetal vegetation that is dominated by annuals began to develop around 200 years ago, as the once almost ubiquitous fallow year became no longer necessary due to the introduction of more effective fertilisers, and was replaced by the cultivation of root or forage crops. Root crops, such as potatoes or turnips require intensive ploughing and are traditionally rotated with cereals. Crop rotations without a fallow year, such as root crop – summer cereal – winter cereal are described as improved three-field crop rotation, and make conditions even harder for the survival of perennial weeds (cf. Eugen 1996). The development of mechanical seed cleaning around 100 years ago also led to a progressive decline in the segetal species specialised on dispersal together with crop seed.

12.2  Environmental Conditions and Habitat Classification

783

Further profound changes in the species composition of segetal communities resulted from the development of new herbicides and harvesting technology. As cereals are now left to ripen for longer before harvesting, some weed species now have time to seed more prolifically than before. In addition, small grass seeds such as those of Apera spica-venti remain on the field. The once very rare Avena fatua (wild oat), for example, has now become a troublesome weed, and has spread to areas in which it was previously unknown. In general, however, the conditions for arable weed species have become tougher. Their communities are becoming increasingly impoverished in both species and individuals, and the segetal communities are considered today as the most threatened element of the Central European flora (see Sect. 12.8.1).

12.2  Environmental Conditions and Habitat Classification Numerous environmental factors shape the species composition of the segetal communities in conjunction with management factors, either directly or indirectly via competitive pressure. The soil moisture, pH and base-richness, as well as the nitrogen and potassium supply and other chemical factors are reflected to various degrees in the species composition. Climatic factors, such as summer temperatures and light intensity at the soil surface, also play a role (Seifert et al. 2015a).

12.2.1  Microclimate Similarly to the case in broadleaved forests, the light intensity under the crop canopy declines rapidly during early summer. In the intensively managed crop fields of today, the minimum relative light intensity at the time of peak crop biomass regularly falls below 8–12 % of incident light, with maize stands being especially dark (around 6 %; Fig. 12.2). Thus, light availability is no higher than in a closed oak forest (see Sect. 4.3.1, Vol. I), which is less than most arable weeds require (Seifert et al. 2014). Higher light intensities in the range of 10–20 % now only exist in field margins or in fields managed for conservation purposes (see Fig.  12.1). At the beginning of the twentieth century, in contrast, these light levels were found in most arable fields. The importance of the competition for light in arable weed communities is shown by the fact that the coverage of crop plants and of weeds often follow opposite trajectories over the course of the growing season (van Elsen 1994). An early and dense crop coverage also delays the warming of the soil in spring and influences the arable weed communities via differences in temperature requirements for germination and early development (Lauer 1953; Günter 1997).

784

12  Vegetation of Arable Fields, Gardens and Vineyards

Fig. 12.2  Seasonal changes in the relative light intensity at ground level in different crop types. Modified from Rademacher (1939). In the intensively managed crop fields of today, the minimum relative light intensity at the time of peak crop biomass regularly falls below 8–12 % of incident light, with maize stands being especially dark (around 6 %; see text). w winter

12.2.2  Soil Moisture Regime The water supply continues to be an important factor in Central European arable habitats, although large-scale drainage in valleys and marshes and irrigation in drier areas have led to a widespread homogenization of the soil moisture regime in arable fields. Water availability depends to a large degree on the soil particle size, whereby e.g. sandy soils have a lower water storage capacity than loam or clay soils. In the Pleistocene lowlands of Mecklenburg-Vorpommern, Manthey (2003) determined a wide range of maximum soil water storage capacities (plant-available water) in the root zones of arable weeds, ranging from 50 to over 500 mm (see Fig. 12.3). Summer drought stress is doubtless an important growth limiting factor for arable plant communities in the more continental regions of Central Europe as well as on sandy soils (see Sect. 12.4.3).

12.2.3  Soil Acidity and Nutrient Supply Central European arable soils can range from highly acidic sandy soils (pH [KCl] < 3.5) to limestone or dolomite soils with a pH of over 7.5. This variety in soil acidity is increased by seasonal fluctuations in pH of up to 2 units, especially in the weakly buffered sandy soils (Ellenberg 1950; Zoldan 1981). This used to correspond to a wide spectrum of availability of N, K and other basic cations, before NPK fertilisation and liming became widespread. Manthey (2003) gives an impression of the variation in arable soil chemical conditions still found in northeastern Germany (see Fig. 12.4), although these differences were certainly much larger 50 years ago. In contrast, the recent survey of soil chemical properties carried out by Seifert et al. (2015b) in two intensively managed agricultural landscapes of central

12.2  Environmental Conditions and Habitat Classification

785

Fig. 12.3  Plant available soil water in 17 different arable weed communities in Mecklenburg-­ Vorpommern (197 stands, median, 25- and 75-percentiles, minima and maxima, N = number of replicates) calculated from the effective rooting depth, the soil type and humus content (the numbers on the x axis refer to the communities; see Fig. 12.4) (Modified from Manthey 2003)

Germany revealed only weak differentiation with respect to pH, C:N ratio and base saturation across a broad range of sites. Field margins may in some cases contain less plant-available P and K than the rest of the field (van Elsen 1994), although Seifert et al. (2015b) found no differences in nutrient concentrations between field interior and margin in 240 central German fields. The topsoil organic C content of central European arable fields typically varies between 0.5 and 2.5 % and that of total N between 0.07 and 0.20 % (Scheffer/ Schachtschabel 2010), which is much less than in forest and grassland soils. For cereal fields in central Germany, Seifert et al. (2015b) report a mean SOC content of 2.0–2.2 %. The stores of total N range between 3000 and 9000 kg N ha−1 in the ploughed Ap horizon (Scheffer/Schachtschabel 2010) and are particularly low in acidic soils: Zoldan (1981) recorded 6600–6800 kg N ha−1 in calcareous and loess fields (at 0–25 cm soil depth), but only 4200–4700 kg N ha−1 in sandy fields. The concentration of dissolved and adsorbed mineral nitrogen (NO3− and NH4+) in arable soils varies greatly depending on plant development, fertilisation and weather. In intensively fertilised fields, the concentrations of nitrate and ammonium in the soil liquid phase are particularly high in spring before germination, especially in root crop fields. Scheffer/Schachtschabel (2010) give typical mineral N contents for the rooted profile in cereal and sugar beet fields of 60–80  kg N ha−1 and for undersown legumes of 100 kg N ha−1. Zoldan (1981) found much higher mineral N contents (150–480 kg N ha−1) in the top 25 cm of soil of central German root crop fields.

786

12  Vegetation of Arable Fields, Gardens and Vineyards

Fig. 12.4  pH, CAL extracted potassium and phosphate, and total N in the topsoil of 17 different arable weed communities in Mecklenburg-Vorpommern (197 stands, 25- and 75-percentiles, minima and maxima, N = number of replicates) (Modified from Manthey 2003)

12.3  Vegetation 12.3.1  Classification of Arable Communities The phytosociological classification of arable, vineyard and garden weed communities has changed continuously over the last 50 years, and remains inconsistent within Central Europe. A particularly controversial question in this context is the extent to which the species composition of the segetal vegetation is influenced by management. Passarge (1964a) and Oberdorfer (1993a) considered management to play such a large role that they separated the communities of cereal fields (class Secalietea) from those of the root crops (class Chenopodietea) at the highest level, combining the latter with the ruderal communities. They thus considered the

12.3 Vegetation

787

Table 12.1  Overview of the classification of arable weed communities at the order and alliance level (following Rennwald 2000) Class Orders

Stellarietea mediae Secalietalia

Aperetalia spicae-venti

Acidic/basic soil Alliances

Rich in lime and/or bases

Acidic

Caucalidion platycarpi Cereal (winter crop) fields

Aphanion arvensis Cereal (winter crop) fields

Main occurrence

Veronico-­ Euphorbion Root crop (summer crop) fields

Spergulo-­ Oxalidion Root crop (summer crop) fields

Lolio remoti-­ Linetalia No preference Panico-­ Setarion Summer-­ warm root crop (summer crop) fields

arable weeds e.g. of a turnip field to be more closely linked to the annual ruderal vegetation of dry soils discussed in Chap. 11 (class Sisymbrietea, following the classification of Rennwald 2000) than to the arable weeds that will develop in the same field in the following year under winter wheat. The similarities with the ruderal communities are mainly related to the fact that many nitrophilic species grow well under root crops and in gardens, because these are (or were) generally more intensively fertilised than cereal fields. The majority of Central European phytosociologists now place all arable weed communities in a single class, the Stellarietea mediae (Tüxen 1950; see Chap. 14: no. 3.4), although there are differences of opinion as to whether to separate cereal and root crops at order or alliance level. Here, we will mainly follow the classification of Hüppe and Hofmeister (1990) and the terminology of Rennwald (2000) developed primarily for Germany. Based on an analysis of over 10,000 relevés from various regions of Germany, Hüppe & Hofmeister came to the conclusion that segetal communities are largely similar in cereal and root crop fields provided they have the same soil type, and that edaphic factors have a greater influence on the species composition than the type of management. Hence, two main orders can be distinguished based on the base richness of the soil, the Aperetalia spicae-venti on base-poor soils, and the Secalietalia (previously Papaveretalia rhoeadis) on base-rich soils. It is only at the level of the alliance that the cultivation system plays a role, e.g. in distinguishing between winter and summer crops, or cereal and root crops (see Table 12.1). Schubert and Mahn (1968) attempted to resolve these issues in their revised classification of the arable communities of Central Europe. In their analysis of 6300 relevés, they differentiated the vegetation units first according to their floristic similarity, and then according to the habitat types and the region. They identified eight associations based mainly on soil properties, and particularly base richness. They also included management measures such as crop rotation, herbicide use and tillage systems, but only at the level of variants of the eight associations.

788

12  Vegetation of Arable Fields, Gardens and Vineyards

The separation of cereal and root crop communities at a higher classification level proposed by Oberdorfer (1993a) may be justified in the Mediterranean region with different agricultural systems. However, it is unhelpful in Central Europe mainly because many of the character species of root crop or cereal weed communities at order or class level have a low level of fidelity. For example, the character species listed by Oberdorfer (1993a) for his root crop class (Chenopodietea) include Capsella bursa-pastoris, Chenopodium album, Tripleurospermum inodorum, Senecio vulgaris and Sonchus oleraceus, which all also occur in winter cereal crops. There are therefore very few class character species with high fidelity. Even some character species in his root crop order of the Polygono-Chenopodietalia are also found in winter cereals, e.g. Galinsoga ciliata, Anagallis arvensis, Erodium cicutarium, Polygonum persicaria and P. lapathifolium, Lamium amplexicaule, Sonchus arvensis and Stellaria media. There are similar overlaps of alliance character species between root crop and cereal communities. In addition to the character species, there are numerous accompanying species that are found in most arable fields regardless of management, e.g. Elymus repens, Galium aparine, Fallopia convolvulus, Poa annua, Tussilago farfara and Vicia hirsuta. There are therefore many links between root crop and cereal communities, which increase towards the north of Central Europe. This is certainly at least partly to do with the fact that these communities developed only within the last few centuries (Willerding 1986a).

12.3.2  Synsystematic Overview Various surveys of arable weed vegetation have been published from different regions of Central Europe, the majority of which were carried out in the 1960s and 1970s. These include: Borowiec et al. (1987), Burrichter (1963), Brun-Hool (1963), Chytrý (2009), Elias (1986), Hilbig (1962, 1966, 1967a, 1973), Hilbig et al. (1962), Hofmeister (1970), Holzner (1970, 1973), Hüppe (1987), Hüppe and Hofmeister (1990), Kläge (1999), Kropac (1981), Kump (1971), Kutschera (1966), Mahn and Schubert (1962), Májeková and Zaliberová (2014), Manthey (2003), Meisel (1966a, b, 1967, 1969a, b), Meisel and Hübschmann (1976), Mucina et al. (1993), Nezadal (1972, 1975), Oberdorfer (1992–1998), Passarge and Jurko (1975), Plakolm (1989), Preising et  al. (1995), Ries (1992), Schiller (2000), Schubert and Köhler (1964), Schubert and Mahn (1968), Tillich (1969), Waldis (1987), Wedeck (1972) & Wiedenroth and Mörchen (1964). The descriptions in the following sections are intentionally based on the situation in Central Europe around 40 years ago, i.e. before agricultural intensification wiped out much of the diversity that existed at the weed species and community level. Classification systems based on the impoverished weed vegetation today would certainly be very different to the one presented below. They would likely contain only few different types of assemblages that are best identified by a small number of dominant taxa rather than characteristic species, and further rapid changes in species composition are to be expected. Maintaining the historical classification

12.3 Vegetation

789

s­ ystems may have the advantage that it will in many cases be possible to assign even highly impoverished weed assemblages to the vegetation units described in these systems, for example by adopting the concept of ‘fragmentary communities’ in the sense of Brun-Hool (1963). 12.3.2.1  A  rable Weed Communities on Acid Soils (Order Aperetalia spicae-venti) Following the classification of Rennwald (2000) for Germany, the weed communities of cereal fields (or, more generally, winter-sown crops) on acid soils in the suboceanic regions of Central Europe are assigned to the Aphanion arvensis. This alliance includes the Sclerantho-Arnoseridetum minimae, once widespread in rye fields on acid sandy soil, which is characterised by the small Asteraceae Arnoseris minima and by Aphanes inexpectata, Anothoxanthum aristatum and Teesdalia nudicaulis. This community was abundant in the northwest German regions of ‘permanent’ rye cultivation since the early Middle Ages (Behre 1993). The Papaveretum argemone occurs in winter cereal fields on somewhat base-richer loamy or sandy soils mainly in northeastern Germany and northern Poland, and is coloured red and blue in early summer by poppies (Papaver argemone and P. dubium) and cornflower. The most abundant weed community of loamy fields in both the lowlands and uplands is the Aphano-Matricarietum chamomillae, which is promoted by fertilisation and often dominated by Matricaria recutita and the spreading Tripleurospermum inodorum. In more continental climates, the Matricarietum is replaced by the Papaveretum argemones, in montane zones by the Holco-­ Galeopsidetum, and on very poor sands by the Sclerantho-Arnoseridetum minimae (Hüppe and Hofmeister 1990; Preising et al. 1995). The winter cereal weed communities of the Aphanion alliance have a corresponding community under summer or root crops on acid soils, namely the Spergulo-Oxalidion (roughly equivalent to the Polygono-Chenopodion described by Oberdorfer 1993a). Characteristic species of this community include the North American neophyte Oxalis stricta (syn. O. fontana) as well as Chenopodium polyspermum, Cerastium glomeratum and in some areas Galeopsis speciosa. This alliance contains the Chenopodio-Oxalidetum fontanae and the Galeopsietum speciosae occurring on mesic, lime-poor loamy soils. Root crop and summer cereal fields on base-poor soils in the warmer lowlands of Central Europe support the alliance Panico-Setarion, containing many introduced grass species. The C4 grass Digitaria ischaemum (smooth crabgrass) is characteristic for the Digitarietum ischaemi, whilst Echinochloa crus-galli (cockspur) and Setaria spp. (bristle grasses) characterise the Setario-Galinsogetum parviflorae and Glebionis segetum the Spergulo-Chrysanthemetum segetum, which occur mainly in root crop and maize fields on sandy soils. The panicoid grasses are promoted by herbicide application at the expense of forbs such as Chenopodium species (Dierßen 1996a). Further Central European weed communities are listed, for example, by Kropac (1981), Oberdorfer (1993b), Pott (1995), Mucina et  al.

790

12  Vegetation of Arable Fields, Gardens and Vineyards

(1993) and Chytrý (2009); the latter authors also cover weed communities under the influence of a more continental Pontic climate and those growing on saline soils in Austria, Czech Republic and Slovakia. 12.3.2.2  A  rable Weed Communities on Base-Rich Soils (Order Secalietalia) Arable, vineyard and garden weed communities on calcareous or relatively base-­ rich loam and clay soils are assigned to the order Secalietalia (‘poppy fields’, formerly also termed the Papaveretalia rhoedis or Centauretalia cyani). Within this order, the communities of the alliance Caucalidion platycarpi contain many thermophilic character species of Mediterranean or continental origin, and colonise cereal fields on relatively dry, calcareous soils. They are distinguished from the cereal weed communities on acid soils (order Aphanion) by the lack of several indicators of acid conditions, but mainly by the occurrence of numerous calcicolous character species of the Secalietalia order or the Caucalidion alliance: Still relatively abundant: Anagallis foemina Consolida regalis Euphorbia exigua Rare: Adonis aestivalis Bifora radians Caucalis platycarpos Lathyrus tuberosus Legousia hybrida Legousia speculum-veneris Melampyrum arvense Silene noctiflora Scandix pecten-veneris Almost extinct (+: lost in large regions of central Europe): Adonis flammea (+) Asperula arvensis (+) Bupleurum rotundifolium Conringia orentalis Turgenia latifolia (+) Vaccaria hispanica (+) Some of these ‘lime indicators’ also occur on acid soils, such as Euphorbia exigua.

12.3 Vegetation

791

Many of the previously used character and differential (D) species of the Caucalidion are now rare, e.g. Sherardia arvensis, Legousia speculum-veneris, Valerianella dentata (D), Campanula rapunculoides (D), Buglossoides arvensis, Chaenarrhinum minus (D), Ranunculus arvensis, Falcaria vulgaris (D), Stachys annua, Galeopsis angustifolia (D), G. ladanum (D), Galium tricornutum and Melampyrum arvense. The once highly species-rich Caucalido-Adonidetum flammeae (syn. Caucalido-Scandicetum pecten-veneris) colonises shallow, calcareous stony soils, whilst the Euphorbio exiguae-Melandrietum noctiflori occurs on moderately decalcified loamy fields mainly in subcontinental Central Europe. The Kickxietum spuriae is found on calcareous, occasionally waterlogged fields in more subatlantic climates (Preising et al. 1995). In areas of more intensive fertilisation and on base-poor soils, the Melandrietum is replaced by the Aphano-­ Matricarietum chamomillae (see above). Root crop and summer crop fields on base-rich loamy and clay soils support communities of the Veronico-Euphorbion alliance. These are characterised by nutrient-demanding herbs such as Euphorbia helioscopia, E. peplus and Fumaria officinalis. The Thlaspio-Fumarietum officinalis is one of the most common associations in this alliance, and is found mainly in turnip fields. The Mercurialietum annuae occurs in fertile and frequently worked gardens, fields and vineyards. 12.3.2.3  Weed Communities of Vineyards and Other Cultivated Land Central European vineyards differ in several respects from crop fields and gardens as habitats of herbaceous weeds (e.g. Issler 1942; Link 1954; Roser 1962; Hilbig 1967b). This is related to the manner of growth and cultivation of the vines (Vitis vinifera), the wild form of which naturally occurs in Mediterranean floodplains. The following habitat characteristics play a major role in shaping the weed vegetation (Wilmanns and Bogenrieder 1992): –– Trained vines develop leaves high above the soil surface, so that it is possible to work the soil directly beneath the vine and around the stem throughout the year. –– Vine leaves develop relatively late and shade the soil only from the end of May till the end of October. As the winters in grape-growing areas are mild, the herb layer experiences a relatively long growing period from November until May in which disturbance is rare. –– As individual vines can live for several decades, and vineyards are often cultivated continuously for centuries, even weeds with low reproductive rates and dispersal abilities can colonise and persist, including many geophytes. Wilmanns and Bogenrieder (1992) studied the weed community composition in southwest German vineyards with either traditional or modern cultivation (see Table  12.2). In the former, the slopes were fertilised with manure and harrowed, tilled or ploughed to combat weeds. However, even in the areas with traditional management, only few species of the once rich assemblages of bulb and tuber plants remained, among them Allium vineale and Ornithogalum nutans and O.

12  Vegetation of Arable Fields, Gardens and Vineyards

792

Table 12.2  The Geranio-Allietum of vineyards in Southwestern Germany that were managed in a traditional way, with intensive herbicide application or by maintaining dense grass cover on the soil. Excerpt from Wilmanns and Bogenrieder (1992)a Relevé no. Region Aspect Slope angle (°) Year of survey (19..) Geophytes Allium vineale Ornithogalum umbellatum Muscari racemosum Ornithogalum nutans Tulipa sylvestris Corydalis solida C. cava Ranunculus ficaria Indicators of bare soil Stellaria media Fumaria officinalis Euphorbia helioscopia Mercurialis annua Lamium amplexicaule Herbicide resistant plants Galium aparine Polygonum aviculare Valerianella carinata Bromus sterilis Cardamine hirsuta Poa trivialis group Poa trivialis Lolium perenne Agrostis stolonifera Trifolium repens Glechoma hederacea Ranunculus repens R. acris Potentilla reptans Festuca arundinacea Stellarietea sensu lato Veronica persica V. hederifolia

Traditional 1 2 M A SW S 15 12 86 83

Herbicides 3 4 W I SW S 6 20 83 90

+ 2 (+) 2

1 1

1

Grass cover 5 6 A A NE W 5 3 86 83

2 1 2

2 1 2 + 1 2 2

4 1 + 2

+ +

+

+ + 1

+ 1 + 4

1 +

2

2

+

2

1

4 + 1 + 2 3 1

5 1 + + +

1 1 1 +

2 +

1 3

2 +

+ +

2

Life form and Indicator value Lf.

T

M N

G G G G G G G G

7 6 7 7 7 6 6 5

4 5 3 4 4 5 6 6

7 7 6 7 5 7 8 7

T T T T T

× 6 × 7 6

× 5 5 4 4

8 7 7 8 7

T T T T TH

6 6 7 6 6

× 4 4 4 5

8 6 × 5 7

HC H H CH GH H H H H

× 6 × × 6 × × 6 5

7 5 7 5 6 7 6 6 7

7 7 5 6 7 7 × 5 5

T T

× 6

5 5

7 7

(continued)

12.3 Vegetation

793

Table 12.2 (continued) Relevé no. Region Aspect Slope angle (°) Year of survey (19..) Convolvulus arvensis Senecio vulgaris Lamium purpureum Sonchus oleraceus Malva neglecta Capsella bursa-pastoris Lathyrus aphaca Tripleurospermum perforatum Other species Taraxacum officinale Elymus repens Poa annua Cirsium arvense Vicia angustifolia Urtica dioica Crepis capillaris Cardaria draba

Traditional 1 2 M A SW S 15 12 86 83 1 + 1 + 2 2 1 1 1 + + +

Herbicides 3 4 W I SW S 6 20 83 90 + 1 2 + + + + +

1 1

1

1 + +

2 1 2 1

Grass cover 5 6 A A NE W 5 3 86 83 + + + +

1

1 2

+ 2

2 2

+ 2 1

Life form and Indicator value Lf. GH TH TH TH TH T T T

T 6 × 5 6 6 × 7 6

M 4 5 5 4 5 5 3 ×

N × 8 7 8 9 6 3 6

H G TH G T T H HG

× 6 × 5 6 6 6 7

5 × 6 × × 5 5 3

8 7 8 7 × 8 4 4

Plots 1 and 2: mainly mechanical cultivation and organic fertiliser, i.e. traditional management; plots 3 and 4: herbicide treatment several times per year; plots 5 and 6: with a grass layer that is mown and mulched several times a year. All plots recorded in the Markgräflerland (southwestern Germany), in the districts of A=Auggen, M=Mauchen (SSW of Müllheim), I=Istein (NW of Lörrach), W=Weil am Rhein (SW of Lörrach), T=temperature, M=moisture, N=nitrogen IV from Ellenberg et  al. (1992). Life forms: T=therophyte, G=geophyte, H=hemicryptophyte, C=chamaephyte

a

umbellatum. Geranium rotundifolium, after which the vineyard community of the Geranio-Allietum vinealis is named, had already been absent for a long time, and the formerly almost ubiquitous Muscari racemosum (blue grape hyacinth) was rare. Plot 1 in Table 12.2 had been replanted recently and thus had a relatively species-­ poor weed community. Plot 2 was in an older traditional vineyard with almost all humus-rich and fertile soil being covered by Stellaria media, which fruits and germinates several times a year. Vineyards with herbicide treatment had a lower abundance of therophytes but not of the geophytes. Bromus sterilis was particularly promoted by this management and Wilmanns and Bogenrieder (1992) termed this variant of the Geranio-Allietum a ‘Bromus sterilis-agroform’, in contrast to the Stellaria-rich ‘Stellaria media-agroform’. In the last 30 years, Central European vineyards have increasingly used sown or spontaneously growing grass cover to avoid the need to remove weeds, which also reduces soil erosion. The grass is occasionally mulched to prevent it from growing

794

12  Vegetation of Arable Fields, Gardens and Vineyards

too high and shading the vines. Vineyard geophytes are unable to penetrate this dense grass sward, and the Geranio-Allietum is replaced by a ‘Poa trivialis-Lolio-­ Potentillion community’. The species composition of vineyard weed communities varies somewhat with geology, climate and cultivation methods. However, the rapid shift since the mid twentieth century from traditional to intensive management with herbicide application, and finally to grass cover management, has had a similar effect in all central European vine growing regions. Almost all species of vineyards are indicators of high nutrient availability (EIV N 6–9). In addition, many taxa are relatively thermophilic (EIV T 6 or 7) or have a broad temperature tolerance. However, highly thermophilic species do not occur in Central European vineyards, probably because of the shade cast by the vines. Many weeds are also drought-tolerant (EIV M 3–4 or indifferent). Hopfields are typically managed as permanent cultures and are replanted every 15 years. Like vineyards, they are rarely colonised by typical weed species of cereal fields, as shown by Brandes (1988) in the Hallertau region in Bavaria. The soil in hopfields is often covered throughout the year by Stellaria media, which prevents erosion in sloping areas in the same way as the grass cover in vineyards. The once widespread weed communities of flax fields formed their own order (the Lolio remoti-Linetalia). Several weed species that were exclusively dispersed with flax seed (e.g. Camelina alyssum, Cuscuta epilinum, Lolium remotum, Silene linicola, Spergula arvensis subsp. linicola) are now practically extinct in Central Europe with the decline of flax cultivation (Schneider et  al. 1994). Flax fields are still found in some regions today, but contain the typical weed species of root crop fields.

12.4  Adaptations to the Environment 12.4.1  Arable Plant Functional Types In order to persist as an arable or garden weed, i.e. as an undesirable plant from the land manager’s perspective, it must possess the following attributes: 1. If it is a short-lived therophyte, it must develop rapidly from germination to fruiting. 2. If it is not able to germinate or grow at any time of year, its developmental rhythm must fit in with the cultivation rhythms of the field or the garden as well as with the climatic rhythms of the region. 3. It must be able to tolerate periods of shade during the summer or, like Convolvulus, Vicia or Fallopia, to escape it by climbing (see Fig. 12.5). 4. Whether it is annual or perennial, it must be able to regenerate quickly following mechanical damage or if it is covered by earth after ploughing.

12.4  Adaptations to the Environment

795

Fig. 12.5  The effects of different light intensity on the dry matter production of four arable weed species in monoculture and in mixed culture with 1, 2 or 3 other species (Modified from Bornkamm (1961). All species grow best in monoculture at 100 % incident light, and worst at the lowest light level) The tall Agrostemma githago (corncockle) is a superior competitor to the other species and maintains roughly the same productivity as in monoculture. The other behavioural extreme is Anagallis arvensis (scarlet pimpernel), which is suppressed under full light conditions by all the other species, and particularly by the combination of all three. It therefore grows better at lower light levels, as its competitors are hindered in their development (i.e. it tolerates shade better than the other species tested here, but is not directly promoted by it). Bromus secalinus (rye brome) and Sinapis alba (white mustard) have an intermediate strategy

796

12  Vegetation of Arable Fields, Gardens and Vineyards

5. Today, the most important adaptation for a weed is a certain degree of herbicide resistance or a way of circumventing herbicide effects by phenological plasticity, e.g. by late germination or shoot formation. Therophytes, geophytes and hemicryptophytes occur at varying proportions in weed communities depending on crop rotation and management intensity. Perennial species are more frequent in perennial crops such as clover and lucerne, whilst annual species dominate intensively managed root crops and garden habitats. Ellenberg (1950) and Koch (1970) distinguished the following developmental types of segetal plants: A. Annuals (a) Winter annuals that germinate in autumn and survive the winter in a vegetative state, finishing their development in spring or summer (e.g. Apera spica-venti, Veronica hederifolia, Viola arvensis, Centaurea cyanus); these species can also germinate in spring. (b) Species that remain green over winter, which can germinate, flower and set seed throughout the year (e.g. Stellaria media, Lamium purpureum, Veronica persica, Poa annua, Capsella bursa-pastoris, Thlaspi arvense). (c) Summer annuals, which only rarely germinate in autumn because they do not survive frosts. These include: c1 spring annuals, which develop early in spring and have mostly died back by the summer (e.g. Veronica triphyllos, Ornithogalum umbellatum, Tulipa sylvestris, Myosurus minimus); c2 species that germinate under low temperatures and start to photosynthesise in late winter, sometimes until the beginning of the next winter (e.g. Sinapis arvensis, Raphanus raphanistrum, Galeopsis tetrahit, Avena fatua, Polygonum persicaria, Vicia hirsuta); c3 summer-green plants that germinate late and only under warm conditions (e.g. Echinochloa crus-galli, Setaria, Digitaria and Galinsoga species, Solanum nigrum, Amaranthus retroflexus, Mercurialis annua). Schneider et  al. (1994) further subdivide the annuals into eleven developmental types based on their germination date. B. Perennials Perennial herbs and grasses (e.g. Elymus repens, Cirsium arvense, Sonchus arvensis, Equisetum arvense and the late-shooting geophyte Fallopia convolvulus). A few of these species are parasitic (e.g. Melampyrum arvense). The vast majority of the arable weeds still able to survive in intensively managed arable fields are therophytes, but even these are often destroyed by harrowing, tilling or herbicide application before they can produce ripe seeds. It is therefore an advantage for these species if their seeds do not all germinate at the same time, and if the dormant seeds remain viable for as long as possible. The arable weeds today are therefore dominated by species with variable dormancy lengths.

12.4  Adaptations to the Environment

797

In contrast, some tall weed species that used to be harvested together with the cereal or flax crops have been unwittingly bred by humans to follow the opposite germination and fruiting strategy. These include e.g. Lolium temulentum, Bromus secalinus and B. grossus, Agrostemma githago and Avena fatua. Thellung (1925) and other authors consider such ‘crop plant attributes’ to be: 1. short seed dormancy, i.e. rapid germination of all viable seeds after sowing (as is the case e.g. for Agrostemma githago and Bromus secalinus); 2. loss of germinability after less than 1 year even with storage in dry conditions (also the case for the species listed in 1); 3. lack of natural dispersal features (e.g. the fruits of the flax weed Camelina alyssum only open under pressure, the seeds of Agrostemma do not fall easily out of their capsule, and the fruits of Avena strigosa do not spontaneously drop off the plant); 4. large seeds or other similarly shaped propagules that are difficult to separate from those of the crop plant (e.g. the fruits of Camelina alyssum in linseed, or the pods of Raphanus raphanistrum in cereals). Agrostemma githago, Bromus secalinus and other species in which the germinability of their seeds declines after only a few months in the soil, depend on being sown each year with the crop seed. Since the drastic improvements in seed cleaning technology in the last 50 years, these once ubiquitous cereal weeds have almost disappeared from Central Europe. In addition to such seed mimics in the stricter sense, seed cleaning has also largely caused the declines of species such as Centaurea cyanus, although its seeds remain viable for longer. Cornflower, which was once considered the most permanent of arable weeds, became a botanical rarity at the end of the twentieth century due to the use of broadband herbicides. It was only after the introduction of more selective herbicides in cereals and oil seed rape that Centaurea cyanus has returned to some areas. Biennial and perennial species are only able to survive on arable fields when they are capable of rapid vegetative reproduction. Species that are well-adapted in this respect are e.g. Cirsium arvense, Equisetum arvense and Tussilago farfara. Generally, plants with high relative growth rates and high seed production (R strategists) are much more common among the arable weeds than the slow-growing k strategists that produce low numbers of seeds. Arable weed R strategists include Chenopodium album and Tripleurospermum inodorum, whilst k strategists are Lithospermum arvense or Scandix pecten-veneris, which produce only a few hundred seeds per plant (Schneider et al. 1994; Bischoff 1996).

798

12  Vegetation of Arable Fields, Gardens and Vineyards

12.4.2  G  ermination Conditions and Light and Temperature Requirements The seeds of most arable weeds do not germinate immediately, but must first undergo a primary dormancy that, depending on the species, can last for 2–3 weeks or up to 5 months (Börner 1995). Most segetal plants germinate only in the uppermost 3–7  cm of topsoil once the light, moisture and oxygen levels have reached the levels required by the species. However, the most influential factor is the temperature. Most root crop weeds require high temperatures to break their dormancy, and therefore cannot germinate in winter (Salzmann 1939; see Sect. 12.4.1). In the species studied by Lauer (1953), the germination optimum is above 20 °C and the minimum temperature for germination is 15 °C (see Fig. 12.6). However, not all the character species of root crop communities need warmth to germinate, and in fact the most common of these, such as Chenopodium album and Stellaria media (see Fig. 12.6), do not follow this pattern. Such species with a wide amplitude of germination temperature are also regularly found in cereal fields, but do not grow as well there as they do in root crops. These are plants with a high nutrient demand, and root crops have, at least historically, generally received more fertiliser than winter cereals. The cereal weeds studied by Lauer (1953), in contrast, mainly had their optimum germination at low temperatures (< 10 °C). Their germination is stimulated by the falling temperatures in autumn. If these and the indifferent species have already colonised the field, then the warm-germinating species may not be able to develop due to lack of light even once the soil temperatures have risen high enough for them. It is then only after harvest that enough light reaches the ground for them to grow. If the final soil cultivation of the year takes place in late summer, then the warm-germinating species have a head start (Koch 1970). As they generally develop very quickly, they easily gain the upper hand (Lauer 1953). Not only the germination, but also the growth of the young plants is highly dependent on the temperature. For example, Bupleurum rotundifolium is only competitive if the young plants can achieve high growth rates in a warm spring (Günter 1997). Although the warm-germinating root crop weeds are not able to persist at higher elevations, the temperature-indifferent weeds are particularly frequent in the montane belt. Areas where crops are not cultivated due to long-lasting snow cover or other reasons, lack many typical crop weeds (see e.g. Waldis 1987). Communities of the root crop alliances Veronico-Euphorbion and Spergulo-Oxalidion once reached up to the upper limit of crop cultivation in the Alps (> 2000  m a.s.l.; Hügin 1995), although they contained fewer species. Communities of the cereal weed alliances Caucalidion platycarpi and Aphanion arvensis, in contrast, were only found up to around 1200  m a.s.l. Today, crop cultivation is no longer economically viable at these elevations and has mostly disappeared. A comparison of the average indicator values of the segetal species of calcareous fields with those of deep loam fields gives the impression that the soil influences the temperature requirements of the species. The frequently dry calcareous soils indeed

Fig. 12.6  The germination rates of several segetal species mainly occurring in summer crops and gardens (top) and in winter cereals (bottom) at constant temperature levels (From data from Lauer in Ellenberg (1963)). The curves show the percentage of all seeds germinating at each temperature level. Warm-germinating seeds generally germinate between 20 and 30 °C (dark grey) or even over 30 °C (black). Cold-germinating seeds require temperatures below 20 °C (medium grey), and in some cases even under 7 °C (light grey). The species with a flatter curve (e.g. Raphanus and Capsella) are largely indifferent to temperature

800

12  Vegetation of Arable Fields, Gardens and Vineyards

warm up faster than the more mesic loamy soils with generally higher water content. However, the higher average temperature indicator values of the calcareous arable communities are probably the result of the shallow soils in this habitat and the related low coverage of the crop plants (Ellenberg 1950). All of the arable weed species tested by Bornkamm (1961) grow best in full sunlight, and are therefore sensitive to some degree to shading by the crop. Light availability was the most important determinant of growth for five endangered arable weed species across a range of fertiliser levels (Kleijn and van der Voort 1997). The least light-demanding arable weeds are the nitrophytic herbs of the dense cereal crops on loamy soils. The optimum and tolerated values for light intensity, but also the pH and other factors, show that many arable weeds are forced into suboptimal habitats by superior competitors (see Fig. 12.6). In mixed cultures, tall species such as Centaurea cyanus and Sinapis suppress the low-growing ones, so that these must persist in relatively dark conditions.

12.4.3  Adaptations to Moisture and Aeration of the Soil As very wet or very dry soils are unsuitable for arable farming, most segetal communities grow under the mesic conditions that promote most plant species. Communities adapted to drier or wetter conditions only exist as subassociations, variants or subvariants, and their species composition is mainly determined by other factors. Nevertheless, segetal communities can be used in the same way as grassland or forest communities to divide the landscape into areas of different moisture levels (Meisel and Wattendorf 1962; Manthey 2003). Arable weed communities on acid soils can be found from Portugal to Estonia, and show the gradient of continentality and water supply just as clearly as the much less anthropogenic forest communities (Malato-Belitz et al. 1960). Adaptations to Waterlogging  Periodically waterlogged and hypoxic soils increase the competitiveness of species that are otherwise rare in crop fields and gardens. Three groups of these species can be distinguished based on their behaviour and growth forms, and they often occur as differential species of subunits, either in combinations or alone: 1. deep-rooting perennial species that usually indicate temporary waterlogging of the subsoil: Equisetum arvense, Polygonum amphibium var. terrestre, Tussilago farfara; 2. (a) shallow-rooting perennial species adapted to waterlogging in the topsoil: Agrostis stolonifera, Potentilla anserina, Equisetum sylvaticum, Ranunculus repens, Mentha arvensis, Rorippa sylvestris, Poa trivialis, Stachys palustris; (b) species with response patterns similar to 2a, but less restricted to waterlogged soils: e.g. Aphanes arvensis, Poa annua, Alopecurus myosuroides, Sonchus arvensis and S. asper, Apera spica-venti, Ranunculus arvensis, Matricaria recutita, Tripleurospermum inodorum;

12.4  Adaptations to the Environment

801

Fig. 12.7  Germination, seedling growth and adult vegetative growth are influenced by soil moisture in different ways in weed species. This is the same whether the plants are indicators of high (M7 and 8), intermediate or low (M4 and 3) soil moisture. Data from Snoy (see Ellenberg and Snoy 1957), moisture values (M) from Ellenberg (1979). The pot experiments were carried out in three series. In the dry treatment (d, dotted bars), the soil moisture tension remained around 10 bar (1 MPa); in the wet treatment (w, black bars) the water level was always close to the soil surface. The moist treatment (m, white bars) was watered moderately every day, and served as a control (= 100 %). Gnaphalium uliginosum and Sagina procumbens germinate under wet conditions (the germination rate in the wet treatment is several times that in the moist treatment). However, the young seedlings start to be damaged by the water (average dry matter production per pot in w is at most 50 % of that in m). Later, the waterlogged conditions cause them to die. Polygonum hydropiper, and particularly Plantago major subsp. intermedia, Poa annua and Alopecurus myosuroides, follow the opposite trend. Chaenarrhinum minus and Falcaria vulgaris, which are mainly found on free-draining soil, only germinate under moist conditions. Older plants of Falcaria can tolerate wet conditions, but the species usually does not exploit this ability under natural conditions

3. shallow-rooted, short-lived species germinating under wet conditions that can only develop if there is sufficient moisture in the topsoil, but are not reliable indicators of waterlogging (although they are sometimes described as such in the literature), and mainly found in the communities of the Nanocyperetalia (Sect. 10.1.2): e.g. Gnaphalium uliginosum, Polygonum hydropiper, Juncus bufonius, Sagina apetala, S. procumbens, Plantago major subsp. intermedia. Ellenberg and Snoy (1957) experimentally studied the physiological behaviour during germination, seedling growth and further development till maturity of many of the species listed above, as well as some species of drier conditions. They came to the conclusion that the moisture requirements of many of the studied species change over the course of their lifetime, and in fact often reverse (see Fig. 12.7). Gnaphalium uliginosum, one of the most common species of group 3, requires wet conditions to germinate and its seedlings grow best on soils that are constantly waterlogged up to the surface. However, the older the plant becomes, the more it suffers from the lack of oxygen found in such soils. If the waterlogged conditions persist, then it forms only a very shallow root system and is more likely to be

802

12  Vegetation of Arable Fields, Gardens and Vineyards

p­ revented from fruiting by a fungal infection. Juncus bufonius and Polygonum hydropiper follow a similar pattern. These species do not grow well on constantly wet soil, and their growth is best on deep, well-aerated but always moist soils. It is therefore wrong to describe them simply as indicators of waterlogged conditions, and they should instead be considered indicators of damp topsoil, especially as these species can also be found in damp years on soils that are certainly not waterlogged. In contrast, the perennial species of groups 1 and 2a tolerate the hypoxia related to constant waterlogging, for example Ranunculus repens and Equisetum arvense. Like swamp plants, their roots or rhizomes possess large-pored aerenchyma tissue. This allows them to penetrate deeper into waterlogged and anoxic soils than most other plants. These are therefore true indicators of waterlogged conditions. Adaptations to Drought  Summer drought plays a role in many crop fields, not only in regions with pronounced summer drought such as the Upper Rhine Plain or Burgenland. Based on their indicator values for soil moisture, the majority of arable weeds can be considered as largely drought-tolerant. Several communities of the Caucalidion platycarpi, the Panico-Setarion and the neophyte-dominated Eragrostion minoris contain particularly large numbers of indicators of dry conditions. However, the other arable weed alliances also generally have both drier and wetter communities at the level of variants and subvariants (Meisel and Wattendorf 1962; Schubert and Mahn 1968). According to measurements by Kudoke and Kaussmann (1973), the topsoil of arable fields in northeastern Germany with drier variants of segetal communities has less than 10 vol%, and in some cases even under 5 % soil moisture, whilst fields with mesic and damp variants never drop below 10 %. Drought stress can be avoided through the formation of a deep root system, leading Wehsarg (1931) and Ellenberg (1950) to classify 175 arable weed species from central Germany according to their typical root depth: (a) topsoil-rooting species, the roots of which rarely go deeper than 10 cm (e.g. Juncus bufonius, Sagina procumbens, Gnaphalium uliginosum); (b) shallow-rooting species, the roots of which reach to around 20  cm (e.g. Galeopsis tetrahit, Aphanes arvensis, Lamium amplexicaule); (c) moderately deep-rooting species with roots down to around 30 cm (max. 50 cm) (e.g. Agrostemma githago, Adonis aestivalis, Erodium cicutarium); (d) deep-rooting species, the roots of which are always 30–50 cm deep, and rarely down to 1  m (e.g. Equisetum arvense, Taraxacum officinale, Artemisia vulgaris); (e) subsoil-rooting species that often reach depths of over 1  m (e.g. Cirsium arvense, Convolvulus arvensis, Falcaria vulgaris). The maximum rooting depth depends strongly on the soil type (Kutschera 1960). Of the 175 segetal species studied, Ellenberg (1950) identified 9 % as topsoil-­ rooting species, 41 % as shallow-rooting, 38 % as moderately deep-rooting, 10 % as deep-rooting and 2 % as subsoil-rooting. Almost 90 % of the species therefore ­usually rooted down to less than 30 cm and rarely down to 50 cm, making it almost

12.4  Adaptations to the Environment

803

impossible for them to reach the damper subsoil. Many arable weeds, such as the spring annuals, therefore avoid the summer drought in that they complete their life cycle early in the year. Arable weed communities of dry soils contain increasing numbers of thermophilic introduced C4 grasses (e.g. Digitaria ischaemum, Setaria species, Eragrostis species, Echinochloa crus-galli). These have spread most rapidly in maize fields where crop cover develops late in spring.

12.4.4  A  daptations to the Nutrient Regime and the Effects of Fertiliser Application The main soil properties that influence (or influenced until recently) segetal communities are soil moisture (see Sect. 12.4.3), base saturation and pH.  Table  12.3 displays several vegetation relevés from crop fields in southwestern Germany occurring along a gradient of soil acidity. Many species of the Sclerantho-Arnoseridetum minimae are calcifuge, and thus restricted to fields with acid soils, including Teesdalia nudicaulis, Hypochaeris glabra, Ornithopus perpusillus, Arnoseris minima. In contrast, many species of the Caucalidion alliance are calcicole (e.g. Caucalis platycarpos, Lathyrus tuberosus, Bupleurum rotundifolium). However, the apparent affinity to certain pH ranges is not universal, as is shown by results from Finland where some of these species behave differently (Borg 1964). In his analysis of a broad variety of segetal habitats in northeastern Germany, Manthey (2003) found a strong influence of base saturation and pH on the species composition (see Fig.  12.4), even though these areas are now also intensively limed and fertilised. Nitrogen availability also influences the species composition of the segetal communities (Pyšek and Lepš 1991; Storkey et al. 2012), although most fields are now so intensively fertilised that differences in soil N supply have diminished. On the landscape scale, N is often less important for arable weed communities than P availability, soil moisture and base saturation. All arable weeds are relatively nutrient-­ demanding, and moderate fertilisation (e.g. 40  kg N ha−1) causes them to grow better, to have a higher germination rate, and often also to produce greater quantities of seeds (Wells 1979; Mahn 1984; Günter 1997). However, fertiliser application can also delay the transition in arable weeds from vegetative to generative phases and increase mortality (Mahn 1992; Bischoff 1996). Normal levels of fertiliser today (> 150 kg N ha−1) only promote the growth of a few of the segetal species, but often seem to be beyond optimal growing conditions. Some species (e.g. Consolida regalis and Centaurea cyanus) have been found to react to these conditions with a reduction in productivity and flower formation (Svensson and Wigren 1982). As shown by the experiments of Mayser (1954), nitrogen fertilisation promotes the growth of ruderal plants and root crop weeds (e.g. Amaranthus retroflexus, A. lividus, Solanum nigrum, Setaria verticillata, S. glauca and ­

Serial no.: pH value (mean of First 1948) digit decile Pronounced acidity indicators: Rumex acetosella Scleranthus annuus Spergula arvensis Acidity indicators in a wider sense: Aphanes arvensis Raphanus raphanistrum Apera spica-venti Indifferent: Matricaria recutita Poa annua ± preferring calcareous soil: Sinapis arvensis

2 3 4 5

2

2 2 2 2

1 + + 3

2

1

1

2 +

2 + 1 + 1

1

1 1 2

1

2

1

2

+

3

2 2 2 2 1 1 2

1 +

2

+

+ 2 + 1

2

2 2

6. 1

3

3

1

2 1

+

+

4

2

1

+

4

2

2 +

2

5

1

2

+

5

2

1 1

+

7

1

1

9

2

+

7. 0

1

2

3

2

1

2

1

2

2

3

1

3

4

3 3

3

2 2

1 1 1

1950 R1-5

8

5 ×

4

4 4

2 2 2

1974 R1-9

8b

5 ×a

5

×a 4

2 2 3

1992 R1-9

Indicator value for 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 pH (EIV-R)

4. 5. 5 6 6 8 1 2 3 5 7 8

1

Table 12.3  pH amplitude of several arable weeds in southwestern Germany and Central Europe according to measurements in 25 plots

804 12  Vegetation of Arable Fields, Gardens and Vineyards

1

2 3 4 5 1 + + +

1 +

1 +

2 3 1 +

2 + +

3

1 1

1 + 1

3

+

3 2

1 1 +

1 + 3

+ 2 +

1

+ 2 3

+

1

1 1

+ 1

1

1

1

5

5 5 5

4 4 4

9

8 8 9

6 7 8

9

8 8 9

6 7 8b

Indicator value for 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 pH (EIV-R)

Modified from Ellenberg (1950) and Ellenberg et al. (1992). The pH values are averages calculated from numerous measurements taken throughout 1948 on each plot. The numbers in the columns 1 to 25 give the highest coverage of the species that was found in the 25 plots in the course of 1948. The derived indicator values for soil acidity/reaction (EIV-R) in the last three columns refer to three different categorisations from 1950 (original 5-scale system), 1974 and 1992 (9-scale system with later adjustments) (R1 = indicators of very high acidity, never found on weakly acid or alkaline soils; R9 = indicators of base richness, always found on calcareous soils) a The species should be considered indifferent, because it often also occurs on calcareous soils if they are well fertilised. The five-point scale for 1950 is based on data only from southwestern Germany. Later results from Germany and Switzerland suggested a nine-point scale to be more appropriate b Based on the data used here, this species had to be classified as EIV-R6 or R7, but generally occurs mainly on base-rich soils. The changes in the ecological indicator value over time stem from the accumulation of additional data from across Central Europe

Fumaria officinalis Papaver rhoeas Sonchus oleraceus Indicators of calcareous soil: Consolida regalis Galium tricornutum Caucalis platycarpos Anagallis foemina

Serial no.:

12.4  Adaptations to the Environment 805

806

12  Vegetation of Arable Fields, Gardens and Vineyards

Fig. 12.8  The effects of increased nitrate and ammonium fertilisation of monocultures of arable weeds and ruderal plants in pots (Modified from data from Mayser in Walter 1963). The pot experiments were carried out with a mixture of loam and sand under constant water supply (60 % of soil water capacity) and PK fertilisation, either with no additional nitrogen (N0) or with a large amount of N in the form of NO3− or NH4+. The bars show the yield increase in multiples of N0 (above-­ ground dry weight production). Apart from Urtica urens and Sisymbrium, both forms of N have similar positive effects on the yield. Ruderal plants (Arctium, Amaranthus and Sisymbrium) profit most, and garden and root crop weeds (Urtica urens, Echinochloa, Setaria) also benefit, although not all of them (e.g. Chenopodium polyspermum). Silene nutans and vulgaris, i.e. two species that occur in nutrient-poor grasslands, react least to N fertilisation. Capsella bursa-pastoris has a wide amplitude for N under natural conditions

Echinochloa crus-galli) more than weeds of cereal crops and those indifferent to management type (e.g. Ranunculus arvensis, Bupleurum rotundifolium, Valerianella dentata, Sherardia arvensis and Sonchus arvensis, see Figs. 12.8 and 12.9). Small weeds are particularly disadvantaged, as they are unable to keep pace with the fastgrowing Chenopodiaceae, Galium aparine, Stellaria media and large-leaved root crop and garden weeds (Günter 1997). These often reach high coverage values in highly fertilised fields (Menck and Behrendt 1974) such as intensive vegetable cultures (see Fig. 12.10), but flower and seed less often than when they grow under less nutrient-­rich conditions (Snaydon 1984; Franz et al. 1990). The lush growth of both crop plants and weeds caused by nitrogen application almost always leads to a decline in weed species richness (Böhnert 1979; Pulcher-Häussling 1989; Kulp 1993). Schmitzberger et al. (2005) found a linear decrease in the number of species in Austrian crop fields with an increase in N fertilisation from 50 to 150 kg ha−1. Whether fertilisation generally leads to an increase (as found by Do van Long 1978 and Alkämper et al. 1979) or to a decrease in the richness and coverage of weeds (e.g. Radics 1990; Verschwele and Niemann 1992), depends mainly on how

12.4  Adaptations to the Environment

807

Fig. 12.9  The effects of increased ammonium and nitrate fertilisation on a mixed stand of arable weeds (Modified from data from Mayser in Walter 1963). The mixed cultures were each planted on 1 m2 of garden soil with the same numbers of individuals of each species, and either not fertilised (N=0) or given different amounts of N in the form of NH4+ or NO3−. The proportion of each species in the above-ground harvested biomass is given in %. Chenopodium album is the dominant species on soil with no additional fertilisation as well as on soil fertilised with NH4+ or NO3−, and its growth response to increasing N supply is also the strongest. Chenopodium polyspermum, Solanum nigrum and Echinochloa crus-galli also increase in coverage with increasing fertilisation. The remaining species are increasingly suppressed, although they are all promoted by increased N when grown without competitors. Comparing this with Fig. 12.8 also shows that Chenopodium album reacts less strongly to N fertilisation in monoculture than Echinochloa and Solanum nigrum

successfully the weeds can compete with the crop species for light. The outcome of this depends on the weather, the crop type and the species composition of the segetal community (Schuboth and Mahn 1994; Bischoff 1996). If the crop is strongly ­promoted by fertilisation, as is the case for modern high-yielding varieties, then the

808

12  Vegetation of Arable Fields, Gardens and Vineyards

Fig. 12.10  Intensive cultivation of various root crops (mostly vegetables) and ornamental plants on the outskirts of the city of Hamburg. Note the small size of fields. A small fruit tree plantation is still managed at low intensity and increases habitat diversity

weeds will be quickly shaded out. At most, climbing species such as Fallopia convolvulus and the Vicia species or Galium aparine, will profit, as they can escape the shady soil surface. The effect of N on the vegetation is determined to a large extent by the amount of mineral N available to the plants during the growing season. In arable fields, this depends not only on the mineralisation rate, but to a large extent on the intensity of fertilisation. The amount of fertiliser applied, in turn, depends on the type of crop and this will change over time with the annual crop rotation (see Sect. 12.7.1). Changing N availability may be one of the reasons why Manthey (2003) (in contrast to the situation in forest and mire habitats) did not find a clear relationship between the C:N ratio in the topsoil and the segetal vegetation, although this ratio is usually closely linked to the N mineralisation rate. Continued intensive fertilisation has certainly also reduced any previous fertility differences in crop habitats. However, the study by Manthey (2003) found that the total N content of the soil had a clear influence on the composition of the vegetation. The C and N content of the soil is in turn closely related to the particle size and thus the geology. In the relatively poor soils of northeastern Germany, the plant-available P content of the soil apparently has only a relatively small influence on the species composition of segetal communities, as Manthey (2003) found only a weak relationship between the calcium acetate lactate (CAL)-soluble P and the vegetation of crop fields (see also Trautmann 1954). As with N, this may be due to the fact that P is now no longer a limiting factors in most Central European arable fields (apart from

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Table 12.4  Arable weeds with similar behaviour regarding the soil P content (calcium acetate lactate extraction) Very high P availability (5) Convolvulus arvensis Galinsoga parviflora Sonchus oleraceus Chaenarrhinum minus

Fairly good P availability (3) Matricaria discoidea Fumaria officinalis Erodium cicutarium Viola tricolor Plantago major

High P availability (4) Lamium purpureum Tussilago farfara Urtica urens Sisymbrium officinalis Arenaria serpyllifolia Senecio vulgaris Consolida regalis

Poor P availability (2) Aphanes arvensis Rumex crispus Polygonum lapathifolium Sagina procumbens Vicia tetrasperma Spergula arvensis Anthemis cotula Raphanus raphanistrum

Very poor P availability (1) Juncus bufonius Galium uliginosum Ranunculus repens Arabidopsis thaliana Valerianella locusta Scleranthus annuus Indifferent species (×) Galium aparine Polygonum aviculare Poa annua Vicia hirsuta Veronica arvensis Galeopsis tetrahit Centaurea cyanus Equisetum arvense

From Nowack (1990)

a few sandy fields) due to intensive fertilisation (Scheffer and Schachtschabel 2010). However, the type of crop species matters for P and K availability, as maize fields typically receive much larger amounts of P and K (and Ca and Mg) fertiliser than winter cereal fields due to the much greater use of organic fertilisers, in central Germany around 3–5 times more (Seifert et al. 2015b). Manthey (2003) found K availability to be low only in the acidic soils supporting the Sclerantho-­ Arnoseridetum minimae community, which has mostly now disappeared. A different impression of the role of P is given by the results of Nowack (1990), who found a clear relationship between the occurrence of certain segetal species and the soil content of CAL-P in central and northwestern German crop fields. He was even able to identify several weed species as regional indicators for 6 classes of P supply (see Table 12.4). However, the differences in P availability in segetal vegetation can only be accurately characterised once the role of arbuscular mycorrhizae in these communities is better understood.

12.4.5  The Effects of Cultivation The frequent occurrence of root crop or cereal crop weeds in different years in the same field is doubtless caused by the promotion of different groups of segetal species by different crop plants and management regimes. So what are the factors causing these inter-annual fluctuations in the vegetation? The growth form of the crop appears to play a relatively minor role, and only very few weed species show a clear preference for a particular crop species (van

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Fig. 12.11 Average species richness in cereal and root crop relevés in Bavaria in 1951–1968, and again in 1986–1988, showing the relative proportions of character species of cereal and root crop weed communities in the stands (Modified from Albrecht 1989)

Elsen 1994). For example, winter wheat, winter barley and oil seed rape have very similar coverage and therefore provide very similar light levels, and even summer cereals and root crops now have only a slightly lower coverage (van Elsen 1994; Seifert et  al. 2015b). The differences between the microclimate within different crop types are now therefore relatively unimportant (Seifert et al. 2015a). The species richness of root and cereal crops has also become increasingly similar (see Fig. 12.11; see also May 1986). Old farming lore referring to a linkage between certain weeds and crops can also be explained via other mechanisms. For example, rye fields blue with cornflowers mainly reflect the concentration of both rye cultivation and Centaurea cyanus on nutrient-poor sandy soils, whilst the link between wheat fields and poppies is caused by the fact that both occur on richer soils (van Elsen 1994). In addition to the nutrient supply, the type of cultivation is also more important than the crop type itself: summer wheat, oats and maize are cereals, but their segetal flora is more similar to that of root crops. Nevertheless, there is no direct relationship between the mechanical disturbance and species composition of the crop weeds. ‘Root crop weeds’ are not any more resistant to frequent disturbance than most ‘cereal weeds’. If anything, they are less resistant, as an experiment in crop fields near Stuttgart with 1 to 16 rounds of tillage per year showed (Ellenberg, unpublished). In fact, it is more the time at which the last intensive soil cultivation occurs. If this is in late autumn or winter, then a typical ‘cereal weed community’ forms, whether cereal is planted or not. If the soil is only worked in May, June, or even later in summer, or if the weeds are controlled with herbicides around this time, then the same crop field will develop a ‘root crop weed community’. However, this will only occur if the soil had been planted with alternating root and cereal

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crops for several decades previously. Fields with conservation tillage (i.e. with no ploughing) may have a different species composition of the segetal flora, with a strong increase in the proportion of geophytes and a decrease in therophytes (Froud-­ Williams et al. 1983).

12.4.6  The Effects of Herbicide Application The application of selective herbicides has largely replaced mechanical weed control in the last 40 to 50 years, becoming almost ubiquitous in Central Europe. In intensive crop management, 15 or more different pesticides (mostly herbicides and fungicides) are used during a growing season, with higher use intensity in regions with more favourable conditions for arable land use (Andert et al. 2015). This practice causes probably the most fundamental changes to segetal communities of any management technique (Helmecke and Mahn 1984). The tolerance of arable weeds to herbicides partly depends on their germination period. Cold-germinating species, i.e. the winter cereal weeds, are particularly sensitive, such as Consolida regalis, Adonis aestivalis, Myosurus minimus, Legousia speculum-veneris, Scandix pecten-veneris, Ranunculus arvensis, Lithospermum arvense, Veronica triphyllos, Centaurea cyanus, Veronica hederifolia, Papaver rhoeas, Valerianella locusta and Vicia hirsuta. Species that do not require low temperatures to germinate react in various ways to herbicide application. For example Stellaria media and Capsella bursa-pastoris are relatively insensitive, because they are able to germinate throughout spring, summer and autumn and have a very short developmental cycle. Several grasses are also largely unaffected by herbicides, including Alopecurus myosuroides, Digitaria ischaemum, Apera spica-venti, Avena fatua and Poa annua, which all have a short developmental cycle and thus a high seed production rate, guaranteeing sufficient reproduction even with intensive herbicide treatment. Species with a flexible life cycle that can react to any type of disturbance may also be successful, allowing them to produce large quantities of seeds and maintain a diaspore bank under any conditions. Species with strong vegetative dispersal such as Elymus repens can only be controlled with broad-band herbicides. The grasses often even profit from the reduced competition from dicot weeds on sprayed fields, as growth regulating herbicides primarily affect the broadleaved dicots. Local taxa can adapt to avoid the effects of the herbicides (Pötsch and Busch 1985; Sturny et al. 1984; Börner 1995). Chenopodium ficifolium, for example, now germinates over a much wider temperature range than it did 50 years ago. This adaptation allows it to germinate in both spring and summer so that it can largely escape the effects of herbicide application (Otte 1991). van Elsen (1994) found that if herbicides are applied to a crop field after several years of being unsprayed, then the number of arable weed species sinks to a quarter of its previous figure, and their coverage is almost zero (see Fig. 12.12). The herbicides often almost completely wipe out the weeds in the centre of the field, so that

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Fig. 12.12  Species richness in margins of arable fields on Triassic sandstone (Buntsandstein) in northern Hesse that had not previously been treated with herbicide before, during and after herbicide application (Modified from van Elsen 1994)

only a few species such as Veronica arvensis, Viola arvensis, Stellaria media or Galium aparine with wide ecological amplitudes are able to survive. For example, maize fields in Germany are on average colonised by only nine weed species with Chenopodium spp., Stellaria media, Fallopia convolvulus and Echinochloa crus-­ galli being the most frequent ones (de Mol et al. 2015). These assemblages have no character or differential species, and are usually impossible to assign to a segetal community (van Elsen 1994). Ten years of herbicide use in a Bavarian crop field led to the loss of the diaspore bank of many species (Albrecht 1989), making a rapid regeneration after the cessation of herbicide use no longer possible. In other regions, the typical segetal flora had largely disappeared even after 5 years of intensive herbicide and fertiliser application (Bischoff 1996). Modern conventional arable farming thus leads to crop monocultures with typically only 3–6 nitrophilic weeds per relevé (100 m2) that are adapted to the use of herbicides; the weed cover is often 200 kg N ha−1 in inorganic and organic form) to the mineralisation gives an annual N supply in the range of around 250 up to >400 kg N per ha and growing season, whereby bacterial mineralisation contributes on average less than 50 % of the total amount. Legumes as catch crops can further increase the N input. Ploughing and the resulting aeration of the soil mean that the nitrification rate is very high in almost all arable fields. Mineral N fertiliser is generally first immobilised by microbes and only available for plant uptake after a delay (Körschens and Mahn 1995); in clayey soils, large amounts of ammonium (up to 100 kg N ha−1 in the top 30 cm) are also fixed on the clay minerals (Nieder 1996). Fertiliser application can slow bacterial mineralisation (Zoldan 1981). If a large amount of easily decomposable organic material is available in the soil under a humid climate, then large amounts of nitrate can be leached from fertilised arable fields and fallows (up to 100  kg N ha−1 year−1 and more; Waldhardt and Schmidt 1993; Scheffer/ Schachtschabel 2010). Nieder (1996) found that the degree of leaching decreased from potatoes to cereals to sugar beet. Thriving weed communities can help to reduce this by taking up excess nitrate and storing it through the growing season, in the same way as winter cover crops. However, N fertiliser application is now somewhat lower than in the late twentieth century, and modern fertilisation practices have helped to reduce the nitrate leaching in general. Nieder (1996) calculated average NO3− losses for Lower Saxony of 16 kg N ha−1 year−1 on heavy soils and 63 kg

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N ha−1 year−1 on light soils. Eighty percent of the N input in freshwater systems in Germany originates from agricultural land (SRU 2015). Denitrification leads on average to a loss of 7 % of the fertiliser as N2O emission, i.e. release rates of 3–30  kg N ha−1 year−1 (Scheffer/Schachtschabel 2010; SRU 2015). Jungkunst et al. (2006) give maximum values for N2O emissions from Central European arable fields as 17 kg ha−1 year−1. The considerable N accumulation in intensively fertilised agricultural systems is demonstrated by an approximate N balance calculated for the utilised agricultural area of western Germany (arable and grassland) for 1986, when N fertilisation was highest: total input was 218  kg N ha−1 year−1 (synthetic fertiliser: 126, imported fodder: 47, atmospheric deposition and biological N fixation: 45) and average net loss in form of harvested plant and animal products was only 51  kg. In that time, the soil accumulated around 48 kg N ha−1 per year. 44 kg year−1 was lost as gaseous ammonia and 25 kg year−1 as nitrous oxide via denitrification. A further 45 kg N ha−1 year−1 was lost via seepage into the groundwater and 5 kg through erosion (Isermann 1990; Baeumer 1992). In 2010, the arable land of Germany was burdened with an average surplus of 70 kg N ha−1 year−1, after subtracting the N export with harvest and denitrification from the N input (SRU 2015). An estimate of important nitrogen fluxes on a national level is given for Germany in Fig. 3.35 in Volume I. In the case of phosphorus, large decreases in P fertilisation have led to lower excesses of P and a decrease in P pools in many arable ecosystems in Central Europe in the last two to three decades. The associated P load on freshwater and marine ecosystems has decreased as well (see Figs. 4.33 and 1.29).

12.7  Vegetation Dynamics 12.7.1  I nterannual Fluctuations and Changes with Crop Rotation As many of its species are short-lived, the arable weed communities react more strongly to annual changes in habitat conditions than most other plant communities. A wet spring, for example, promotes species that germinate in damp soils, a dry summer promotes perennial weeds, because the annuals can produce fewer ripe seeds than in normal years (Koch 1955), whilst high levels of fertiliser application allow particularly nutrient-demanding species to spread. However, the segetal communities are primarily influenced by the yearly change in the crop species as well as the cultivation regime and the herbicides used. This causes the formation of what Rademacher (1948) described as a series of ‘aspects’ of the segetal communities. Changes in the composition of arable weed communities caused by crop rotation on the same field were studied by van Elsen (1994) in mostly base-rich arable land in central Germany and the Eifel. As expected, summer annuals of root

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crop communities (alliance Veronico-Euphorbion) were promoted by summer cereal cultivation, whereby the Spergulo-Chrysanthemetum segetum was even restricted to this cultivation form. Winter cereal cultivation, in contrast, promoted species of the Papaveretum argemone such as Papaver argemone and Veronica triphyllos (see Fig.  12.15). If cereals were replaced by potatoes in the following year, then Mercurialis annua, Sonchus asper, Fumaria officinalis and Atriplex patula became dominant (see Fig. 12.16). However, these annual changes did not generally produce different associations, but only a shift in the proportions of species within the community.

12.7.2  Secondary Succession on Arable Fallows Whilst the management of easily accessible and large arable fields has intensified in recent decades, areas close to cities or of low-yielding soils have in places been abandoned, allowing spontaneous succession to occur. These examples of old-field succession have been the subject of many permanent plot or chronosequence studies in Central Europe, e.g. in the Czech Republic (e.g. Leps 1987; Prach 1982; Osbornova et al. 1990; Prach et al. 2007), Poland (Symonides 1985; Falinski 1986a) and Germany (e.g. W. Schmidt 1988, 1993; van Elsen and Günther 1992; Tischew 1994; Schmiedeknecht 1995; Hilbig 1996; Toetz 2000). The direction and speed of secondary succession and its basic processes are largely determined by the site conditions, the previous management of the field, and the surrounding vegetation and its species pool. As a result, very variable combinations of species occur particularly in the first few years after abandonment, dominated by species of segetal and ruderal communities. However, these variable initial stages become increasingly similar with the development in the direction of woody vegetation. Within 25 years, a pioneer forest of over 10 m in height can form on an arable field with mesic soils, as was shown by an old field experiment in Göttingen (Schmidt 1981, 1993), as well as a comparative study of different successional stages in karst landscapes in western Czech Republic (Osbornova et al. 1990). In the initial successional stages in the Göttingen experiment, the vegetation was dominated by species of the classes Stellarietea mediae and Sisymbrietea, i.e. annual arable weeds and ruderal plants. However, these were quickly replaced by the class Artemisietea, i.e. perennial ruderal plants and arable weeds (see also Krumbiegel et al. 1995). Species of other communities of disturbed ground (namely the Agropyretalia intermedio-repentis, the Plantaginetea and Agrostietalia stoloniferae) were relatively rare. Instead, a larger proportion of the tall herbs in years 3–9 of the succession was made up of species of agricultural grassland (Molinio-Arrhenatheretea), heathland and Nardus grassland (NardoCallunetea), and dry grassland (Festuco-Brometea). Many of the species in these classes are helped in establishing by having good dispersal mechanisms. Species of forest clearings (Epilobietea angustifoliae) are also known for their dispersal ability, and the proportion of these species, as well as that of the herbaceous

824 Fig. 12.15  Fluctuations in the arable weed vegetation caused by the alternation between winter and summer crops in a calcareous field in northern Hesse (central Germany). Excerpt from van Elsen (1994). Sowing: s = summer, w = winter. Herbicides: y = yes, n = no

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825

Fig. 12.16  Fluctuations in the arable weed vegetation caused by the alternation between cereal and root crops in a calcareous field in northern Hesse (central Germany), including relevés from the field margin (M) and interior (I). Excerpt from van Elsen (1994). Herbicides: n = no herbicides. Sowing/planting: A = autumn, S = summer, W = winter

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fringes of scrub and forest (Trifolio-Geranietea) increased particularly after 10 years of succession. The decisive role, which is played in the first 30 years by local factors such as the diaspore pool, the species pool in the surrounding area, the degree of herbivory and the previous management (Krumbiegel and Klotz 1996; Prach et al. 2007), is visible when comparing the dominant plant species over time in the three old field succession experiments in Göttingen (on mesic loess loam) and western Czech Republic (on mesic Terra fusca or dry Cambisols). In Göttingen, these were (with co-­dominant species after the slash and sub-dominants in brackets): Papaver rhoeas (Sonchus spp.) – Galium aparine (Conyza canadensis) – Epilobium spp. – Tussilago farfara – Picris hieracioides  – Solidago canadensis  – Salix caprea  – Fraxinus excelsior (Betula pendula). In the mesic Czech fields: Tripleurospermum inodorum/Stellaria media – Elymus repens – Arrhenatherum elatius/Festuca pratensis – Prunus spinosa/Crataegus spp. In the dry Czech fields: Papaver rhoeas – Carduus acanthoides  – Elymus repens/Artemisia vulgaris  – Poa angustifolia/Festuca rupicola (Schmidt 1981, 1988; Osbornova et  al. 1990). In strongly disturbed landscapes, invasive species such as Solidago canadensis, Calamagrostis epigejos and Epilobium and Aster species can form monodominant stands after only a few years (Prach et al. 2007). Landscape-scale comparisons show that soil moisture is a major determinant of the course of old-field succession (Ejrnaes et al. 2003; Prach et al. 2014a). In general, the effect of soil conditions is highest in early successional stages, while that of the structure of the surrounding landscape increases toward later stages (Prach et al. 2014b). Despite these differences, it is still possible to make some generalisations about the course of Central European arable succession and its drivers: 1. The initially dominant therophytes are replaced by biennial and then by perennial grasses, forbs and finally long-lived woody species, i.e. plant life-span increases rapidly (see Figs. 12.17a and 12.18). The perennial species are usually first dominated by species with effective vegetative dispersal such as Elymus repens and Solidago canadensis (Schmiedeknecht 1995). These changes are reflected in a transition in plant strategy types (Grime 2001) from ruderals (R) to competitors (C–R, C) and finally to stress-tolerants (S). The establishment of woody vegetation can be delayed by several decades by lack of water, dense vegetation of forbs and grasses, as well as a high grazing pressure (cf. Sect. 8.7.2). A rapid increase in the plant biomass therefore only occurs in successions with early establishment of woody vegetation. 2. In the initial stages, germination success and therefore the species composition of the stand is mainly determined by the small-scale heterogeneity in soil properties. In later stages, competitive ability plays a more important role. 3. The volume of seed production decreases continuously throughout succession (Symonides 1986; Albrecht 2004). The seed bank reaches its maximum after around 10 years, after which it begins to decline again (see Fig. 12.17b). 4. The species diversity of vascular plants may increase or decrease over the course of old-field succession. Changes in diversity are tightly linked to the dom-

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Fig. 12.17  Changes in the relative proportions of plant life forms ((a) dry soils, (b) mesic soils), above-ground biomass (c) (X: dry soils, M: mesic soils), number of seeds (d) and soil temperature (e) (A: max., B: min.) during the first 60 years of succession after abandonment of an arable field in western Czech Republic (Modified from Osbornova et al. 1990; with permission of Springer Science+Business Media)

inant species and their degree of dominance (Monk 1983). In the old-field succession in Göttingen, the highest species richness was found in the 4th year, during the tall herb stage (see Fig. 12.18), in the Czech Republic this was in the 3rd year (see Fig. 12.17c). The number of species dropped again with the formation of a shrub layer in the tall herb-scrub stage (around 5 to 15 years after abandonment). 5. In the first 30 years, the processes of species change mainly follow the ‘tolerance’ and ‘inhibition’ models of Connell and Slatyer (1977), whilst the establishment of woody species increases the importance of ‘facilitation’ (see Sect. 6.7.3 for the heathland-to-forest succession). 6. If old field succession follows a sequence of plant strategy types from ruderal to competitive and finally stress tolerant, then the speed of vegetation change will decrease with the progression of succession, and the stability of the vegetation will increase (Leps 1987). At the same time, its resistance to external disturbance

Fig. 12.18  Changes over time in different layers of the vegetation (top), in the proportions of different plant life forms (middle: P phanerophytes, C chamaephytes, N nanophanerophytes, H hemicryptophytes, G geophytes, T therophytes) and in species richness (bottom; solid line, Shannon-Index = dotted line) over almost 40 years of undisturbed succession in an arable fallow on fertile loam in the Experimental Botanic Garden in Göttingen (From W. Schmidt (1993), Dölle and Schmidt (2007) and W. Schmidt (unpublished))

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829

will also increase, whilst its resilience, i.e. the speed at which it recovers after disturbance, decreases. 7. The maximum soil surface temperature decreases as the stand becomes denser, whilst the depth to which the ground freezes decreases (see Fig. 12.17d). 8. Abandonment can initially lead to an increase in nitrate leaching from arable fields, because large amounts of nutrients are no longer removed with the harvest (Jagnow and Söchtig 1981; Russow et  al. 1995). However, in the long-term, nitrogen is fixed in long-lived biomass, thereby reducing the losses. 9. After several decades of arable succession, more phosphorus is present in organic and occluded forms and is therefore less available for plant uptake than at the beginning of succession (Osbornova et al. 1990; Schmidt 1993a). A reduction in potassium availability has also been observed over this time frame (Schmidt 1981, 1993; Wurbs and Glemnitz 1997), while the soil nitrogen content generally declines very slowly (Büring 1970; Karrer et al. 1997). Over long time spans, succession from arable land to forest leads to humus accumulation and rising soil carbon and nitrogen pools (cf. Sect. 6.7.3). Successional trends in nutrient pools and availability depend on the fertilisation level of the fields prior to succession.

12.8  Human Influence 12.8.1  T  he Recent Collapse of Arable Weed Populations and Its Causes Patterns of Change  The weed communities of Central European crop fields have been constantly changing since the Neolithic, i.e. for as long as this habitat has existed, experiencing both gains and losses in species (Burrichter et al. 1993; Meyer et al. 2015c). However, changes in crop types and the introduction of mechanical seed cleaning meant that various segetal species that had still been widespread in 1860 had already been largely lost by 1950, as observed by Koch (1955) in the Stuttgart region. The rapid mechanisation of soil cultivation and the beginning of widespread herbicide use in the 1950s/1960s caused the loss of species from segetal communities to accelerate. Tüxen first noted the loss of character species of arable weed communities in 1962, and in the following 50 years, most regions of Central Europe reported dramatic declines in the segetal flora, e.g. in Stolzenau on the Weser (Meisel 1966a, b), Southern Germany (Bachthaler 1968), Northwestern Germany (Meisel and Hübschmann 1976), Southern Lower Saxony (Wagenitz and Meyer 1981), the Münster region (Reuß 1981), Northern Hesse (Hotze and van Elsen 2006), the Ore Mountains (Köck 1984), the Lower Rhine region (Kutzelnigg 1984), Saxony-Anhalt (Hilbig 1985), various regions of Eastern Germany (Pötsch and Busch 1985), Bremen (Kulp and Cordes 1986), Bavaria (Otte 1984; Braun 1988; Albrecht 1989), Austria (Ries 1992), Northwestern Poland (Borowiec 1988), Eastern Germany (Kläge 1999), Czech Republic (Chytrý 2009), Slovakia (Májeková

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et al. 2010), Thuringia (Kohlbrecher et al. 2012), and central and Northern Germany (Meyer et al. 2013, 2015b). Indeed, no other habitat in Central Europe has experienced such drastic reductions in population sizes and diversity in the last 40–50 years as the arable fields. In many regions, the collapse of segetal communities exceeds that of the mesic and moist grassland habitats (see Sect. 8.8.1). In Germany, around 120 of the approximately 350 segetal taxa are now threatened, and 15 to 18 are extinct (Korneck and Sukopp 1988; Schneider et al. 1994). 92 of these threatened species (around 69 %) are segetal species in the stricter sense, i.e. they are mainly found in crop fields, vineyards or gardens. The majority of these are species of cereal fields (73 taxa). In a meta-analysis of 13 studies on the decline in the segetal flora of Germany between the 1950s and 1980s (Albrecht 1989), 67 of 96 arable weed species showed on average across the studies a significant decline in abundance or a reduction by at least one constancy class. Only 9 species (Avena fatua, Echinochloa crus-galli, Galium aparine, Lamium purpureum, Lapsana communis, Matricaria discoidea, Tripleurospermum inodorum and the Galinsoga species) showed on average an increase (see Table 12.7). Kläge (1999) compared the frequency of segetal species in around Cottbus in eastern Germany in the 1990s with data from older floristic studies, and found that of the 282 segetal species occurring in the region, today only 32 are common and 35 frequent, whilst 47 are sporadic, 61 rare and 20 very rare. 90 species that were once found in crop fields in the region no longer occur there (at least in arable habitats). Only 28 % of the species occurred at the same frequency as they did before agricultural intensification, whilst 72 % had declined to some extent. Albrecht (1989) repeated earlier surveys in 195 crop fields in seven landscapes in Bavaria. Whilst a total of 209 vascular plant species were found in 1951–1968 in the fields, by 1986–1988 there were only 168, meaning an average decrease of 20 %. 39 species had suffered a significant population decline. Meyer et al. (2013) conducted a re-sampling study in 392 semi-permanent plots in central and northern German fields and found a reduction in the regional species pool by 23 % since the 1950s/60s. In other regions of Germany (around Bremen, eastern Germany, Swabian Jura), decreases in the total number of arable weed species of between 17 and 52 % were found (Mittnacht 1980; Pötsch and Busch 1985; Kulp 1993). Arable communities only remained stable for longer in a few areas such as the Lower Rhine region (Kutzelnigg 1984) and in the Thuringian Orla valley (Xylander 1987). Not only the regional species pool, but also the species diversity of individual stands has declined significantly in recent decades, for example by 30 % in Bavaria from an average of 23 species per plot (1951/1968) to 16 (1986/1988, Albrecht 1989), and from 24 species (median) in the 1950s/1960s to 7 species in the field interior in 2009 in central and northern Germany (Meyer et al. 2013; see Fig. 12.19). Other Central European authors found decreases in the average species richness of 20–50 % (Hilbig and Bachthaler 1992a and b; Ries 1992), and in extreme cases of 65 %, as in the case of Stolzenau on the Weser (Meisel 1979). Whilst characteristic species combinations could be found on areas of 25–30  m2 in the 1950s, by the 1970s these could only be distinguished on 500–1000 m2 (Meisel 1979).

12.8  Human Influence

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Table 12.7  Changes in the frequency of various arable weeds species in 13 regions of Germany over the course of 30 years (roughly 1950–1965 to 1979–1988) Marked decrease: Adonis aestivalis, Agrostemma githago, Anagallis arvensis, Centaurea cyanus, Cerastium holosteoides, Consolida regalis, Convolvulus arvensis, Daucus carota, Equisetum arvense, Medicago lupulina, Neslia paniculata, Odontites verna, Scleranthus annuus, Sherardia arvensis, Sonchus oleraceus, Vicia sativa Decrease: Achillea millefolium, Agrostis stolonifera, Alopecurus myosuroides, Anagallis foemina, Anthemis arvensis, Aphanes arvensis, Arenaria serpyllifolia, Arnoseris minima, Campanula rapunculoides, Chaenarrhinum minus, Chenopodium polyspermum, Cirsium arvense, Euphorbia exigua, Gnaphalium uliginosum, Holosteum umbellatum, Juncus bufonius, Kickxia elatine, K. spuria, Lathyrus tuberosus, Legousia speculum-veneris, Lithospermum arvense, Mentha arvensis, Myosotis arvensis, Myosurus minimus, Oxalis fontana, Papaver argemone, P. rhoeas, Plantago intermedia, Ranunculus arvensis, Raphanus raphanistrum, Rhinanthus alectorolophus, Rumex acetosella, Sagina procumbens, Scandix pecten-veneris, Sedum telephium, Setaria viridis, Silene noctiflora, Sinapis arvensis, Sonchus arvensis, S. asper, Spergula arvensis, Stachys palustris, Trifolium repens, Valerianella locusta, Veronica arvensis, V. hederifolia, V. triphyllos, Vicia angustifolia, V. hirsuta, V. tetrasperma Mostly constant: Aethusa cynapium, Apera spica-venti, Atriplex patula, Capsella bursa-pastoris, Chenopodium album, Elymus repens, Euphorbia helioscopia, Fallopia convolvulus, Fumaria spp., Galeopsis spp., Lamium amplexicaule, Matricaria chamomilla, Poa annua, Polygonum aviculare, P. hydropiper, P. persicaria, Stellaria media, Taraxacum officinale, Thlaspi arvense, Veronica persica, Viola arvensis Increase: Avena fatua, Echinochloa crus-galli, Galinsoga spp., Galium aparine, Lamium purpureum, Lapsana communis, Matricaria discoidea, Tripleurospermum perforatum From various authors compiled in Albrecht (1989)

Unsurprisingly, the coverage of arable weed communities has also declined s­ignificantly along with their species richness. Many crop fields are now almost completely free of weeds, particularly in the centre of the field. The success of the battle against weeds is shown by Seifert et al. (2015b), who found in 2011/2012 an average weed coverage of