Ecosystems of disturbed ground

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ECOSYSTEMS OF THE WORLD 16

ECOSYSTEMS OF DISTURBED GROUND

ECOSYSTEMS OF THE WORLD Editor in Chief: David W. Goodall Centre for Ecosystem Studies, Edith Cowan University, Joondalup, W.A. (Australia)

I. TERRESTRIAL ECOSYSTEMS

1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13. 14. 15. 16.

A. Natural Terrestrial Ecosystems Wet Coastal Ecosystems Dry Coastal Ecosystems Polar and Alpine Tundra Mires: Swamp, Bog, Fen and Moor Temperate Deserts and Semi-Deserts Coniferous Forests Temperate Deciduous Forests Natural Grasslands Heathlands and Related Shrublands Temperate Broad-Leaved Evergreen Forests Mediterranean-Type Shrublands Hot Deserts and Arid Shrublands Tropical Savannas Tropical Rain Forest Ecosystems Wetland Forests Ecosystems of Disturbed Ground

17. 18. 19. 20. 21.

B. Managed Terrestrial Ecosystems Managed Grasslands Field Crop Ecosystems Tree Crop Ecosystems Greenhouse Ecosystems Bioindustrial Ecosystems

II. AQUATIC ECOSYSTEMS

22. 23.

A. Inland Aquatic Ecosystems River and Stream Ecosystems Lakes and Reservoirs

24. 25. 26. 27. 28.

B. Marine Ecosystems Intertidal and Littoral Ecosystems Coral Reefs Estuaries and Enclosed Seas Ecosystems of the Continental Shelves Ecosystems of the Deep Ocean

29.

C. Managed Aquatic Ecosystems Managed Aquatic Ecosystems

III. UNDERGROUND ECOSYSTEMS 30.

Subterranean Ecosystems

ECOSYSTEMS OF THE WORLD 16

ECOSYSTEMS OF DISTURBED GROUND Edited by Lawrence R. Walker Department of Biological Sciences, University of Nevada, Las Vegas, 4505 Maryland Parkway, Box 454004, Las Vegas, NV 89154-4004, USA

1999 ELSEVIER Amsterdam – Lausanne – New York – Oxford – Shannon – Singapore – Tokyo

Elsevier Science B.V. Sara Burgerhartstraat 25 P.O. Box 211, 1000 AE Amsterdam, The Netherlands © 1999 Elsevier Science B.V. All rights reserved. This work is protected under copyright by Elsevier Science B.V., and the following terms and conditions apply to its use: Photocopying Single photocopies of single chapters may be made for personal use as allowed by national copyright laws. Permission of the publisher and payment of a fee is required for all other photocopying, including multiple or systematic copying, copying for advertising or promotional purposes, resale, and all forms of document delivery. Special rates are available for educational institutions that wish to make photocopies for non-proﬁt educational classroom use. Permissions may be sought directly from Elsevier Science Rights & Permissions Department, PO Box 800, Oxford OX5 1DX, UK; phone: (+44) 1865 843830, fax: (+44) 1865 853333, e-mail: [email protected]. You may also contact Rights & Permissions directly through Elsevier’s home page (http://www.elsevier.nl), selecting ﬁrst ‘Customer Support’, then ‘General Information’, then ‘Permissions Query Form’. In the USA, users may clear permissions and make payments through the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, USA; phone: (978) 7508400, fax: (978) 7504744, and in the UK through the Copyright Licensing Agency Rapid Clearance Service (CLARCS), 90 Tottenham Court Road, London W1P 0LP, UK; phone: (+44) 171 631 5555; fax: (+44) 171 631 5500. Other countries may have a local reprographic rights agency for payments. Derivative Works Tables of contents may be reproduced for internal circulation, but permission of Elsevier Science is required for external resale or distribution of such material. Permission of the publisher is required for all other derivative works, including compilations and translations. Electronic Storage or Usage Permission of the publisher is required to store or use electronically any material contained in this work, including any chapter or part of a chapter. Except as outlined above, no part of this work may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without prior written permission of the publisher. Address permissions requests to: Elsevier Science Rights & Permissions Department, at the mail, fax and e-mail addresses noted above. Notice No responsibility is assumed by the Publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. Because of rapid advances in the medical sciences, in particular, independent veriﬁcation of diagnoses and drug dosages should be made. First edition 1999 Library of Congress Cataloging-in-Publication Data Ecosystems of disturbed ground / edited by L.R. Walker. − − 1st ed. p.

cm. − − (Ecosystems of the world ; 16)

Includes bibliographical references. ISBN 0−444−82420−0 R.

1. Ecology. 2. Nature− −Effect of human beings on. II. Series.

QH545.A1E2824

I. Walker, L.

1999

577.27− −dc21

99-41984 CIP

ISBN 0 444 82420 0 (Volume) ISBN 0 444 41702 8 (Series) ∞ The paper used in this publication meets the requirements of ANSI/NISO Z39.48-1992 (Permanence of Paper).

Printed in the Netherlands

PREFACE As the human population inexorably grows, its cumulative impacts on the earth’s resources are hard to ignore. The ability of the earth to support more humans is dependent on the ability of humans to manage natural resources wisely. Because disturbance alters resource levels, effective management requires understanding of the ecology of disturbance. Editorship of this book was undertaken with several goals in mind. First, I wanted to present an organized summary of the many types of disturbances that impact the earth, with as global a focus as the existing literature allowed. The book is organized into chapters that deal primarily with natural disturbances (Chapters 2–13), anthropogenic disturbances (Chapters 14–20), overviews of natural processes that occur across disturbance types (Chapters 21–27), and human interactions with and responses to disturbance (Chapters 28–30). Chapter 31 explores a hierarchical view of disturbance; Chapter 32 examines the concept that the consequences of growth of the human population themselves represent the ultimate disturbance, and suggests ways to ameliorate human impacts. “Natural” and “anthropogenic” disturbances generally are interrelated. The focus of this book is on disturbances that have a direct physical impact on terrestrial systems, excluding primarily atmospheric phenomena such as acid rain or increases in carbon dioxide and decreases in ozone (but not wind), and primarily aquatic phenomena such as cultural eutrophication or chemical, thermal, and bacterial pollution of waterways. A second purpose was to enhance understanding of the concept of disturbance in order to manage it better. The development of theory related to disturbance is in its infancy. One of the few book-length overviews of the topic was that provided by Pickett and White (1985), in which they focused on alterations of relatively pristine habitats. With the wealth of examples from around the world presented in this volume, perhaps we can make further progress toward a general theory of disturbance. Alternatively, one may recognize that site-speciﬁc characteristics do not allow such generalizations. To those ends, authors were given some freedom to interpret disturbance as they deemed appropriate. Authors addressing a particular

disturbance type or disturbance in a particular habitat (Chapters 2–20) were asked to address, as far as possible, the disturbance regime, the damage caused by the disturbance, the responses of the biota to the disturbance, and interactions between disturbance types and between disturbance and humans. Authors of process chapters (Chapters 21–27) were to seek generalizations from a global perspective of short- and longer-term responses across various types of disturbed ecosystems. In the ﬁnal Chapter 33, Willig and I evaluate common threads in the previous chapters and make contributions toward the development of a theory about disturbance. A third goal was to provide insights for land managers on how to incorporate lessons about disturbance into their efforts. Some of the chapters (e.g., Chapters 11, 14–16, 18–20) explicitly address managed systems such as pasture-land, urban habitats, and agriculture; other chapters address management issues more generally. Any level of generalization that can be made about disturbance responses will aid managers of disturbed ecosystems. Human well-being on this increasingly crowded planet depends on the success to which land management policies (see Chapter 30) apply such lessons. I wish to thank the many chapter authors for their patience, hard work, excellent insights, and chapter reviews; the 36 non-author peer reviewers for their constructive criticisms that made my job easier; the series editor David Goodall for overall guidance and careful editing; Rachel Lawrence and Dorothy Dean for assistance with correspondence and prooﬁng; the University of Nevada, Las Vegas (UNLV) for providing substantial logistical support; colleagues and graduate students of ecology at UNLV for engaging discussions; Michael Willig for helping me to synthesize the whole volume; and my wonderful wife, Elizabeth Powell, for her continual support. During the development of the ideas presented here, I was supported by NSF grants BSR-8811902 and DEB-9411973 to the Institute for Tropical Ecosystem Studies, University of Puerto Rico, and the International Institute of Tropical Forestry, as part of the Long-Term Ecological Research Program in v

vi

PREFACE

the Luquillo Experimental Forest. Additional support was provided by the U.S. Fish and Wildlife Service, the U.S. National Park Service, and the U.S. Forest Service. Finally, the completion of this book was facilitated by a sabbatical leave from UNLV. Lawrence R. Walker Editor

REFERENCES Pickett, S.T.A. and White, P.S., 1985. The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Orlando, Florida, 472 pp.

LIST OF CONTRIBUTORS E.B. ALLEN Department of Botany and Plant Sciences University of California Riverside, CA 92521-0124, USA

M.L. CADENASSO Institute of Ecosystem Studies Box AB Millbrook, NY 12545-0129, USA

M.F. ALLEN Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

C.R. CARROLL Institute of Ecology University of Georgia Athens, GA 30602, USA J.A. COOKE School of Life and Environmental Sciences University of Natal Durban 4041, South Africa

D. BAINBRIDGE Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

C.M. D’ANTONIO Department of Integrative Biology University of California Berkeley Berkeley, CA 94720, USA

A.H. BALDWIN Natural Resources Management Program Department of Biological Resources Engineering University of Maryland College Park, MD 29742, USA

R. DEL MORAL Department of Botany University of Washington Box 355325 Seattle, WA 98195-5325, USA

C.J. BARROW Centre for Developmental Studies University College of Swansea University of Wales Swansea SA2 8PP, United Kingdom

S. DEMARAIS Department of Wildlife and Fisheries Mississippi State University Mississippi State, MS 39762, USA

D. BINKLEY Department of Forest Sciences Colorado State University Fort Collins, CO 80523, USA

M.B. DICKINSON Department of Biological Sciences Florida State University Tallahassee, FL 32306-2043, USA

I.K. BRADBURY Dept. of Geography University of Liverpool P.O. Box 147 Liverpool L69 3BX, United Kingdom

C. DOLJANIN Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

N.V.L. BROKAW Manomet Center for Conservation Sciences P.O. Box 1770 Manomet, MA 02345, USA vii

viii

LIST OF CONTRIBUTORS

T.L. DUDLEY Department of Integrative Biology University of California Berkeley Berkeley, CA 94720, USA

G.S. HARTSHORN Organization for Tropical Studies Box 90630 Durham, NC 27708-0630, USA

G.E. ECKERT 1434 Pine Street Norristown, PA, USA

C. HARVEY Department of Entomology Cornell University 5126 Comstock Hall Ithaca, NY 14853-0901, USA

F. EDWARDS Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA O. ENGELMARK Swedish Centre for Ecological Sustainability (Swecol) S-901 87 Ume¨a, Sweden C.M. GHERSA Departamento de Ecolog´ıa Facultad de Agronom´ıa, Universidad de Buenos Aires Av. San Martin 4453 1417 Buenos Aires, Argentina M. GIAMPIETRO Istituto Nazionale della Nutrizione Unit of Special Food Technology Via Ardeatina 546 00178 Rome, Italy S.Yu. GRISHIN Institute of Biology and Pedology Russian Academy of Sciences Vladivostok 690022, Russia P.J. GUERTIN Environmental Division U.S. Army Construction and Engineering Research Lab Box 4005 Champaign, IL 61820, USA S. HARNEY Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

C. HINKSON Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA R.J. HOBBS CSIRO Wildlife and Ecology Private Bag, PO Wembley WA 6014, Australia D.W. JOHNSON Biological Sciences Center Desert Research Institute P.O. Box 60220 Reno, NV 89506, USA E.E. JORGENSEN Department of Range, Wildlife and Fisheries Management Texas Tech University Lubbock, TX 79409, USA ´ ´ V. KOMARKOV A Villa Elisabeth 5 CH-1854 Leysin, Switzerland ´ R.J.C. LEON Departamento de Ecolog´ıa Facultad de Agronom´ıa, Universidad de Buenos Aires Av. San Martin 4453 1417 Buenos Aires, Argentina J. LORETI Facultad de Agronom´ıa, Universidad de Buenos Aires Av. San Martin 4453 1417 Buenos Aires, Argentina

LIST OF CONTRIBUTORS

ix

M.D. LOWMAN The Mary Selby Botanical Gardens 811 S. Palm Ave. Sarasota, FL 34236, USA

S.T.A. PICKETT Institute of Ecosystem Studies Box AB Millbrook, NY 12545-0129, USA

R. MACALLER Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

D. PIMENTEL Department of Entomology Cornell University 5126 Comstock Hall Ithaca, NY 14853-0901, USA

M. MACK Department of Integrative Biology University of California Berkeley Berkeley, CA 94720, USA

M. RILLIG Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

J.A. MACMAHON College of Science Utah State University UMC 4400 Logan, UT 84322-4400, USA

P.W. RUNDEL Department of Biology University of California Los Angeles Los Angeles, CA 90024, USA

J.A. MATTHEWS Department of Geography University of Wales Swansea Singleton Park Swansea SA2 8PP, United Kingdom

T.D. SCHOWALTER Entomology Department Oregon State University Corvallis, OR 97331-2907, USA

M.A. MCGINLEY Ecology Program Department of Biological Sciences Texas Tech University Lubbock, TX 79409-313, USA

B. SCHULTZ Biological Sciences Center Desert Research Institute P.O. Box 60220 Reno, NV 89506, USA

K.L. MCKEE National Wetlands Research Center 700 Cajundome Blvd. Lafayette, LA 70506, USA

M. SEMMARTIN Facultad de Agronom´ıa, Universidad de Buenos Aires Av. San Martin 4453 1417 Buenos Aires, Argentina

M. OESTERHELD Departamento de Ecolog´ıa Facultad de Agronom´ıa, Universidad de Buenos Aires Av. San Martin 4453 1417 Buenos Aires, Argentina

¨ C. SIGUENZA Department of Botany and Plant Sciences University of California Riverside, CA 92521-0124, USA

J.M. PARUELO Facultad de Agronom´ıa, Universidad de Buenos Aires Av. San Martin 4453 1417 Buenos Aires, Argentina

R.E. SOJKA USDA Agricultural Research Service Northwest Irrigation and Soils Research Lab 3793N-3600E Kimberley, ID 83341, USA

x

LIST OF CONTRIBUTORS

U. STARFINGER ¨ Institut f¨ur Okologie Technische Universit¨at Berlin Schmidt-Ott Strasse 1 D-12165 Berlin, Germany

F. WIELGOLASKI Department of Biology, University of Oslo, Box 1045, Blindern N-0316 Oslo, Norway

H. SUKOPP ¨ Institut f¨ur Okologie Technische Universit¨at Berlin Schmidt-Ott Strasse 1 D-12165 Berlin, Germany

M.R. WILLIG Ecology Program Department of Biological Sciences and the Museum Texas Tech University Lubbock, TX 79409-313, USA

D.J. TAZIK Environmental Division U.S. Army Construction and Engineering Research Lab Box 4005 Champaign, IL 61820, USA

S.D. WILSON Department of Biology University of Regina Regina, Saskatchewan S45 0A2, Canada

L.R. WALKER Department of Biological Sciences University of Nevada, Las Vegas 4505 Maryland Parkway Box 454004 Las Vegas, NV 89154-4004, USA S.L. WEBB Biology Department Drew University Madison, NJ 07940-4000, USA D.F. WHIGHAM Smithsonian Environmental Research Center Box 28 Edgewater, MD 21037, USA J.L. WHITMORE Vegetation Management and Protection Research USDA Forest Service, Box 96090 Washington, DC 220090-6090, USA

J. WU Department of Life Sciences Arizona State University West Box 37100 Phoenix, AZ 85069, USA L.C. YOSHIDA Department of Botany and Plant Sciences University of California Riverside, CA 92521-0124, USA T.A. ZINK Department of Biology Soil Ecology and Restoration Group San Diego State University San Diego, CA 92182-0057, USA

CONTENTS PREFACE

. . . . . . . . . . . . . . . . . . . . . . .

LIST OF CONTRIBUTORS

. . . . . . . . . . . . . .

Chapter 1. AN INTRODUCTION TO TERRESTRIAL DISTURBANCES by L.R. Walker and M.R. Willig . . . . . . Chapter 2. DISTURBANCE REGIMES AND ECOSYSTEM RESPONSE ON RECENTLY-DEGLACIATED SUBSTRATES by J.A. Matthews . . . . . . . . . . . . .

v vii

1

17

Chapter 3. STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS by V. Kom´arkov´a and F.E. Wielgolaski . .

39

Chapter 4. ECOLOGICAL EFFECTS OF EROSION by D. Pimentel and C. Harvey . . . . . . .

123

Chapter 5. VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY by R. del Moral and S.Yu. Grishin . . . . .

137

Chapter 6. BOREAL FOREST DISTURBANCES by O. Engelmark . . . . . . . . . . . . . .

161

Chapter 7. DISTURBANCE BY WIND IN TEMPERATEZONE FORESTS by S.L. Webb . . . . . . . . . . . . . . . Chapter 8. BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS by D.F. Whigham, M.B. Dickinson and N.V.L. Brokaw . . . . . . . . . . . . . . . Chapter 9. FOREST HERBIVORY: INSECTS by T.D. Schowalter and M.D. Lowman

. . .

Chapter 10. DISTURBANCE IN MEDITERRANEANCLIMATE SHRUBLANDS AND WOODLANDS by P.W. Rundel . . . . . . . . . . . . . . Chapter 11. GRAZING, FIRE, AND CLIMATE EFFECTS ON PRIMARY PRODUCTIVITY OF GRASSLANDS AND SAVANNAS by M. Oesterheld, J. Loreti, M. Semmartin and J.M. Paruelo . . . . . . . . . . . . . . . .

187

Chapter 12. DISTURBANCE IN DESERTS by J.A. MacMahon . . . . . . . . . . . .

307

Chapter 13. DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS by K.L. McKee and A.H. Baldwin . . . . .

331

Chapter 14. MINING by J.A. Cooke

. . . . . . . . . . . . . . .

365

Chapter 15. DISTURBANCE ASSOCIATED WITH MILITARY EXERCISES by S. Demarais, D.J. Tazik, P.J. Guertin and E.E. Jorgensen . . . . . . . . . . . . . . .

385

Chapter 16. DISTURBANCE IN URBAN ECOSYSTEMS by H. Sukopp and U. Starﬁnger . . . . . .

397

Chapter 17. DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS by C.M. D’Antonio, T.L. Dudley and M. Mack . . . . . . . . . . . . . . . . . .

413

Chapter 18. DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE by D. Binkley . . . . . . . . . . . . . . .

453

Chapter 19. ANTHROPOGENIC DISTURBANCE AND TROPICAL FORESTRY: IMPLICATIONS FOR SUSTAINABLE MANAGEMENT by G.S. Hartshorn and J.L. Whitmore . . .

467

Chapter 20. SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA by C.M. Ghersa and R.J.C. Le´on

. . . . .

487

Chapter 21. PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND by R.E. Sojka . . . . . . . . . . . . . . .

503

Chapter 22. SOIL MICROORGANISMS by M.F. Allen, E.B. Allen, T.A. Zink, S. Harney, L.C. Yoshida, C. Sig¨uenza, F. Edwards, C. Hinkson, M. Rillig, D. Bainbridge, C. Doljanin and R. MacAller . . . . . . .

521

Chapter 23. RESPONSES OF CARBON AND NITROGEN CYCLES TO DISTURBANCE IN FORESTS AND RANGELANDS by D.W. Johnson and B. Schultz . . . . . .

545

223

253

271

287

xi

xii Chapter 24. DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS by I.K. Bradbury . . . . . . . . . . . . . . Chapter 25. PATTERNS AND PROCESSES IN PRIMARY SUCCESSION by L.R. Walker . . . . . . . . . . . . . . Chapter 26. PLANT INTERACTIONS DURING SECONDARY SUCCESSION by S.D. Wilson . . . . . . . . . . . . . . Chapter 27. THE RESPONSE OF ANIMALS TO DISTURBANCE AND THEIR ROLES IN PATCH GENERATION by M.R. Willig and M.A. McGinley . . . . Chapter 28. HOW HUMANS RESPOND TO NATURAL OR ANTHROPOGENIC DISTURBANCE by C.J. Barrow . . . . . . . . . . . . . . .

CONTENTS

571

585

611

Chapter 31. PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS by S.T.A. Pickett, J. Wu and M.L. Cadenasso

707

Chapter 32. ECONOMIC GROWTH, HUMAN DISTURBANCE TO ECOLOGICAL SYSTEMS, AND SUSTAINABILITY by M. Giampietro . . . . . . . . . . . . .

723

Chapter 33. DISTURBANCE IN TERRESTRIAL ECOSYSTEMS: SALIENT THEMES, SYNTHESIS, AND FUTURE DIRECTIONS by M.R. Willig and L.R. Walker . . . . . .

747

GLOSSARY

769

. . . . . . . . . . . . . . . . . . . . . .

633 SYSTEMATIC LIST OF GENERA 659

AUTHOR INDEX

. . . . . . . . . .

773

. . . . . . . . . . . . . . . . . . .

777

SYSTEMATIC INDEX

Chapter 29. RESTORATION OF DISTURBED ECOSYSTEMS by R.J. Hobbs . . . . . . . . . . . . . . .

673

Chapter 30. ENVIRONMENTAL POLICIES AS INCENTIVES AND DISINCENTIVES TO LAND DISTURBANCE by G.E. Eckert and C.R. Carroll . . . . . .

689

GENERAL INDEX

. . . . . . . . . . . . . . . .

821

. . . . . . . . . . . . . . . . . .

833

Chapter 1

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES Lawrence R. WALKER and Michael R. WILLIG

is now 5.8×109 and is projected to reach 10–12×109 by the year 2040. What are the consequences of such growth? What is the carrying capacity of the earth (Cohen, 1995)? Can human intelligence and technology prevent or even postpone a global collapse? Estimates of the ecological footprint (a concept that calculates how much arable land is needed to sustain a given level of energy consumption per member of a human population; Wackernagel and Rees, 1996) of those countries with the highest standards of living already are 15 times greater in area than the geographical space they occupy. Clearly, the world does not have the resources to sustain the entire human population at a standard of living similar to that in the more afﬂuent nations of the world. Giampietro (Chapter 32, this volume) explores ways in which wise resource management and curtailment of resource abuse can improve the future prospects of humans and the biosphere.

WHY STUDY DISTURBANCE?

Dramatic, large-scale natural disturbances (e.g., volcanic eruptions, ﬁres, hurricanes, ﬂoods) are important to understand because they destroy property, cause human injury, and disrupt emotional lives. Human interference with natural disturbances (e.g., ﬁre suppression) may actually make them more destructive (e.g., larger, hotter ﬁres: Bond and van Wilgen, 1996). Disturbances are also important to all living organisms because they have beneﬁcial effects such as nutrient recycling, resetting of successional pathways, and maintenance of species diversity (Luken, 1990). The exponential increase in human population density guarantees that more people are affected by natural disturbances every year. It is clear that one needs to continue efforts to predict and avoid disturbances, minimize damage, and maximize the ability of human society to restore degraded systems. Some anthropogenic disturbances are well publicized (e.g., spills of oil or toxic waste, bomb explosions). Yet the more gradual disturbances that do not receive as much attention, such as urbanization, excavation of minerals, soil erosion as a result of agriculture, or logging of forests, may have far greater consequences. In fact, anthropogenic disturbances are ubiquitous and all ecosystems of the world are disturbed at least partially by human activities. Both natural and anthropogenic disturbances clearly impact the entire earth. Understanding how to live with or mitigate natural disturbances, and moderate the consequences of human actions, is imperative (Thomas, 1956; Botkin et al., 1989). The consequences of increased human population represent the ultimate disturbance. Humans currently consume or utilize 40% of the earth’s primary production (Vitousek et al., 1986). The human population

PERSPECTIVES ON DISTURBANCE

Disturbances have been the subject of many myths and legends. Gods have been associated with disturbances such as volcanoes (Ixtocewatl and Pococatepetl in Mexico; Vulcan in ancient Rome; Pele in Hawaii), windstorms (Luquillo in Puerto Rico; Hurakan in Mayan culture), ﬂoods (Janaina in Brazil; Poseidon in ancient Greece), and ﬁre (Loki in Norse mythology; Prometheus in ancient Greece). The biblical Noah dealt with a ﬂood, and Moses’ enemies were subjected to a herbivore (locust) outbreak. Disturbances have directly altered human history. Volcanoes have destroyed cities (e.g., Pompeii in Italy; St. Pierre in Martinique) and altered world climates (Krakatau in Indonesia) (Sheets and Grayson, 1979; 1

2

Simkin and Fiske, 1983). Hurricanes have repeatedly damaged buildings and biota (e.g., Hurricane Hugo in the Caribbean and the eastern United States: B´enitoEspinal and B´enito-Espinal, 1991; Finkl and Pilkey, 1991; Walker et al., 1991, 1996). Fertile soils along river ﬂoodplains (e.g., the Nile, Tigris, or Euphrates) have nurtured civilizations, but often at the cost of extensive losses of lives and property (Ofﬁcer and Page, 1993). Famous ﬁres have altered the histories of cities such as Chicago, Rome and San Francisco, and the vegetation of entire continents (Komarek, 1983). Biotic disturbances are perhaps most damaging. The Black Death killed one-third of all people in medieval Europe, and many Native Americans died from diseases such as smallpox and malaria introduced by Europeans (cf. Crosby, 1986; Ofﬁcer and Page, 1993). Cultural and environmental concerns traditionally have been shaped by the interplay between resource availability and the local disturbance regime. Degradation of land caused by erosion and deforestation was noted by Greek and Roman writers, and Confucianism in China addressed environmental concerns (Barrow, 1991). Humans typically have responded to natural disturbances by management (use of ﬁre by many native cultures), exploitation (use of early-successional plants for food), or avoidance (minimal use of deserts, lava ﬁelds, and glacial valleys). Attitudes toward natural resources can evolve from exploitation to conservation when human population densities reach local carrying capacities. However, the demise of some societies [e.g., the Maya in Central America, the Hohokam in Arizona (U.S.A.), and the Assyrians in Mesopotamia] has been attributed in part to the collapse of the local resource base from over-exploitation (Thomas, 1956). The remarkable ability of humans to accommodate to naturally or anthropogenically caused environmental change (or to migrate out of disturbed areas – as with the Dust Bowl in Oklahoma, U.S.A.: Worster, 1979) suggests that most disturbances modify but do not destroy cultures. Most landscapes are now the product of a long history of human land use (e.g., the Mediterranean basin: Rundel, Chapter 10, this volume). For the last 100–200 years, Western cultures have been systematically recording observations about various natural disturbances (e.g., volcanoes: Whittaker et al., 1989; glaciers: Chapin et al., 1994) and anthropogenic disturbances (e.g., changes in levels of atmospheric carbon dioxide: Vitousek, 1994) and ecosystem responses to disturbance (e.g., succession:

Lawrence R. WALKER and Michael R. WILLIG

Clements, 1928). Such long-term observations allow an examination of disturbance on various time scales with the partitioning of short-term ﬂuctuations from longerterm cycles (Magnuson, 1990). They also facilitate the distinction of human impacts from natural ﬂuctuations. Recognition of the role of humans in global warming or acid rain, and the growing impacts of mining, agriculture, and urbanization have increased environmental awareness in recent decades. This awareness has fostered the growth of environmental politics (e.g., the Green Parties in Europe), entrepreneurism (e.g., the purchase of natural areas by private agencies such as the Nature Conservancy operating from the United States), and cooperation at the local level (restoration activities), the regional level (credits to companies that reduce pollution), and the global level (relief of national debt in exchange for establishment of nature reserves). Interactions of culture and disturbance are further discussed in this volume by Ghersa and Leon (Chapter 20), Barrow (Chapter 28), Hobbs (Chapter 29), Eckert and Carroll (Chapter 30) and Giampietro (Chapter 32).

DEFINITIONS OF DISTURBANCE

As the literature on disturbance ecology has proliferated in the last two decades, so too has the lexicon. Nonetheless, maturation of the science requires a precise use of terminology along with straightforward clariﬁcation when terms are used in different ways. At the same time, terms should be sufﬁciently general so that they are useful to an appreciable segment of the practitioners in the discipline. On occasion, growth of a discipline can be stymied signiﬁcantly by vague or illdeﬁned terminology, in part because synthesis requires incisive understanding and in part because confusion over terminology can lead to division among practitioners who disagree about deﬁnitions. Such semantic differences can give the impression of disagreement over substantive or conceptual issues, lead to heated or senseless debate, and delay the maturation of a scientiﬁc discipline. We do not attempt to resolve such semantic and conceptual differences here. Indeed, authors contributing to this volume were given broad latitude in the use of terms so as to engender individual creativity. Nonetheless, we follow White and Pickett (1985) and provide an introduction to widely accepted meanings of selected terms in the lexicon of disturbance ecology,

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES

so that the general reader will have an appreciation of the scope of the discipline, and specialists will be motivated to provide more detailed deﬁnitions or alternate terminology as appropriate (see Pickett et al., Chapter 31, this volume). A disturbance is a relatively discrete event in time and space that alters the structure of populations, communities, and ecosystems. It can do so by altering the density, the biomass, or the spatial distribution of the biota, by affecting the availability and distribution of resources and substrate, or by otherwise altering the physical environment. It often results in the creation of patches and the modiﬁcation of spatial heterogeneity. Disturbance is a relative term that requires explicit delineation of the system of concern, including the spatial and temporal scale of the components of interest. The cause of a disturbance may be thought of as the agent or entity initiating the changes in the structure of the ecological system of interest. For example, highspeed winds are agents of disturbance for hurricanes. If the cause originates outside the system of interest, as is the situation for hurricanes, the disturbance is considered to be exogenous, whereas if the cause of the disturbance originates inside the system of interest, as when a tree-fall results from natural senescence, the disturbance is considered to be endogenous. Clearly, deﬁnition of the system of interest is integral to such considerations, and a clear distinction is not always possible. The likelihood of an exogenous disturbance may be affected by the state of the system of interest and characteristics of endogenous disturbances may be affected by characteristics of previous exogenous disturbances. Indeed, the dichotomy between purely endogenous and exogenous disturbances might more appropriately be considered as a continuum of intermediate possibilities. Disturbances are most often characterized by the central tendency, variability, and distribution of three attributes: frequency, extent, and magnitude. Frequency measures the number of events per unit of time or the probability that an event will occur. Extent is the actual physical area affected by a disturbance. It can be estimated from the area of a single event (e.g., a tree-fall), or from the sum of the areas affected by equivalent events over a particular time period (e.g., gap area created by all tree-falls in a year). Extent is often reported as the proportion of an entire landscape in which a particular disturbance occurred

3

in a given time period. Magnitude includes two interrelated attributes: intensity and severity. Intensity is the physical force of an event (e.g., wind-speed for hurricanes), whereas the impact on or consequences to the system of interest is the severity (e.g., the biomass of trees that were killed by passage of a hurricane). Intensity and severity are usually correlated, and the terms often are used interchangeably, at least in part, because the physical forces of many disturbances, especially those generated by the biota (e.g., treefalls, rodent mounds, insect outbreaks) are difﬁcult to quantify. Clearly, severity reﬂects the response of the biota to the disturbance and may not be fully documented until a considerable time has elapsed since the disturbance event impinged on the system of interest. Most systems are simultaneously subjected to a number of disturbances (e.g., hurricanes, landslides, tree-falls, herbivory, droughts, and human activities all affect the structure and function of Caribbean forests). The sum of all disturbances at a particular place and time is termed the disturbance regime. The different disturbance events enhance or diminish the frequency, extent, or magnitude of other disturbances. Such interactions are considered synergisms, and are important considerations to address in understanding disturbance and recovery in ecological systems.

TYPES OF DISTURBANCE

Because virtually every habitat experiences some level of disturbance, no book can easily cover the entire topic. This book focuses on disturbances that physically impact the ground. It does not address atmospheric or aquatic disturbances. Primarily natural disturbances (Chapters 2–13) can be categorized by the four classical elements: earth, air, water, and ﬁre (Table 1.1). Disturbances linked to the earth are independent of all causal factors other than tectonic forces (del Moral and Grishin, Chapter 5, this volume). Disturbances involving air, water, and ﬁre are primarily driven by an interplay of climatic, topographic, and soil factors. In addition, biotic variables inﬂuence ﬁre and are represented by both non-human disturbances (e.g., herbivory) and human disturbances (Table 1.1). Disturbances often trigger other disturbances, so that there is an interlacing web of disturbance interactions (for a detailed example, see Fig. 33.2 below). For instance, volcanoes can trigger earthquakes, earthquakes

4

Lawrence R. WALKER and Michael R. WILLIG

Table 1.1 Examples of some of the major types of disturbance of the earth 1 Element

Primary disturbance 2

Earth (tectonic)

earthquake (1) erosion (>50) volcano (1)

Air

hurricane (15) tornado (<1) tree-fall (nd)

Water

drought (30) ﬂood (15) glacier (10)

Fire

ﬁre (>50)

Biota – non-human

herbivory (nd) invasion (nd) other animal activity 3 (nd)

Biota – human

agriculture (45) forestry (10) mineral extraction (1) military activity 4 (1–40) transportation 5 (5) urban (3)

1

Data from many sources; nd = no data available. Approximate percent of earth’s terrestrial surface regularly affected by each disturbance is in parentheses. 3 Includes building, excavating, waste products, movement, death, diseases, parasites. 4 U.S.A., 1%; Vietnam, 40%. 5 Includes motorized and non-motorized transportation. 2

or hurricanes can trigger landslides, hurricanes or landslides can induce ﬂooding, and ﬂooding can cause landslides. These interactions may augment, diminish, or neutralize the interacting disturbances. Anthropogenic disturbances are, of course, always interacting with natural disturbances (e.g., road-building can trigger a landslide). A hierarchical view of disturbance types (cf. O’Neill et al., 1986; Pickett et al., 1987) may be most useful in examining disturbance interactions, and in making spatial and temporal scales explicit for each disturbance under consideration. When the common types of disturbance of the world (from Table 1.1) are compared by frequency, extent, and severity using a subjective ranking procedure (1, least; 5, most), several patterns emerge (Fig. 1.1). Primarily anthropogenic disturbances are usually greater in extent (mean score = 3.2) than

Fig. 1.1. The frequency, spatial extent, and severity of 19 types of disturbance throughout the world based on their subjectively ranked scores from 1 (least) to 5 (most). Intensity and severity scores were highly correlated and thus are represented on a single axis. Disturbances are: AG, agriculture; AN, animal activities; DR, drought; EA, earthquakes; ER, erosion; FI, ﬁre; FL, ﬂooding; FO, forestry; GL, glaciers; HE, herbivory; HU, hurricanes; IN, invasions; MI, mining; ML, military; TF, tree falls; TO, tornadoes; TR, transportation; UR, urban; VO, volcanoes. Anthropogenic disturbances are shaded. Uncircled letters occupy the same location as adjacent circles (VO, GL; EA, TO; IN, AG).

natural disturbances (mean score = 2.1), presumably because of the cosmopolitan distribution of humans. Anthropogenic disturbances are also slightly more severe (mean score = 3.8) than natural disturbances (mean score = 3.0), but similar in frequency (mean scores 3.1 and 2.8, respectively). Of the ﬁve most severe disturbance types (score = 5), natural disturbances (glaciers and volcanoes) were less extensive and frequent than anthropogenic disturbances (mining, transportation, urban development). Transportation was rated uniquely high in both extent and severity. Other outliers were herbivory, tree-falls, and animal activities, all of which received very low scores for severity, but high scores for frequency. Most important, perhaps, is the broad range of extent, severity, and frequency among the disturbance types, particularly those representing natural disturbances. At relatively large spatial scales (~104 –1010 m2 ) and long temporal scales (~102 –104 yr), many areas of the earth are dominated by only one or a few major disturbance types. Inside the front cover of this book, we have mapped areas where disturbances related to

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES

earth, air, water, and ﬁre predominate on terrestrial surfaces of the earth. Volcanoes and earthquakes result from plate tectonics (earth element) and predominate around the rim of the Paciﬁc Ocean and in central Asia. Hurricanes (air element) develop in the tropics, but occasionally reach latitudes >45º N or S. Tornadoes reach further inland than hurricanes. Less severe windstorms are nearly ubiquitous at smaller spatial scales and were not included in the map. Floods or ice (excess of the water element) are important disturbances along river corridors and in boreal and polar regions. Drought (deﬁciency of the water element) is primarily a factor in mid-latitude, hot deserts, but also in northeastern Brazil (Mares et al., 1985). Droughts and ﬂoods are dictated largely by ocean currents, global wind patterns, and regional topography, although human activities often inﬂuence both droughts (e.g., desertiﬁcation) and ﬂooding (river channelization). Fire is the most ubiquitous type of terrestrial disturbance after human urban and agricultural activities (Bond and van Wilgen, 1996). It is important in tundra, coniferous forests, temperate grasslands and shrublands, and tropical grasslands and savannas, although only the most ﬂammable biomes (coniferous forests and Mediterranean-climate shrublands) are shown. Biotic disturbances can be considered a ﬁfth category of disturbance. Non-human biotic disturbances include plant and animal invasions, herbivory, and other animal activities (e.g., excavating, building, movement, waste products, disease, and parasitism). These activities are too ubiquitous and small in scale to map globally. In contrast, anthropogenic disturbances, equally ubiquitous but occurring at larger spatial scales,

5

can more readily be mapped globally (Fig. 1.2). There is a strong similarity between the distributions of human population (Fig. 1.2A) and common human disturbances (Figs. 1.2B, 1.2C, 1.2D). Current anthropogenic disturbances reﬂect human land-use patterns that are a consequence of historical settlements based primarily on the presence of soils suitable for agriculture (Fig. 1.2B), and appropriate waterways or land routes for transportation. More recent urbanization reﬂects primarily transportation centers (Fig. 1.2C) that have excellent access to power sources or to agricultural products (cf. Cronon, 1991). Many humans (45%) now live in or near cities, and this trend is accelerating. Nevertheless, some human activities such as mineral extraction (Fig. 1.2D) and military installations may actually promote low human population densities, but still represent severe disturbance (e.g., northern Alaska, northern Venezuela, eastern Saudi Arabia). Inside the back cover of this book, we have mapped all human inﬂuences together, using four hemeroby classes (see Sukopp and Starﬁnger, Chapter 16, this volume) representing degrees of human inﬂuence: (1) minimal: mountains, tundra, undeveloped forest; (2) moderate: low human population densities, some agriculture; (3) major: moderate human population densities, intense agriculture (e.g., deep plowing, clearcutting, biocides); and (4) maximal: high urban population densities, sealed or poisoned land surfaces. This measure of combined inﬂuences of humans emphasizes that most damage occurs where population densities are high. Agriculture and resource extraction, although often locally severe, do not alter the environment as much as pavement and urban buildings.

6

Lawrence R. WALKER and Michael R. WILLIG

Fig. 1.2A. Global distribution of four aspects of anthropogenic disturbance. A. Human population distribution. Grey levels indicate different population densities: <2 km−2 ; 2–20 km−2 ; 20–100 km−2 ; >100 km−2 . Various sources were used, including: Oxford World Atlas (1973) and The Times Atlas of the World (1990, 1995).

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES

Fig. 1.2A (continued).

7

8

Lawrence R. WALKER and Michael R. WILLIG

Fig. 1.2B. Agriculture excluding forestry. Grey levels indicate relative intensity:

sparse;

low;

moderate;

high.

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES

Fig. 1.2B (continued).

9

10

Lawrence R. WALKER and Michael R. WILLIG

Fig. 1.2C. Surface transportation (roads and railroads). Grey levels indicate relative intensity:

sparse;

low;

moderate;

high.

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES

Fig. 1.2C (continued).

11

12

Lawrence R. WALKER and Michael R. WILLIG

Fig. 1.2D. Mineral resource extraction (including oil, gas and coal). Grey levels indicate relative intensity: high.

sparse;

low;

moderate;

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES

Fig. 1.2D (continued).

13

14

Lawrence R. WALKER and Michael R. WILLIG

Table 1.2 Distribution of topics discussed in each chapter. Parentheses indicate minor topics Chapter

Element 1

Geographic region 2

Ecoregion 3

Trophic level 4

Theme 5

2

1,2,3

1,2,(3),(4),5,6

1,5

1,(2),3

1,3,(4),7,8

3

1,2,3,4,5

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

1,2,4,6

1,2,(3)

2,(4),5,6,7,9,12

4

1,2,3,5

(1),3,(5),6,7

(2),5,8

1,(2),3

1,5,6,10

5

1,(2),3

3,4,(5),6,8

4,5

1,2,3

1,2,3,8,10,12

6

2,4,5

3,5,6

5

1,2

2,3,4,7,(10)

7

2,5

3,4,5,6,7,8

5,6

1,2,3

1,2,3,(4),5,8,10,13

8

2,(4)

1,3,4,6,7,8

5,6

1

1,2,3,5,6,8,14

9

5

4,6

5

1,2

1,2,3,5,6,9,11

10

4,5

1,4,5,6,7

1,2,4,5,8

1

1,10,12

11

4,5

1,6,7

(1),2,3,(5)

1,2,(3)

1,5,6,7,9,14

12

2,3,4,5

1,3,4,5,6,7

1

1,2,(3)

1,2,3,(4),14

13

(2),3,4,5

3,6,8

6

1,2

1,2,3,4,5,6,(7),9,10,(14)

14

1,5

1,(3),4,5,6,(7)

1,2,3,5

1

1,3,5,6,13

15

5

4,5,6

1,2,4,5,6

1,2

1,10,13

16

5

(3),5,6

5,7

1,2

(2),3,(5),10,12

17

(1),(2),(3),4,5

1,4,5,6,8

1,2,3,4,5,6,8

1,2

1,3,4,5,(10),12

18

1,2,4,5

4,5,6

5,8

1,2

2,3,5,6,(10)

19

1,2,3,4,5

1,3,4,6,7,8

5

1,2

1,2,3,9,10,12,13

20

5

7

2,8

1,(3)

2,3,4,5,7,(8),10,12,14

21

1,3,5

6

(6),8

1,2

1,5

22

1,4,5

5,6

1,4,5

1,2,3

3,5,13,14

23

5

1,5,6

2,4,5

1,2,3

1,3,4,5,(6)

24

1,4,5

1,4,5,6

2,4,5

1,2

1,5,6

25

1,2,3,5

1,2,3,4,5,6,8

5

1,2,3

1,3,4,5,6

26

5

(4),6

2,5,(6),8

1

2,3,4,5,6

27

2,4,5

4,6,8

2,4,5,8

1,2,3

1,2,3,4,5,6,10,14

28

4,5

4,5,6,7

(2),(4),(5)

2

1,(2),5,7,9,13,14

29

5

4,6,7

2,4,5,(6),8

1,2,3

1,5,7,9,11,12,13,14

30

5

1,3,6,7,8

(1),2,5,6,8

1,2,3

3,7,9,10,13,14

31

1,2,3,4,5

6,7

5

2

2,(3),7,(13),14

32

5

1,3,4,5,6

8

1,2

5,6,8,14

1

Element: 1, earth; 2, air; 3, water; 4, ﬁre; 5, biota. Geographic region: 1, Africa; 2, Antarctica; 3, Asia; 4, Australasia (Australia, New Zealand, Micronesia); 5, Europe; 6, North America; 7, South America; 8, Islands. 3 Ecoregion: 1, desert; 2, grassland; 3, savanna; 4, shrubland; 5, forest; 6, wetland; 7, urban; 8, agroecosystem. 4 Trophic level: 1, producer; 2, consumer; 3, decomposer. 5 Theme: 1, interactions; 2, spatial heterogeneity; 3, succession; 4, competition; 5, nutrient cycling; 6, productivity; 7, stability and resilience; 8, predictability; 9, thresholds; 10, biodiversity; 11, functional redundancy; 12, invasive species; 13, restoration and management; 14, modeling. 2

AN INTRODUCTION TO TERRESTRIAL DISTURBANCES DISTURBANCE THEMES

The chapters in this volume approach the topic of disturbance from many perspectives (Table 1.2). Each chapter addresses how at least one of the four basic elements or ethers (earth, air, water, ﬁre) and the biota may be an agent of disturbance, and a few chapters address all of them. The most frequently covered type of disturbance is biotic (particularly human). The most frequently described geographical region is North America, but all regions of the world are discussed (more than simply a reference or brief mention) in the following rank order: North America Australasia = Europe > Africa = Asia = South America > islands Antarctica. This representation probably reﬂects both author bias and available literature, although all regions except Antarctica are discussed in at least ten chapters. Ecological regions were discussed in the order: forests grasslands > shrublands = agroecosystems > deserts = wetlands savannas > urban areas, again suggesting the distribution of available literature (and humans), and despite the global importance of urbanization. Most chapters address effects of disturbance on primary producers and consumers, but a substantial fraction also consider decomposers. Fourteen themes emerge in the following order: nutrient cycling > interactions = succession spatial heterogeneity > productivity = biodiversity > competition = modeling > stability and resilience > thresholds = invasive species = restoration and management > predictability functional redundancy. This order suggests that disturbances often interact and that there are intimate links between disturbance and nutrient cycling, succession, spatial heterogeneity, productivity, and biodiversity. In the last chapter of the volume, we examine the lessons learned from earlier chapters in this volume about the relationships between disturbance and these important themes. Ecologists have made great strides in understanding the role of disturbance in shaping natural systems. Successional responses and competitive interactions among species have generated particularly large numbers of papers. Other responses to disturbance (notably belowground processes) have received very little attention. Land managers have developed a broad base of knowledge about practical issues relating to intentional human disturbances such as agriculture. However, neither ecologists nor land managers have developed a

15

robust set of predictions about the consequences of disturbances. Much more integration of management and theory is needed in order to address the environmental challenges which humans face. Especially important to understand are the consequences of irregular natural disturbances such as hurricanes or volcanoes, the recent and overwhelming human impacts such as erosion or clear-cutting, and the interactions among them. This global compendium of examples of disturbed ground offers a sampling of the types of data that are available and some preliminary generalizations and conceptual models. We hope this book will stimulate more longterm monitoring of disturbed ground, experiments that address the mechanisms behind biotic responses to disturbance, and studies that compare responses within and among various types of disturbances (and ideally across gradients of disturbance severity). Such types of data are needed to provide the basis for predictions about disturbance. REFERENCES Barrow, C.J., 1991. Land Degradation. Cambridge University Press, Cambridge, 295 pp. B´enito-Espinal, D.B. and B´enito-Espinal, E. (Editors), 1991. L’Ouragan Hugo: genese, incidences g´eographiques et e´ cologiques sur la Guadeloupe. Co-editors: Parc National de la Guadeloupe, D´el´egation R´egionale a l’Action Culturelle, and Agence Guadeloup´eene de l’Environnement du Tourisme etdes Loisirs. Imprimerie D´esormeaux, Fort-de-France, Martinique. Bond, W.J. and van Wilgen, B.W., 1996. Fire and Plants. Chapman and Hall, London, 263 pp. Botkin, D.B., Caswell, M.F., Estes, J.E. and Orio, A.A. (Editors), 1989. Changing the Global Environment: Perspectives on Human Involvement. Academic Press, New York, 459 pp. Chapin III, F.S., Walker, L.R., Fastie, C.L. and Sharman, L.C., 1994. Mechanisms of primary succession following deglaciation at Glacier Bay, Alaska. Ecol. Monogr., 64: 149–175. Clements, F.E., 1928. Plant Succession and Indicators. H.W. Wilson, New York, 453 pp. Cohen, J.E., 1995. How Many People Can the Earth Support? Norton, New York, 532 pp. Cronon, W., 1991. Nature’s Metropolis: Chicago and the Great West. Norton, New York, 530 pp. Crosby, A.W., 1986. Ecological Imperialism – The Biological Expansion of Europe, 900–1900. Cambridge University Press, Cambridge, 368 pp. Finkl, C.W. and Pilkey, O.H. (Editors), 1991. Impacts of Hurricane Hugo: September 10–22, 1989. J. Coastal Res., 8: 356 pp. Special issue. Komarek, E.V., 1983. Fire as an anthropogenic factor in vegetation ecology. In: W. Holzner, M.J.A. Werger and I. Ikusima (Editors), Man’s Impact on Vegetation. Dr W. Junk, The Hague, pp. 77–82. Luken, J.O., 1990. Directing Ecological Succession. Chapman and Hall, London, 251 pp.

16 Magnuson, J.J., 1990. Long-term ecological research and the invisible present. BioScience, 40: 495–501. Mares, M.A., Willig, M.R. and Lacher Jr., T.E., 1985. The Brazilian caatinga in South American zoogeography: tropical mammals in a dry region. J. Biogeogr., 12: 57–69. Ofﬁcer, C. and Page, J., 1993. Tales of the Earth. Oxford University Press, Oxford, 226 pp. O’Neill, R.V., DeAngelis, D.L., Waide, J.B. and Allen, T.F.H., 1986. A Hierarchical Concept of Ecosystems. Princeton University Press, Princeton, NJ, 253 pp. Oxford World Atlas, 1973. Saul B. Cohen (Geographical Editor), Oxford University Press, Oxford, 190 pp. Pickett, S.T.A., Collins, S.L. and Armesto, J.J., 1987. Models, mechanisms, and pathways of succession. Bot. Rev., 53: 335–371. Sheets, P.D. and Grayson, D.K., 1979. Volcanic Activity and Human Ecology. Academic Press, New York, 644 pp. Simkin, T. and Fiske, R.S., 1983. Krakatau 1883: The Volcanic Eruption and its Effects. Smithsonian Institution Press, Washington, DC, 464 pp. Thomas, W.L., 1956. Man’s Role in Changing the Face of the Earth. University of Chicago Press, Chicago, IL, 1193 pp. Times Atlas of the World, 1990. Times Books, London, 395 pp. Times Atlas of the World, 1995. Times Books, London, 374 pp.

Lawrence R. WALKER and Michael R. WILLIG Vitousek, P.M., 1994. Beyond global warming: ecology and global change. Ecology, 75: 1861–1876. Vitousek, P.M., Ehrlich, P.R., Ehrlich, A.H. and Matson, P.A., 1986. Human appropriation of the products of photosynthesis. BioScience, 36: 368–373. Wackernagel, M. and Rees, W., 1996. Our Ecological Footprint. New Society Publishers, Philadelphia, PA, 160 pp. Walker, L.R., Brokaw, N.V.L., Lodge, D.J. and Waide, R.B. (Editors), 1991. Ecosystem, plant and animal responses to hurricanes in the Caribbean. Biotropica, 23: 313–521. Walker, L.R., Silver, W.L., Willig, M.R. and Zimmerman, J.K. (Editors), 1996. Long term responses of Caribbean ecosystems to disturbance. Biotropica, 28: 414–614. White, P.S. and Pickett, S.T.A., 1985. Natural disturbance and patch dynamics: an introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Orlando, Florida, pp. 3–13. Whittaker, R.J., Bush, M.B. and Richards, K., 1989. Plant recolonization and vegetation succession on the Krakatau Islands, Indonesia. Ecol. Monogr., 59: 59–123. Worster, D.E., 1979. Dustbowl: the Southern Plains in the 1930s. Oxford University Press, Oxford, 277 pp.

Chapter 2

DISTURBANCE REGIMES AND ECOSYSTEM RESPONSE ON RECENTLY-DEGLACIATED SUBSTRATES John A. MATTHEWS

Elven, 1978a, 1980; Matthews, 1978, 1979a; Paternoster, 1984). More recent developments include the ordination of communities (e.g., Matthews, 1979b–d; Matthews and Whittaker, 1987; Whittaker, 1989, 1991, 1993; Crouch, 1993; Vetaas, 1994) and autecological studies (St¨ocklin, 1990; St¨ocklin and B¨aumler, 1996). Studies outside of North America and Europe have been relatively few in number and, with the notable exception of work associated with the ice-free Antarctic landscapes (e.g., Smith, 1993; Walton, 1993; WynnWilliams, 1993; Lyons et al., 1997) and in New Zealand (e.g., Wardle, 1980, 1991; Sommerville et al., 1982; Burrows, 1990), generally less detailed. However, some investigations have also taken place in South America (Lawrence and Lawrence, 1959; Heusser, 1960, 1964; Rabassa et al., 1981; Veblen et al., 1989; Jordan, 1991), Africa (Coe, 1967; Spence, 1989; Mahaney, 1990), Irian Jaya (Hope, 1976) and Asia (e.g., Turmanina and Volodina, 1978; Solomina, 1989). Although recently-deglaciated terrain originates as a result of major disturbance (nudation) and is subjected to a wide range of disturbances, both the disturbances and their ecological effects are, paradoxically, poorly understood. Rather, emphasis has been placed on primary succession as ecosystem development following nudation. With few exceptions, most notably Oliver et al. (1985) and Whittaker (1991), development of these ecosystems has been viewed as a largely biological, deterministic process in a static physical environment, which is relatively immune from subsequent disturbance. This traditional, oversimpliﬁed view was challenged by Matthews (1992), who enlarged upon the concept of recently-deglaciated terrain as a geoecological landscape in which biological processes are coupled with abiotic processes of the physical

INTRODUCTION

Recently-deglaciated terrain holds a special place in the history of ecology, largely as a result of the classical studies on primary succession and soil development that have been carried out in these landscapes, in North America (e.g., Cooper, 1923a–c, 1939), the Alps (L¨udi, 1921, 1945, 1958; Braun-Blanquet and Jenny, 1926; Friedel, 1934, 1937, 1938; Negri, 1934) and Scandinavia (Fægri, 1933). These were some of the ﬁrst detailed investigations of primary succession and soil development, and they took advantage of the spatial chronosequence of ecosystems that exists in front of a retreating glacier. Over the last 40 years, recently-deglaciated landscapes have continued to inspire important empirical and theoretical contributions to understanding ecological succession. Cooper’s early work at Glacier Bay, Alaska, has been followed by detailed research on the vegetation itself (including the long-term monitoring of change), on the soil changes that accompany vegetation change, and on the mechanisms of change (e.g., Crocker and Major, 1955; Lawrence, 1958, 1979; Mirsky, 1966; Lawrence et al., 1967; Reiners et al., 1971; Bormann and Sidle, 1990; Chapin et al., 1994; Fastie, 1995). Similar investigations have been made at other glaciers in North America (e.g., Scott, 1974; Birks, 1980; Sondheim and Standish, 1983; Fitter and Parsons, 1987; Blundon and Dale, 1990; Blundon et al., 1993; Helm and Allen, 1995). In Europe, studies of plant communities of recently-deglaciated substrates have generally involved phytosociological approaches, which have traditionally emphasized the classiﬁcation and mapping of communities as a basis for understanding environmental relationships (e.g., Jochimsen, 1963, 1970; Persson, 1964; Richard, 1973; 17

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environment. The term ‘geoecology’ was coined by Troll (1971) with the aim of placing the study of ecosystems in the broader context of the geographical landscape (German: ‘Landschaft’). It emphasizes the spatial organization of ecosystems in the landscape, and the broader framework of environmental processes, both natural and cultural, with which biological and ecological processes interact. Thus, successional changes on deglaciated substrates can be viewed as a result of the interaction of dynamic physical processes, including disturbance, with biotic processes. In this chapter, I focus on the importance of disturbance in the recently-deglaciated landscape, with particular attention to ecological effects on ecosystems. Of necessity, the chapter relies heavily on the studies reviewed in Matthews (1992) with the addition of the rather limited amount of more recent research. However, there has been no previous comprehensive review, ecological or otherwise, of disturbance in these environments. After summarizing the global distribution and general characteristics of recently-deglaciated terrain, this chapter develops three main themes: (1) the classiﬁcation of disturbance types and their characteristics; (2) observed effects of disturbances on ecosystems, particularly the direct effects on the substrate and on plants; and (3) the role of disturbance as a driving force in primary succession.

GLOBAL DISTRIBUTION AND GENERAL CHARACTERISTICS OF DEGLACIATED SUBSTRATES

Although about 96% of Earth’s glacier ice is in Antarctica and Greenland, ice bodies – glaciers, ice caps and ice sheets – and hence deglaciated substrates occur on all continents in a remarkably wide range of polar and alpine environments. Glaciers accumulate wherever there is a surplus of snowfall over snowmelt for sufﬁcient years to allow the consolidation of snow into ice; ice wastes away, glacier margins retreat and deglaciated substrates are exposed when melting (ablation) exceeds accumulation. On a global scale, the latitudinal and altitudinal distribution of glaciers is primarily dependent on low air temperatures: thus, glaciers occur at or close to sea level in some parts of the Arctic and Antarctic, but only above about 5000 m on tropical mountains. At a continental scale, glaciers are found at increasing altitudes from oceanic coasts towards continental interiors primarily

John A. MATTHEWS

in response to decreasing snowfall, and they may be absent from many cold but dry continental interiors because of precipitation starvation. Locally, other factors are inﬂuential, particularly wind and aspect through their effects on snow-drifting and topographic shading. A zone of recently-deglaciated terrain – the glacier foreland – currently occurs in front of most glaciers world wide. This has been produced by glacier retreat from the ‘Little Ice Age’ glacial maximum attained at various times during the last few centuries. In southern Norway, for example, most glaciers attained their ‘Little Ice Age’ maxima around the middle of the eighteenth century (Matthews, 1991; Bickerton and Matthews, 1993) (Fig. 2.1); in the Alps, some reached their maxima around AD 1600, others in the middle of the nineteenth century (Grove, 1988). Successively older terrain occurs with increasing distance from the glacier front, and a range of techniques are of potential use in dating glacier forelands with varying levels of accuracy (see Matthews, 1992). In some cases, older substrates may be available as a result of pre‘Little Ice Age’ (Neoglacial) glacier advances and retreats. The extent of the glacier foreland varies with the magnitude of glacier retreat. On the one hand, large valley glaciers in maritime climates tend to be relatively dynamic; glacier retreat rates may reach >100 m yr−1 , and the foreland may extend for several kilometres from the glacier; some small cirque glaciers in continental climates, on the other hand, are characterized by a narrow deglacierized zone <100 m wide. Glaciers that calve into lakes or other water bodies, have the greatest retreat rates and their forelands tend to be the most extensive: for example, at Glacier Bay, Alaska, retreat has totalled more than 100 km since about AD 1760 (Chapin et al., 1994). Recently-deglaciated landscapes are relatively barren (Fig. 2.2) but, as Jochimsen (1962) has aptly described them: “Das Gletschervorfeld – keine W¨uste” [the glacier foreland – not a desert]. Except on the most recently created land surfaces, close inspection reveals the presence of plant life, and the vegetation cover usually increases rapidly with increasing terrain age. Distinctive landforms, the products in particular of glacial and glacio-ﬂuvial erosional and depositional processes, introduce a variety of habitats (e.g. Sugden and John, 1976; Drewry, 1986). The substrate itself is typically a till, that is, a diamicton or unsorted mixture

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Fig. 2.1. The glacier foreland of Bøverbreen, Jotunheimen, Norway: the relatively barren, recently-deglaciated zone between the glacier and the ‘Little Ice Age’ moraine ridge dating from about AD 1750. Note the mature mid-alpine vegetation in the foreground on terrain deglaciated about 9000 years B.P., as estimated by 14 C measurements.

of textures ranging from boulders to clays that have been deposited directly by glacier ice (e.g., Boulton and Eyles, 1979; Dreimanis, 1988). This is often absent in areas of glacier erosion, where bedrock outcrops occur, or have been reworked and replaced, most notably by sorted glacio-ﬂuvial sands and gravels deposited by meltwater streams (Maizels, 1979; Maizels and Petch, 1985; Gurnell and Clark, 1987). Superimposed on the regional polar or alpine climate, and in addition to any climatic effects resulting from the local topography, are meso-climatic inﬂuences due to the proximity of a glacier. These include the likelihood of glacier winds, which are particularly strong on open terrain and where conﬁned by valleys (e.g. Hoinkes, 1954; Nickling and Brazel, 1985). In summary, the environment may be described as relatively inhospitable to plants even without any consideration of disturbance. However, as deglaciated terrain is varied and occurs in a wide range of macroclimates, conditions for plant growth and ecosystem recovery are not easy to generalize and should not be viewed as uniform.

TYPES OF DISTURBANCE

Four types of disturbance can be recognized as affecting recently-deglaciated substrates and their ecosystems: (1) direct glacial disturbance, deﬁned as disturbance caused by glacier contact; (2) glacier-dependent disturbance, caused by the proximity of a glacier; (3) glacier-conditioned disturbance, substrate disturbance conditioned by previous glacier presence; and (4) glacier-independent disturbance, which happens to occur on deglaciated terrain. Direct glacial disturbance Direct glacial disturbance is an effect of direct contact between the glacier and the ecosystem. Erosional disturbance of this type commonly accompanies a glacier advance, when ecosystems may be overridden and substrates bulldozed (Fig. 2.3). Also included is the direct deposition of debris from glacier ice, which may slide, roll, or ﬂow down the glacier

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John A. MATTHEWS

Fig. 2.2. The glacier foreland of Stegholbreen, an outlet of the Jostedalsbreen ice cap, Norway. Note the sharp trim-line between the glacier foreland and the mature, low-alpine, Betula pubescens woodland.

surface onto the proglacial area or may melt out from ice. The most important disturbance in the latter category is probably the sediment ﬂows that originate as supraglacial debris liqueﬁes in response to the melting of underlying ice (Lawson, 1981, 1982). A third, more unusual category follows glacier retreat and includes rock slides (Sigurdsson and Williams, 1991) and the slumping of lateral moraines as a result of the removal of glacial support (Fig. 2.4), and the slow melting of buried glacier ice, either in ice-cored moraines or beneath other parts of the glacier foreland (Wright, 1980; Mattson and Gardner, 1991; Krainer and Poscher, 1992; Winkler and Hagedorn, 1994). Whole ecosystems were destroyed during ‘Little Ice Age’ glacier expansion, which can be regarded as major nudation events on a global scale. Relics of these former ecosystems in the form of sub-fossil tree trunks, exhumed plant communities, and/or soil have occasionally been exposed during subsequent glacier retreat (e.g., Elven, 1978b; Smith, 1982; Bergsma et al., 1984). Smaller-scale destruction followed the minor readvances and accompanying moraine-ridge deposition that have punctuated the general retreat from

‘Little Ice Age’ limits (e.g., Bickerton and Matthews, 1993) and are currently occurring at many alpine and Scandinavian glaciers (e.g., Matthews et al., 1995; see also Fig. 2.3). Glacier-dependent disturbance Glacier proximity may result in indirect disturbance even though there is no direct contact between the glacier and the ecosystem. Glacio-climatic, glaciohydrologic and glacio-ﬂuvial disturbance may be recognized as sub-types. Glacio-climatic disturbance is induced by local climatic changes as the glacier advances and retreats. Glacier winds are the most obvious manifestation of the glacier climate but others include the lowering of air temperatures, especially during the day, and the elevation of minimum air temperatures at night (Hoinkes, 1954; Lindr¨oth, 1971). As such effects are limited to a distance of about 200 m from the snout of cirque and small valley glaciers, they affect the ecosystems when the glacier approaches to within this threshold distance. Similarly, there is a soilmoisture gradient in summer associated with glacier

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21

Fig. 2.3. An advancing glacier bulldozing the substrate at Bergsetbreen, an outlet of the Jostedalsbreen ice cap, Norway. Note the small ‘push-moraine’ ridge, the damaged vegetation, and additional disturbance from blocks of ice fallen from the near-vertical ice front.

proximity, because there is continuous groundwater recharge from the melting ice close to the glacier snout. This gradient is likely to be much steeper than the soil temperature gradient, with high soil-moisture deﬁcits occurring on well-drained sites even within 40 m of the glacier snout (Ballantyne and Matthews, 1982). Glacioﬂuvial disturbance by meltwater streams extends much further downstream and is present to a greater or lesser extent on all glacier forelands (Maizels, 1979; Odland et al., 1991). The characteristic braided-channel pattern of an outwash plain (plain sandur: extensive, unconﬁned area of glacio-ﬂuvial deposits) or a valley train (valley sandur: an area of glacio-ﬂuvial deposits conﬁned by valley sides) is a response to the high sediment load of the streams, their variable discharge and their non-cohesive banks (Miall, 1983), which ensure that the anastomosing (inter-twining) complex of shallow, shifting channels remains active almost indeﬁnitely (Fig. 2.5). Glacier-conditioned disturbance Recently-deglaciated substrates are susceptible to some

kinds of disturbances because they have been covered by glacier ice in the recent past. Such glacierconditioned disturbances represent the adjustment of the newly-deglaciated landscape to exposed, subaerial environmental conditions. They may also be termed paraglacial disturbances because they result from nonglacial processes that are directly conditioned by glaciation (cf. Church and Ryder, 1972; Matthews et al., 1998). Many kinds of disturbance are most effective immediately after deglaciation, and become less effective with the passage of time. These disturbances include mass movements, such as soliﬂuction (Rose, 1991) and debris ﬂow (Ballantyne and Benn, 1994; Ballantyne, 1995), cryogenic (frost-heave and frostsorting) processes (Ballantyne and Matthews, 1982, 1983), aeolian deﬂation and loess-like redeposition by wind (Boulton and Dent, 1974; Rose, 1991), and pervection (the down-proﬁle removal of ﬁne soil particles by soil water percolation). The newly-deglaciated substrates of proglacial areas are particularly susceptible to these disturbances, because of such characteristics as unconsolidated and meltwater-saturated sediments that dry out periodically, steep slopes, and the lack

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¨ Fig. 2.4. Large-scale disturbance of the eastern lateral moraine of the Marzellferner, Otztal Alps, Austria. To the left, sections of the moraine over 50 m in length have slipped towards the valley ﬂoor as a result of the removal of support following glacier retreat. Note also the extensive gully erosion on the proximal slope of the moraine, and the debris-covered glacier snout in the valley bottom (right).

of a signiﬁcant vegetation cover. These disturbances decline as slopes stabilize (Welch, 1970; Sharp, 1984), sediments consolidate (Boulton and Paul, 1976) or are exhausted, the glacial-meltwater source retreats, and a protective vegetation cover becomes better established. Selected examples are illustrated in Figs. 2.6 and 2.7.

by humans (e.g., agriculture, forestry, tourism and scientiﬁc activity) which is best classiﬁed as exogenous glacier-independent disturbance, although tourists and scientists are often attracted by the presence of the glacier and hence in one sense may be considered as glacier-dependent.

Glacier-independent disturbance The occurrence of some disturbances on recentlydeglaciated substrates is entirely coincidental; they occur elsewhere and there is no reason to expect a greater frequency or magnitude on glacier forelands. The wide range of such disturbances that might occur may be classiﬁed as exogenous (e.g., those avalanches and debris-ﬂows that descend onto the glacier foreland from valley-side slopes, and nivation processes associated with late-lying snowbeds) or endogenous (e.g., wind-throw disturbance on wooded forelands; Noble et al., 1984; Bormann and Sidle, 1990). The latter also includes animal infestations (e.g., Pignatelli and Bleuler, 1988), and disturbance

EFFECTS OF DISTURBANCE REGIMES ON ECOSYSTEMS

Because most ecological studies on recently-deglaciated terrain focus on succession, disturbances of all types have generally been actively avoided rather than investigated. Nevertheless, there is accumulating evidence of the diverse effects of a very wide range of disturbances on glacier-foreland ecosystems. Following the availability of evidence, emphasis in this section is given to the observed direct effects of disturbance on the substrate and on plants.

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Fig. 2.5. Glacio-ﬂuvial disturbance on the outwash plain (sandur) in front of Veobreen, Jotunheimen, Norway. Note the characteristic braided-channel of the sandur, which is being crossed by a herd of reindeer (Rangifer tarandus) (centre).

Substrate modiﬁcation Disturbance can physically remove substrate material en masse or differentially, can produce soil mixing, and can add substrate material to the soil proﬁle. All these types of substrate modiﬁcation have been observed on glacier forelands, and merge with more gradual soil changes that occur in response to the operation of continuous abiotic soil-forming processes. Soil proﬁles are shallow on recently-deglaciated terrain (Mellor, 1985; Messer, 1988) and many geomorphic processes are active. Truncation or removal of soil proﬁles, whether this is by direct glacial disturbance or other types of disturbance, is not unusual. The most effective processes in this are glacio-ﬂuvial and massmovement processes; the former are particularly active near the valley axis downstream of valley glaciers; the latter near the glacier snout because of ice-melting and paraglacial effects (see above). Differential removal of ﬁne particles is less obvious, but is ubiquitous on unvegetated areas, both close to the glacier and in eroded areas elsewhere. Deﬂation of silt and ﬁne sand by strong winds is thought to be the main

cause of the stony layer that develops rapidly on newlyexposed till surfaces in Iceland (Boulton and Dent, 1974). Surface wash by running water and frost-sorting processes may also contribute to the development of this stony lag. The amount of material removed may sometimes be estimated from exposed plant roots or from stones that remain isolated above the general level of the surrounding terrain (Rose, 1991). Subsurface removal of silt and clay down the soil proﬁle by pervection affects not only the near-surface layer but deeper layers as well. At ﬁrst, this leads to the rapid formation of a sub-surface, silt-rich layer but, after 10– 30 years, the sub-surface layer is degraded (Romans et al., 1980; Frenot et al., 1995). Substrate mixing commonly occurs due to cryoturbation on Norwegian glacier forelands. Heaving and subsequent settling as ice lenses grow and melt is one source of disturbance; additional disturbance results from the differential heave of particles of different size, which produces frost sorting. Frost heave may continue throughout the winter in frost-susceptible, especially silt-rich substrates as frost penetrates into the ground, provided there is an adequate water supply for the

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John A. MATTHEWS

Fig. 2.6. Disturbance due to slow mass movement: geliﬂuction (soliﬂuction) on the glacier foreland of Styggedalsbreen, Jotunheimen. The downslope movement (upper left to lower right) of the substrate is reﬂected in the surface microrelief, and has probably been retarded by a surﬁcial crust of mosses.

growth of ice lenses. Larger particles heave at faster rates than smaller ones, which produces boulder-cored frost boils (Harris and Matthews, 1984) where large boulders break the ground surface, and sorted patterned ground when the larger particles have migrated to the margins (gutters) around domed centres of ﬁnes (Ballantyne and Matthews, 1982, 1983). On sloping terrain, the thawing of ice-rich sediments leads to downslope movement, partly as a result of settling (frost creep) but mostly as a result of ﬂow (geliﬂuction) which occurs following the saturation of the soil and its loss of cohesion. Geliﬂuction and frost sorting are the main processes leading respectively to soliﬂuction lobes and stone stripes (Figs. 2.6 and 2.7). Eroded material tends to be deposited down-slope, where substrate burial can occur. Aeolian deposition of organic material as well as ﬁne mineral particles, removed by wind from close to the glacier or from other exposed substrates, occurs in sheltered localities further away from the glacier and in vegetated areas (Boulton and Dent, 1974; Rose, 1991). Baranowski and P˛ekala (1982) found that mean annual deposition of up to

174 g m−2 , including 4% organic material, occurred in the end-moraine zone of the glacier Werenskioldbreen in Svalbard. Gellatly (1987) described how successive intervals of deposition produce successive buried (fAh) horizons in distal parts of the glacier foreland of Classen Glacier, New Zealand. Plant damage Many processes that disturb the substrate also damage the plants; additional disturbances may damage plants without modifying the substrate. Despite its wide recognition, there have been relatively few descriptions of plant damage as a direct effect of disturbance on recently-deglaciated terrain. Strong winds have major effects on open glacier forelands where plants may be isolated or vegetation cover thin. In Iceland, where a Racomitrium canescens– Stereocaulon spp. moss–lichen heath is the ﬁrst plant community to achieve a complete ground cover, this is easily stripped off once disturbed (Boulton and Dent, 1974). Eroded turf, injuries to shrubs and

DISTURBANCE REGIMES AND ECOSYSTEM RESPONSE ON RECENTLY-DEGLACIATED SUBSTRATES

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Fig. 2.7. Substrate disturbance indicated by small-scale stone stripes produced by a combination of frost heaving and sorting, and downslope movement, on the glacier foreland of Svellnosbreen, Jotunheimen, Norway. Note lens cap for scale.

crippling of shrubs were attributed by Friedel (1936) to the destructive effects of winds in front of the Pasterze Gletscher in the Austrian Alps, where the lower altitudinal limit of Elynetum (an alpine grass– heath community) is depressed by as much as 500 m. Friedel (1936) also described the erosion of moraine crests and the deposition of loess-like ﬁne sediments up to 3 m thick in vegetated areas. Heusser (1956) recognized wind-trained and wind-sheared trees which stand above the winter snow cover near the Columbia Glacier (British Columbia, Canada) (see also, Lutz, 1930; Oke, 1987). The proximate cause of such damage may be physical abrasion by snow or mineral particles transported by wind (niveo–aeolian erosion), or physiological effects due, for example, to winter dessication and frost damage. In several studies, growth suppression has been detected in the width of tree-rings following the approach of an advancing glacier (Lawrence, 1950; Bray and Struik, 1963; Holzhauser, 1984; Villalba et al., 1990). Suppression is manifest in narrow rings, which can be attributed to glacier-dependent disturbance of the local climate of the tree. This effect should be distinguished

from possible growth stimulation if competition is reduced following the thinning-out of surrounding trees (Bray and Struik, 1963). Root breakage and exposure to desiccation, caused by mobility of the substrate, are the main destructive effects on plants of frost-heave, related cryogenic processes, and mass-movement processes. Large roots, such as tap roots, and those of woody shrubs, are likely to be more susceptible to damage by frost-heave than the ﬁbrous roots of grasses. Tap roots can be heaved out of the ground, much like relatively large soil particles. Even roots that extend below the depth of winter frost penetration, and hence become anchored in the substrate, may still be damaged by being snapped below the crown as the soil nearer the ground surface heaves (cf. Perfect et al., 1987, 1988; Goulet, 1995). Glacio-ﬂuvial disturbance is most noticeable after major ﬂood events and along channel banks (Sidle and Milner, 1989). However, more subtle effects are likely to be widespread on ﬂoodplains following the annual spring snow-melt ﬂood, and also after more frequent rainfall-induced summer ﬂooding. For example, data on seedling survival for a three-year period from the

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foreland of the Exit Glacier in Alaska showed only 30% survival in ﬂooded plots compared with 80% survival in non-ﬂooded areas. Mortality was, moreover, greatest among young seedlings less than one month old (Helm and Allen, 1995). Indirect effects on plants The indirect effects of disturbance on plants, and the effects on animals and micro-organisms, are even more poorly understood. Some of the above-mentioned substrate modiﬁcations, for example, have potential impacts on nutrient availability to plants. Flood events may provide nutrient ﬂushes from meltwater. Soil mixing by cryoturbation on recently-deglaciated terrain is likely to counteract not only the loss of ﬁnes through pervection but also the loss of nutrients in solution by leaching. This may be termed a ‘dry ﬂush’ effect (sensu Pearsall, 1971), which is well known in temperate uplands and mountains. On forested terrain that has been deglaciated for longer, wind-throw may perform a similar function in releasing immobilized nutrients and improving decomposition (Bormann and Sidle, 1990; Bormann et al., 1995; see also below). Aeolian deposition of nutrients may also be beneﬁcial. Disturbance not only leads to a high turnover, but may also be necessary for the maintenance of certain species’ populations. This has been argued strongly by Brandani (1983), who suggested that some trees, such as Alnus, Betula and Populus, are favoured by the regular disturbance regime of the glacial river valley, aspects of which include adverse temperature and moisture conditions in winter, and high water tables and ﬂooding in spring. He saw large annual seed crops, small wind- and water-dispersed seeds, lack of seed dormancy, rapid germination, early ﬂowering and seed dispersal, short life-spans, and the ability to sprout when damaged, as some of the disturbance-adapted traits shown in this environment (see also, Ryvarden, 1971, 1975; Chapin, 1993). Whittaker (1993) demonstrated variations in plant size-frequency distributions and behaviour patterns on the foreland of the Storbreen glacier in Norway, which were explicable in relation to patterns of succession, site moisture and disturbance. The most striking behavioural adaptation amongst the shrub species studied was the predominance of sexual reproduction on the younger terrain; establishment of ramets by vegetative means occurred increasingly on older terrain, and it was the characteristic mode of establishment in

John A. MATTHEWS

mature vegetation outside the glacier-foreland boundary. Sexual reproduction by seed seems to be the main reproductive strategy on glacier forelands as well as on fellﬁelds in the High Arctic (Callaghan, 1987). However, vivipary is a prominent reproductive strategy among the pioneer grasses at Storbreen and other Jotunheimen glaciers (e.g., Deschampsia alpina, Festuca vivipara and Poa alpina ssp. vivipara), and, according to Callaghan (1987), 20% of the vascular plant species on the foreland of K˚arsa Glacier in northern Sweden were found to be viviparous. Furthermore, on recently-deglaciated nunataks at Omnsbreen in southern Norway, wind transport of viviparous seedlings across the ground away from the parent plants (pseudo-viviparous tumbling) was found to be important in the initial stage of colonization (Elven, 1980). Observations such as these led Callaghan (1987) to suggest that this combination of sexual and vegetative reproduction is an adaptation to high levels of both stress and disturbance – the missing strategy in Grime’s (1979) classiﬁcation of species into three primary plant strategies (competitors, stress tolerators, and ruderals). Effects on animals and micro-organisms The effects of disturbance on the animals of recentlydeglaciated terrain have received little attention. Although the important early investigations of Janetschek (1949) in the Alps emphasized successional concepts, and hence the relationship between terrain age and the animal populations, recent work by Gereben (1995) on the occurrence of six species of ground beetles (Nebria spp.) suggests that habitat differences, including substrate instability, may be more important than terrain age in accounting for the ground-beetle assemblages. Little is known about the effects of disturbance on micro-organisms. Relevant research has been carried out on recently-deglaciated soils in Antarctica, where Wynn-Williams (1986, 1993) has shown that microbial assemblages of bacteria and algae form raft-like, mucilaginous coatings on frost-disturbed deglaciated soils. However, as with the black-crust phenomenon on glacier forelands in lower latitudes (Worley, 1973), emphasis has been placed on the role of these structures in stabilizing the substrate, rather than as an effect of disturbance. Davey and Rothery (1993) found differences in the species composition of micro-algal communities between frost-sorted polygons on Signy

DISTURBANCE REGIMES AND ECOSYSTEM RESPONSE ON RECENTLY-DEGLACIATED SUBSTRATES

Island that were not correlated with a wide range of environmental factors, and suggested control by the vagaries of the colonization process. However, it is possible that differences in the intensity of disturbance may account for the observed pattern.

PRIMARY SUCCESSION AND DISTURBANCE

It is clear from the above that deglaciated terrain is a highly disturbed, dynamic physical environment. How does this affect the nature of succession? The general question is approached here in terms of ﬁve more speciﬁc questions: (1) Does deglaciation yield a sterile landscape? (2) How does disturbance affect the rate of succession? (3) To what extent does disturbance force successional trajectories? (4) Do seres diverge due to disturbance? (5) What is the relationship between disturbance, stabilization and retrogression? Does deglaciation yield a sterile landscape? A useful distinction can be made between primary disturbance, whereby a new land surface is created, and secondary disturbance that subsequently affects the site (cf. Oliver et al., 1985). The classical distinction between primary and secondary succession insists that there is no vestige of a previous ecosystem in the case of primary succession, and it has usually been assumed that where deglaciation is the primary disturbance this initiates primary succession. The validity of this assumption may be questioned on several grounds. Most importantly, recent research in the Antarctic has suggested that colonization by micro-organisms, particularly Cyanobacteria and algae, may be a necessary prerequisite to a pioneer plant succession (Wynn-Williams, 1986, 1993). These micro-organisms (‘primary colonizers’) are believed to play a role in this weathering of the substrate, in stabilization of the surface, and possibly as an organic- and nutrientenriched medium for ‘secondary colonizers’ such as bryophytes and microlichens (Smith, 1990, 1993). Because bacteria have been discovered in subglacial sediments of glaciers in the French Alps (Moiroud, 1970; Moiroud and Gonnet, 1977), and viable forms that are apparently 12 000 years old have been reported from Antarctic ice cores at −55ºC (Cameron and

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Morelli, 1974; Abyzov et al., 1987; see also, Gilichinsky and Wagener, 1995), it would seem that newlydeglaciated terrain may not be as sterile as previously thought. However, the following arguments can be advanced in favour of continuing to regard the initial deglaciated terrain as an essentially sterile system: (1) it cannot be certain that the samples from both subglacial sites and ice cores had not been contaminated (WynnWilliams, 1993), (2) the global atmosphere may well be a much more important source of propagules than the subglacial sediments (cf. Broady et al., 1987); (3) as microbial populations in the proglacial area are much more abundant in the rhizosphere than in the adjacent free soil (Moiroud, 1975; Moiroud and Gonnet, 1977), the rhizosphere is an alternative source of propagules in the subglacial sediments; and (4) dependence of secondary colonizers on such primary colonization is not proven – the primary colonizers may co-exist with the secondary colonizers in much the same way as cryptogams and phanerogams may colonize at the same time on temperate forelands (Matthews, 1992). Thus, the possible existence of micro-organisms in the substrate prior to deglaciation does not appear to be of any special signiﬁcance for succession. Several other aspects of initial colonization are of relevance to the question of whether deglaciation initiates a true primary succession; but, in the writer’s opinion, these do not negate use of the term “primary succession” in the context of recentlydeglaciated terrain. They nevertheless may alter the rate of succession and possibly its course. First, again from the Antarctic, has come the concept of a glacier propagule bank (Smith, 1993). Newlydeglaciated terrain is strategically located with respect to a pool of viable propagules that have collected on the glacier surface and are released onto the proglacial area as the snow and ice melt. This may best be classiﬁed as an aspect of plant immigration that affects primary succession in ways still to be investigated. Second, whole ecosystems sometimes develop on supraglacial material, may be transported on the surface of the glacier, may undergo succession, may function as a propagule source and may survive deposition in whole or in part on the glacier foreland (Stephens, 1969; Post and Streveler, 1976; Rabassa et al., 1981; Veblen et al., 1989). Such ‘ecosystems-in-transit’ undoubtedly accelerate succession on forelands where they occur. Third, landslides, avalanches and debris ﬂows often operate in a similar fashion as loci of succession

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beneath the steep slopes associated with the forelands of valley glaciers. How does disturbance affect succession rate? Traditionally, the effect of disturbance on succession has been regarded as setting the process back to an earlier stage, or reducing the rate of succession that would occur in its absence. Examples of this, ranging from the mosaic of successional stages that can be observed on the various surfaces of slightly varying height comprising a sandur, to the bare centres of patterned ground disturbed by cryoturbation, are abundant on recently-deglaciated substrates. In addition, disturbance has been viewed as one of the allogenic factors that slows the attainment of a relatively stable (or ‘climax’) state from a few hundred years in the most favourable environments to many thousands of years in the most severe environments (e.g., Turmanina and Volodina, 1978; Solomina, 1989; Matthews, 1992). Much less attention has been given to the possibility that disturbance may accelerate succession [but see del Moral and Bliss (1993), and del Moral and Grishin (Chapter 5, this volume) in the context of volcanic substrates]. Plant damage, examples of which are provided above, is most likely to retard succession. Thus, although the initial colonization seemed as rapid at Exit Glacier as on non-ﬂooded forelands, ﬂooding of the deglaciated area led to a longer initial colonization phase of up to 30 years, an effect of the high seedling mortality (Helm and Allen, 1995). Other examples can be cited where initial colonization has been delayed by substrate disturbance, an outcome that is particularly likely where glacier-conditioned (paraglacial) disturbance is involved. According to Zollitsch (1969), the initiation of succession on the foreland of the Pasterze Glacier may be delayed by disturbance for 20 years. In the Tien Shan, terminal moraines tend to be colonized more rapidly than lateral moraines because the latter are characterized by steeper, more unstable slopes. Similarly, according to Spence and Shaw (1983), succession at the Schoolroom Glacier (Teton Range, Wyoming, U.S.A.) is retarded on moraine slopes compared to relatively stable moraine crests. The slow melting of buried ice appears to produce the longest delay in moraine stabilization and hence the initiation of succession. For example, on the ice-cored moraines of the Klutlan Glacier, Yukon Territory (Canada), it is estimated that 75 years

John A. MATTHEWS

is necessary for substrate stabilization (Birks, 1980; Wright, 1980). Kuc (1964) also has reported a major lag in colonization at Hyrnebreen in Svalbard. The examples relating to initial colonization clearly demonstrate that some disturbances can accelerate the early stages of succession by aiding the immigration of propagules with or without substrate attached. Other processes of succession – ecesis, competition, reaction and gradual allogenic change – may also be affected by disturbance, which hence may slow or accelerate succession. Under less severe, intermediate disturbance regimes, safe sites may be created by disturbance, or damage to one species may create opportunities for another. The ability of a wide variety of early colonizers to co-exist and thrive at the same site in apparently inhospitable conditions, thanks to a combination of a lack of competition with relatively high disturbance levels (Stork, 1963) seems to be a manifestation of this. Similarly, disturbance probably contributes to the early peak in species diversity that characterizes many glacier-foreland successions (e.g., Persson, 1964; Zollitsch, 1969; Elven and Ryvarden, 1975; Matthews, 1978; Sommerville et al., 1982). In summarizing vegetation–environment relationships on the Storbreen glacier foreland in Jotunheimen (Norway), Whittaker (1989) concluded that explanation of the vegetation pattern in the deglaciated landscape was improved by viewing environmental variables as multivariate ‘factor complexes’, and that the disturbance regime (especially as reﬂected in cryoturbation) formed an intimate component of a terrain-age factor complex. The negative correlation, high levels of disturbance being associated with early successional stages, suggests that the main effect of disturbance is to reduce the rate of succession (see also Whittaker, 1987). However, this may be an oversimpliﬁcation, because subsequent analysis of directly measured change over a 12-year time interval (Whittaker, 1991) demonstrated that, although the greatest amount of vegetation change occurred on terrain ages of 10–20 years, this was largely nondirectional in the successional sense. Furthermore, progressive (successional) change was occurring on terrain ages of 20–50 years, where disturbance levels were still high, which leaves open the possibility of an enhanced succession rate due to disturbance. An interesting observation of the interactions between two types of disturbance with opposite effects on succession rate has been made at Austerdalsbreen, southern Norway, by Whittaker and Petch (personal

DISTURBANCE REGIMES AND ECOSYSTEM RESPONSE ON RECENTLY-DEGLACIATED SUBSTRATES

communication). In the terminal-moraine zone of this glacier foreland, vegetation is generally heavily grazed by domestic animals, especially sheep, and succession is arrested at an early stage dominated by grasses and shrubs, even where terrain age is about 250 years. The most advanced vegetation – Betula pubescens woodland – occurs in a part of the foreland deglaciated about 60 years ago. This woodland is growing on an area of abandoned channel bars isolated from the rest of the foreland by major, active meltwater channels. In effect, therefore, ﬂuvial disturbance protects the area from extensive, persistent disturbance from the ‘artiﬁcial’ grazing regime. To what extent does disturbance force successional trajectories? That disturbance is a potential forcing factor in primary succession on recently-deglaciated substrates cannot be doubted. The extent to which allogenic factors in general and disturbance in particular controls the direction or trajectory of succession by deﬂecting it along modiﬁed pathways, or creating very different pathways, is much less clear. Disturbance may exert control by altering the initial conditions of a succession and/or by continual intervention as an inﬂux variable (Jenny, 1961, 1980; see also Matthews, 1992). The former was emphasized by Oliver et al. (1985) in their study of Nooksack Cirque, Washington, where secondary disturbances affected 63% of the total deglaciated area. The vegetation mosaic was inﬂuenced by more than one type of disturbance which, in decreasing order of area affected, were: avalanches (29%), rockslides (22%), intermittent snowﬁelds (late-lying snow in certain years; 19%), creeping snowﬁelds – large snowﬁelds where the deformation of snow under its own weight (snow creep) may disturb underlying vegetation (15%), and glacio-ﬂuvial streams (5%). Although forest development followed a general four-stage succession – the stand initiation, stem exclusion, understorey reinitiation and old-growth stages – signiﬁcant differences were detected in forest composition depending on the type of disturbance. Thus, stands developing after avalanche and rock-fall disturbance were characterized by a higher proportion of Abies amabilis (mean proportion of stand overstorey 69.8%) than stands developing after soil mass movements and glacier retreat (23.9%). The same stands, subject to avalanche and rock fall, were characterized by less Tsuga heterophylla and Alnus

29

sinuata. A second example of the lasting effects of initial disturbance is provided by differences between forests developing in abandoned meltwater channels and forests developing on adjacent moraines (Sigafoos and Hendricks, 1972; Birks, 1980; Dale and MacIsaac, 1989). Qualitative differences in species composition were detectable long after stabilization of the substrate. However, the precise mechanisms by which these differences were brought about are unknown. Some disturbances temporarily stabilize a community; in other cases disturbance precipitates a change. Frequent inundation by glacio-ﬂuvial meltwater in both the Alps and Norway favours a distinct community dominated by the moss Pohlia gracilis (Fægri, 1933; Jochimsen, 1970; Elven, 1978b). This community enters and leaves the succession in response to shifting stream courses, and may be maintained by one or more of the following: (1) the moisture regime; (2) the high concentration of silt; and (3) nutrient supply (a wet-ﬂush effect). Aeolian deposition of silt or sand can have a similar effect or, depending on the rate of deposition, may lead to replacement change. For example, on the foreland of Klutlan Glacier, Yukon Territory, Canada, wind-blown silt trapped by taller plants supports acrocarpous mosses (Birks, 1980); in front of Muldrow Glacier, Alaska, the accumulation of such material is unfavourable to the clump-forming pioneer species, Astragalus nutzotinensis (Viereck, 1966). Particularly on relatively young terrain, local disturbances clearly cause changes in the direction as well as in the rate of succession. Matthews (1992) suggested that allogenic factors, including disturbance, may be more important than autogenic factors in the replacement of pioneer species by later colonizers. The allogenic processes cited include: (1) cryoturbation and other geomorphological processes of erosion and deposition, which physically damage plants; (2) pervection, frost-sorting, and deﬂation, which produce textural changes in the substrate and hence alter the nature and availability of micro-sites for plant establishment; (3) acidiﬁcation and other chemical changes in the substrate produced by leaching, which are likely to be effective in altering nutrient supply even in the absence of signiﬁcant inputs from litter; (4) local climatic amelioration following glacier retreat and consequent distancing from the glacier climate; and (5) reduction of water supply and hence increasing drought (‘desertiﬁcation’) in response to falling water tables following glacier retreat (cf. Whittaker, 1991).

30

Although some allogenic processes (particularly those associated with the paraglacial phase) are less important in later successional stages, and autogenesis increases in importance as vegetation cover, biomass and the inﬂuence of plants on their environment increase, some types of allogenic change continue to inﬂuence successional trajectories. In addition, exogenous and endogenous glacier-independent disturbances (see above) are just as likely to affect the distal foreland as the pioneer zone. Speciﬁc cases of disturbances cited in the literature as either actual or potential causal factors in the later stages of succession include: (1) surface runoff and meltwater disturbances on glacier forelands in the Austrian Alps (Zollitsch, 1969; Jochimsen, 1970); (2) avalanches and grazing by chamois (Rupicapra rupicapra) at the Rhˆone Glacier in the Swiss Alps (Richard, 1973; Schubiger-Bossard, 1988); (3) bears (Ursus arctos) and tourists at Glacier Bay, Alaska (Lawrence, 1979); (4) wood cutting at Nigardsbreen, southern Norway (Fægri, 1933, 1986); (5) species introduction by scientists at Omnsbreen, southern Norway (Elven, 1980); (6) seals and penguins on Heard Island, Antarctica (Scott, 1990); (7) needle ice at Tyndall Glacier, Mount Kenya (Coe, 1967); (8) soliﬂuction at the Tasman Glacier, New Zealand (Archer et al., 1973); (9) needle ice, soliﬂuction, snow creep and interfacial ice (at the interface between snow cover and substrate) at Helm Glacier, British Columbia (Brink, 1964); (10) water-table ﬂuctuations in response to minor climatic changes in the Ben Ohau Range, New Zealand (Archer, 1973); and (11) wind throw at Glacier Bay, Alaska, which may contribute to the formation of muskeg from forest (Noble et al., 1984; see below, p. 31). Do seres diverge due to disturbance? The question of whether succession is convergent or divergent was a major theme of my early research on the Storbreen glacier foreland, Jotunheimen, southern Norway (Matthews 1979b–d). A variety of multivariate statistical techniques of classiﬁcation and ordination were used to analyse within- and between-community variability, and the concept of convergence towards a relatively uniform landscape in later successional stages was tested. These analyses revealed that succession in this alpine environment was characterized by strongly divergent seres: the herbaceous pioneer community exhibited relatively little within-type variability, whereas late-successional heath communities

John A. MATTHEWS

were characterized by greater within-type variability and were relatively discrete due to greater between-type differences. Other research that has suggested divergence in glacier-foreland succession includes studies by Kuc (1964) on Svalbard, Elven (1978c) at Svartisen, northern Norway, and Spence (1989) on Mount Kenya. In contrast, Fægri (1933) and Vetaas (1994) concluded that succession is convergent on the wooded forelands dominated by Betula pubescens in the sub-alpine zone around the Jostedalsbreen ice cap in southern Norway. Birks (1980) also envisaged rapid convergence on the moraines of the Klutlan Glacier, Yukon Territory, but only after Picea glauca became dominant. More complex schemes involving divergence followed by partial convergence have been depicted by Zollitsch (1969) and by Jochimsen (1970) based on alpine forelands in Austria, whereas Schubiger-Bossard (1988) saw habitat-dependent seres developing in parallel on the foreland of the Rhˆone Glacier, Swiss Alps, without any convergent or divergent trend. Matthews (1992) proposed that convergence is largely a function of relatively strong biotic (autogenic) control in the later successional stages, and that divergence is largely associated with allogenic control, which is widespread in the early stages of succession and in relatively severe physical environments. Thus, convergence tends to occur on temperate, boreal and sub-alpine forelands where a small number of tree species are dominant, while divergence is characteristic in mid-alpine and polar regions and also in sites with severe micro-environments where habitat conditions are otherwise favourable to convergence. However, the evidence is not conclusive, especially in the low-alpine zone where both convergent and divergent trends have been inferred. To the writer’s knowledge, the relationship of disturbance sensu stricto to convergence and divergence has never been examined in any detail and is not clear. Three possible hypotheses can be proposed. First, the effect of disturbance may be non-directional in the sense that it slows or disrupts the successional processes that would otherwise occur, without changing the trajectory; this interpretation emphasizes the stochastic nature of disturbance and suggests that trajectories would be less predictable but not necessarily different. Second, the association of early divergent successional change with a highly disturbed landscape, and later convergent succession with lower levels of disturbance and strong biotic control, suggests a causal relationship. Thus, disturbance may

DISTURBANCE REGIMES AND ECOSYSTEM RESPONSE ON RECENTLY-DEGLACIATED SUBSTRATES

provide opportunities for different species to become established, and hence alternative successional pathways to be initiated. This interpretation is consistent with the idea that strong biotic control, especially if associated with a tree cover, reduces disturbance and the opportunities for divergence by imposing its own environmental uniformity. Third, if disturbance is sufﬁciently ubiquitous yet not too disruptive, it may force successional trajectories along a common pathway, and hence it may lead to convergence; frequent low-level cryoturbation in high-alpine and polar environments provides a possible example. The last two interpretations both view disturbance in much the same way as any gradual allogenic environmental forcing factor that is relatively predictable. In stating the problem in terms of competing hypotheses, it is hoped that this will stimulate hypothesistesting. However, it is realized that the three alternatives may apply to different stages of the same succession, and that each may be more or less appropriate on particular forelands depending on regional or local circumstances: indeed, all three may well be found amongst the diverse environments available on glacier forelands in different parts of the globe. What is the relationship between disturbance, stabilization and retrogression? A tendency to progress towards relative stability in the later stages, in the sense that communities are more likely to be replaced by similar communities than different ones, has for long been recognized as a possible effect of the interacting processes responsible for successional change (Clements, 1928; MacMahon, 1980). It is now realized, however, that relatively mature, late-successional communities are not only subject to disturbance but also may require disturbance for the maintenance of species composition (Pickett et al., 1987; Veblen, 1992). Although there is the problem of obtaining sufﬁciently long chronosequences, several studies of glacier forelands have yielded relevant information. Retrogression, or at least retrogressive tendencies, have been described from the later stages of forest succession at Glacier Bay, Alaska (Lawrence, 1958), Klutlan Glacier, Yukon Territory, Canada (Birks, 1980) and Franz Josef Glacier, New Zealand (Burrows, 1990). In each case it would appear that prolonged succession without disturbance is related to such factors as low soil fertility and waterlogging, which lead to retrogression.

31

Signs of retrogression include a long-term decline in productivity (Bormann and Sidle, 1990), and the presence of widely spaced and/or stunted trees (Birks, 1980; Burrows, 1990). Allogenic as well as autogenic processes appear to cause retrogression, which may well occur even in polar ecosystems. For example, on terrain deglaciated for at least 3000–5000 years in front of the Sørsdal Glacier, Vestfold Hills, Antarctica, simple lichen assemblages consisting of only one or two species in low abundance have been attributed to ‘climax collapse’ in response to the accumulation to detrimental or toxic levels of wind-blown salts and loose sand (Seppelt et al., 1988). Perhaps the best-researched example of retrogression is the eventual replacement of Picea sitchensis forest by muskeg (paludiﬁcation) in southeastern Alaska. Several processes are potentially involved. First, the spruce forest may deteriorate as nitrogen, phosphorus and other nutrients are immobilized by accumulation in soil horizons and bolewood after more than 100 years of succession (Bormann and Sidle, 1990). Second, invasion of the forest by Sphagnum spp. may lead to further acidiﬁcation and deterioration of conditions for forest growth leading to its replacement after periods ranging from 800 to several thousand years (Noble et al., 1984). Third, the internal drainage system of the soil may deteriorate due to podzolization and iron-pan development, possibly over a period of about 500 years (Ugolini and Mann, 1979). Fourth, allogenic changes in climate or ﬁre regime may intervene to accelerate or reverse retrogression (Noble et al., 1984). Disturbance, particularly as induced by windthrow, which releases immobilized nutrients and improves soil drainage and decomposition, is seen as a means of spruce-forest maintenance (Bormann and Sidle, 1990), even though wind-throw may itself contribute to muskeg formation by creating pits that aid Sphagnum colonization of the forest ﬂoor (Noble et al., 1984). According to Bormann et al. (1995), windthrow or disturbances that mimic windthrow may be required at intervals of 200– 400 years to maintain the productive capacity of the soil in this and related ecosystems. It would seem, therefore, on the basis of the limited evidence available from glacier forelands, that disturbance may be associated with both the maintenance of stability and retrogression in mature ecosystems. Once again, glacier-foreland ecosystems provide an appropriate venue for further research.

32 CONCLUSION

The ecosystems of recently-deglaciated substrates come into existence following a major type of primary disturbance. They are quite widespread at present due to the world-wide tendency of glaciers to retreat following the ‘Little Ice Age’ episode of glacier expansion, the effects of which are in some respects analogous to the larger-scale glaciations and deglaciations that have periodically affected the Earth’s ecosystems through geological time. Deglaciation presents a more-orless sterile landscape for the invasion of organisms and initiates a primary succession. Throughout this succession, the ecosystems are subjected to several types of secondary disturbance, of which some (direct glacial disturbance, glacier-dependent disturbance and glacier-conditioned disturbance) are unique to recentlydeglaciated terrain and others (glacier-independent disturbance) also occur elsewhere. The direct and indirect effects of these various aspects of the disturbance regime are manifest in the form of substrate modiﬁcation, plant damage, plant behaviour, and the rate and direction of succession. Disturbance is integral to the nature of succession on recently-deglaciated terrain. In general, disturbance tends to delay the initiation of succession, to slow the rate of succession or produce rapid reversals to earlier stages, to cause divergence of successional trajectories (especially in the early stages), and to maintain the productivity and stability of mature communities in the late stages of succession. However, in some circumstances, disturbance may drive forward progressive change, cause successional convergence and bring about the slow retrogression of mature communities towards less productive communities of lesser stature. The ecosystems of recently-deglaciated substrates exist in a glacier-foreland landscape where a generally severe climatic environment (characterized in particular by low temperatures, especially in winter, and also commonly subjected to drought and strong winds) is combined with high levels of disturbance. In addition, many gradual allogenic processes, such as leaching, pervection, deﬂation, and desiccation, affect the newlyexposed substrate and inﬂuence species’ establishment and succession. Polar and high-alpine forelands are characterized by particularly severe climatic conditions, high levels of disturbance associated with frost, snow and wind, and a predominance of allogenic over autogenic change. Initial colonization and the early stages of succession are also characterized by relatively severe

John A. MATTHEWS

physical environmental conditions, and experience the full brunt of the glacier climate concurrently with intensive paraglacial disturbance of the substrate and minimal autogenic inﬂuences. Thus, whereas life in recently-deglaciated landscapes exempliﬁes disturbed ecosystems, there are problems in separating the effects of disturbance sensu stricto from the effects of gradual allogenic processes against a general background of severity of the physical environment. This emphasizes a theme touched on at several points earlier in this chapter – that is, the major deﬁciencies in knowledge concerning both the nature of the disturbance regime and its precise role in the development and maintenance of the ecosystems of recently-deglaciated terrain. First, details on the magnitude and frequency of the different types of disturbance are lacking. This could be remedied by studies of the disturbances per se. Second, the effects of the various disturbances on the ecosystems and the mechanisms by which disturbances bring about change in the ecosystems need to be investigated using a combination of descriptive, monitoring, experimental, and modelling approaches. Glacier-foreland ecosystems provide one of the few opportunities to elucidate causal relationships between disturbance regime and primary succession, and also the broader question of the relative importance of allogenic and autogenic processes in succession. To this end there is an urgent need for the conservation of selected forelands in National Parks or smaller reserves so that the scientiﬁc importance of these unique ecosystems can continue to be exploited.

ACKNOWLEDGEMENTS

I thank Drs Roger del Moral, Lawrence Walker and Robert J. Whittaker for their comments on the manuscript and Alan Cutliffe for preparing the photographs. This chapter constitutes Jotunheimen Research Expeditions, Contribution No. 130.

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Chapter 3

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS ´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

the poles are the polar fronts, which separate polar and mid-latitude wind belts and impede the transfer of heat from warmer lower latitudes. The southern boundary of the Arctic is the Arctic tree-line, located at the average summer position of the Arctic polar front (Bryson, 1966; Barry and Hare, 1974). It may be approximated by the 10ºC mean isotherm of the warmest month (K¨oppen, 1936) and by net radiation and potential evapotranspiration characteristics (Hare, 1950; Budyko, 1956). The Antarctic Polar Front and the associated 10ºC February surface air isotherm are located over the Southern Ocean. The sharply-deﬁned Antarctic Convergence, an expression of the front, is the boundary of the Antarctic zone. It occurs where the temperature minimum of the Antarctic surface water sinks below warmer, less dense, nutrient-poorer Subantarctic surface water (Deacon, 1934; Lutjeharms et al., 1985; Kanda and Kom´arkov´a, 1997). At comparable latitudes, the climate is considerably more severe in the Antarctic than in the Arctic (Schenck, 1928). The mean air temperature in the Arctic is at least 3ºC higher (Cameron, 1969) and the winter sea ice is much less extensive. Mountainous cold regions are delimited by the average position of the lower horizontal boundary of cold air, indicated by the disappearance of trees at the tree-line, by shrubs, or by freeze–thaw ground phenomena. The tree-line rises from sea level in the polar regions to 4500 m above sea level, and snow-line to 6500 m above sea level on the Tibetan Plateau. Treeline and snow-line are higher in warm, dry, continental, and protected conditions and on massive mountain ranges, and lower in cold, wet, oceanic, and severe climatic conditions on less massive mountain ranges (Hermes, 1955; L¨ove, 1970; Troll, 1973). Stress has been deﬁned in terms of energy change (Ivanovici and Wiebe, 1981), and as an unfavorable deﬂection from what is usual or expected (Odum

INTRODUCTION

Cold dominates large areas surrounding the North and South Poles and high elevation lands at all latitudes. For six months of polar winter night, the pole that is turned away from the sun experiences a net heat loss. During the polar summer day, the angle of the sun’s radiation is low, and proportionally more solar radiation is absorbed, reﬂected, and scattered along its longer path through the atmosphere than in the tropics. Mountainous regions are cold because of the low density of heat-absorbing particles, especially water vapor, in the atmosphere of high altitudes. The temperature decrease is about 1 to 3ºC per 300 m rise in elevation. Cold regions present conditions marginal to life. As latitude or elevation increases, resources decrease, the growing season becomes shorter, the complexity of food webs and ecosystems (e.g., numbers of niches or organisms) declines, and productivity and rates of nutrient cycling decrease (Wielgolaski et al., 1981). With increasing cold, the dominant plant groups change from trees to shrubs at the tree-line (Tikhomirov, 1962) to dwarf-shrubs, followed by sedges, grasses and other herbs, then bryophytes and lichens, and ﬁnally to simple ecosystems with only microorganisms. Despite the overall simplicity of high-latitude and high-elevation areas (Fig. 3.1), steep environmental gradients, and varied rocks, surface characteristics, and disturbances can combine to produce high spatial and temporal diversity in ecosystems. This diversity also reﬂects numerous stresses and disturbances ranging from frequent, narrow-amplitude, brief ﬂuctuations of environmental factors to infrequent catastrophic ecosystem destruction, all more common than in most other biomes. The boundaries of cold air masses centered at 39

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Fig. 3.1. The mountain ranges on the Tibetan Plateau, looking northwest from the main chain of the Himalaya on the border between Nepal and Tibet. The elevation of the snow-line is about 5000 m. May 1994.

et al., 1979; Barrett, 1981). “Stress” here includes environmental inﬂuences which impede the function of a biotic unit, “subsidy” includes favorable or function-supporting inﬂuences, and “disturbance” includes destructive, space-freeing, matter-removing, structure-changing events (Grime, 1979; Odum et al., 1979; Grime et al., 1988). Ecosystem resistance (the ability to remain unaffected by disturbance) modiﬁes the type and degree of damage a disturbance will cause. Ecosystem resilience [the ability to recover to a more or less persistent state (Holling, 1973; Pimm, 1984; Neubert and Caswell, 1997)] determines the rate (elasticity: Westman, 1986; Westman and O’Leary, 1986), degree, and direction of recovery from this damage. Persistence refers to presence over time, irrespective of disturbance, and constancy to low amplitude of variation over time. Ecosystems are stable if they return to their original variable states after a disturbance (Boesch, 1974; Boesch and Rosenberg, 1981; Rejm´anek, 1996a). In this chapter, we attempt to characterize the diverse cold-region ecosystems through the stresses and disturbances that affect them, and differentiate them on this basis from the ecosystems of other biomes. We also try to evaluate the importance of the resistance and resilience of cold-region ecosystems to disturbance

for their continued existence under new natural and human-related stresses and disturbances.

NATURAL DISTURBANCES

Periodic partial destruction or constant disturbance pressure are an integral part of ecosystems (Sprugel, 1976; Auerback, 1981; D.M. Raup, 1981). Cold regions have higher levels (intensity, frequency, diversity, scope, duration) of natural stresses and disturbances than most other biomes. In cold regions, unusually large heat- and gravity-related energy exchanges not only produce many disturbances, but also modify and spread disturbances originating from other causes (nutrient availability, surface disturbance by animals or humans, pollution). Heat and gravity gradients contribute to the high energy of wind, to the changes of state of water, and to other factors causing disturbances that vary with slope orientation and insolation (Moser, 1970) Persistent cold stress keeps the number of organisms low, enabling stress-tolerators to dominate in the absence of better competitors that are cold-intolerant. Such superior competitors may be established in coldregion margins, but most of them eventually disappear

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

41

Fig. 3.2. Summary of patch-creating surface disturbances and their intensities, and frequencies in some cold-region landscapes. Infrequent regional disturbances (glaciations), large patch disturbances that occur every few thousands or hundreds of years (lakes, sand dunes, landslides, other mass movement, snow-patches, volcanic ash deposits), and frequent small patch disturbances (animal burrowing and trampling, nutrient deﬁciency or excess, freeze–thaw effects, drought, waterlogging, and freezing) are integral to most cold-region landscapes. After Kom´arkov´a (1993).

during unfavorable climatic ﬂuctuations (Holdgate, 1964; Edwards and Greene, 1973). Cold leads to slow chemical reactions, decomposition, and nutrient cycling. Despite low production, organic matter accumulates in sinks located in relatively warm and moist habitats, but it is exported and redistributed from most areas by disturbance agents (wind, water, gravity). Other disturbances include drought, high precipitation, and waterlogging; deﬁcient light; excessive light and ultraviolet radiation; deﬁcient or excess nutrients; wind; volcanism; and grazing and other animal activities. Material sinks include peats (Schell and Ziemann, 1983) and peaty soils under tussocks and turfs, which are deep only in wet, ﬂat, and relatively warm areas; Antarctic mossbanks (Collins, 1976; Fenton, 1980); ornithogenic soils (Ugolini, 1972; Tatur, 1989); lake stromatolites (Parker et al., 1981; Wharton et al., 1983); lake and ocean sediments; and stable air in the polar regions (Kerr, 1981) and in mountain valleys without drainage. In cold regions at high altitudes or latitudes where cold stress is constant and deep, there are no resident terrestrial higher organisms. Fluctuating environments of cold regions at middle and low altitudes and latitudes, less constantly hostile, require multi-purpose

adaptations, and do not lend themselves to truly opportunistic strategies such as those appropriate for warm deserts, or highly specialized ones as are seen in tropical rainforests. Many resident organisms live at a low metabolic level and, in alpine plants, longterm slow growth produces fewer cells than in plants of warmer environments (Clarke, 1980, 1985; K¨orner, 1994). Low and dense plant growth forms (mats, cushions, tussocks), surface hair, small leaves, and other adaptations trap heat and moisture, and efﬁciently store water and nutrients (Dahl, 1951; Salisbury and Spomer, 1964; Biebl, 1968). Animal migration and periods of reduced activity usually both secure resources (food, space, heat) and avoid stress and potential damage during unfavorable periods (nights, winters). Small organisms live or temporarily hide in habitats under low light and darkness (within or under rocks, soil, water, snow, or ice), where ﬂuctuations of moisture, radiation, and temperature are limited. Marine ﬁsh follow zooplankton into deep water when phytoplankton production at the surface ceases in winter, also avoiding the risk of freezing. However, even normal diurnal, seasonal, and longterm environmental ﬂuctuations create numerous large disturbed patches in populations or on the ground

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(Fig. 3.2). Most disturbances, however, occur during exceptional periods when ﬂuctuations are wider or longer than average: unusually low summer temperatures, ephemeral summer snow and freezing, reduced amounts of snow in winter, late spring and early summer droughts, or unusually heavy rains. Such infrequent disturbances to which organisms cannot adapt cause most damage (Nitecki, 1981), especially to organisms at the edges of their geographical distribution. Some disturbances caused by changes in natural environmental factors may be triggered by human disturbances. These may originate at lower latitudes, but have the greatest effect in polar regions (climatic warming, thinning of the stratospheric ozone layer). Solar radiation as a disturbing factor The pronounced seasonality of light and darkness at high latitudes is unique (Dunbar, 1968, 1985; HolmHansen et al., 1977). Light stress and photoinhibition, which limit productivity, especially in the presence of additional stress factors, are the result of low temperatures, bright light, high albedo, and long periods of daylight (Adamson et al., 1988a, Howard-Williams and Vincent, 1989; Adamson and Seppelt, 1990). Low light and darkness-speciﬁc adaptations are rare, but microorganisms living in habitats which are poorly illuminated even during polar summer may show them: maximum photosynthetic rates at relatively low light irradiance, heterotrophic nutrition under both darkness and low light intensity (Palmisano and Sullivan, 1985), increased absorbency of blue-green light (Soo Hoo et al., 1987), and sensitive photoreceptive pigments (Palmisano et al., 1985; Rivkin and Putt, 1987a,b). They may be exposed to bright light when their habitat is disturbed (Vincent and Quesada, 1994). Having evolved under solar ultraviolet (UV) radiation, especially under the more injurious UV-B (280–320 nm wavelength), most organisms have some ability to avoid or screen it (pigmentation, hair, vertical migration driven by light and nutrient availability) to detoxify the highly reactive oxidants it generates, especially in hyperoxic aquatic environments (Craig et al., 1992), or to repair the damage it causes [e.g., to molecules of ribonucleic acid (RNA) or deoxyribonucleic acid (DNA)]. Protective adaptations are common in alpine regions, where the exposure to UV and shorter-wave radiation is higher than in the lowlands (intensity increases by 14–18% with

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every 1000 m in altitude), especially on sunny slopes. Such adaptations may be less common at high than at low latitudes, perhaps because polar organisms evolved under conditions of lower UV-B intensity, or because avoidance of UV is more common (Calkins and Thordardottir, 1980; Caldwell, 1968, 1979). Antarctic intertidal and subtidal invertebrates and macrophytes, and other aquatic organisms, contain mycosporine-like amino acids (MAA) that strongly absorb ultraviolet radiation (Karentz et al., 1991, 1992; Marchant, 1992). Some marine polar algae, terrestrial lichens and mosses (Adamson and Adamson, 1992), and many Cyanobacteria have other protective compounds, walls, or shells, e.g., Nostoc commune (Scherer et al., 1988), or scytonemin and related pigments, e.g., Phaeocystis (Vincent and Quesada, 1994; see also Marchant et al., 1991; Davidson and Marchant, 1994). Secondary carotenoids potentially limit photoinhibition and photodamage by quenching excess photochemical energy and trapping toxic oxidants such as free oxygen radicals, e.g., the snow alga Chlamydomonas nivalis (Bidigare et al., 1993) and Cyanobacteria (Vincent and Quesada, 1994). Melanin production in microorganisms may be an adaptation speciﬁc to Antarctic conditions (Abyzov, 1993). The polyploidy of many lower organisms in cold regions may mask the effects of single DNA mutations (Melchers, 1946; Favarger, 1961; L¨ove and L¨ove, 1974). Plant surfaces and animal skin may be protected from both cold and UV exposure by hairs or plumage (Grifﬁths and Fairney, 1988; Karentz, 1994). Because of the relatively high surface albedo, the eyes of Antarctic birds have higher thresholds for corneal damage by ultraviolet radiation than the eyes of birds from temperate areas (Hemmingsen and Douglas, 1970; Karentz, 1994). Temperature ﬂuctuations Most cold-region organisms show some behavioral, structural or functional adaptations to cold and to periodic temperature ﬂuctuations (Clarke, 1985, 1990; Block et al., 1992). At middle and high latitudes on land, seasonal ﬂuctuations often have a large amplitude, which narrows with increasing latitude and elevation and the oceanity of the climate. The temperature of free water on land may range from below −10ºC to above 20ºC (Vincent, 1988). In polar oceans, temperatures range within only a few degrees of 0ºC throughout the year. Although bare ground on land heats and cools rapidly, most organisms live in the surface layer, with

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

slower wind speeds, greater warmth, and the protection by snow and ice during unfavorable periods (Sørensen, 1941); the ground beneath a snow layer 50 cm-thick is warmed to about 0ºC by geothermal heat. Most microorganisms, and some other organisms or their tissues, are active below 0ºC; others employ winter hardening (Kainm¨uller, 1975; Sakai and Larcher, 1987), antifreeze mechanisms (Duman and DeVries, 1975; Block and Duman, 1989), countercurrent exchange (seals, humans), are insulated, maintain a high metabolic rate, or tolerate declines in body temperature. Most lower organisms can freeze or become supercooled temporarily without harm (Tyurina, 1957; Larcher, 1980; Schulze et al., 1985). An inverse relation between frost survival and atmospheric pressure may exist in the mountains (Halloy and Gonz´alez, 1993). Small size is energetically advantageous to ectothermic but not to endothermic organisms; however, small endothermic organisms can easily ﬁnd shelter. Small ectothermic amphibians and reptiles can live only on the margins of cold regions. During cold periods, birds and mammals maintain body temperature by raising their metabolic rate biochemically, by shivering, or by other muscular activity; the blood volume and heat loss are lowered through reduced protein and water intake, and the conservation of urea. Down to a certain temperature threshold, the problems of living at middle and high latitudes center on the seasonal availability of resources (Clarke, 1980, 1990), as water and nutrients are frozen during long, cold winters (Ellis-Evans, 1990; Staley and Herwig, 1993). Essential activities (reproduction) may be spread out over several growing seasons, or occur rapidly within one. In long-lived perennial plants (which are often polyploid), reserves and living green tissues facilitate the survival of winters and unfavorable summers, winter pre-formation of shoots and ﬂowers, and rapid growth and development in the spring. Plants accumulate resources over several years until the surplus is sufﬁcient for ﬂower production (Resvoll, 1917). Asexual reproduction may be important when the production of reproductive cells, propagules or seed supply is limited, or if needle ice or drought restrict germination and seedling establishment; but many vascular plants in cold regions spread by seeds even at the edges of their distribution areas, where pollinators may be absent and the taxa autogamous (Walter, 1964, 1968, 1985; Longton, 1972, 1988). In many resident heterotrophic organisms, often longlived, growth in the spring appears to be limited by

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primary production or food availability, leading to slow growth rates and delayed maturity (Mani, 1968, 1974; Østbye et al., 1975). Antarctic marine invertebrates hatch as advanced larvae or miniature adults, avoiding free-living larval stages (Clarke, 1990), and terrestrial ectothermic vertebrates are viviparous. Reproduction coincides with the environmentally and nutritionally most favorable periods; other energy-costly activities (molting) take place after reproduction is over. Many small mammals stay active in winter, as they cannot store enough energy for hibernation on account of their high metabolic rates. Low survival rates of small birds and mammals are compensated for by frequent reproduction, and rapid growth and development when conditions are favorable (Batzli, 1981). During periods of low temperatures, in the polar summer or during nights on tropical mountains, ice crystals tear cells and tissues internally and abrade them externally, and food, water, and nutrients may not be available. Sudden changes may cause die-back or die-off in terrestrial plants, and thermal shock, frostbite, hypothermia, and death in terrestrial animals. Cold spells arrest snowmelt and freeze water again in the spring (Rouse, 1984; Kane et al., 1991a, 1992). Late snowmelt enhances grubbing by lesser snow geese (Chen caerulescens caerulescens) for graminoid roots and rhizomes, with consequent damage (Jefferies, 1988; Jefferies et al., 1992). During cold summers reproduction may not occur, and organisms that have to reproduce annually may be eliminated. During summer snowstorms and hailstorms, large terrestrial wild and domestic herbivores may lack shelter, and are unable to ﬁnd food. Winter ice may persist on usually icefree lakes (McLaren, 1964; Adams et al., 1989), snowpatches or snow may not melt (Moser, 1973), and freezing may produce hypersaline lake water (Schmidt et al., 1991). Abnormally high heat supply leads to the formation of extrazonal ecosystems (Baker et al., 1964; Longton and Holdgate, 1979; Broady, 1993). Too much summer heat may lead to damage or death in insulated, overexerted endotherms, in plants with heataccumulating growth forms, or in seedlings on bare soil (Pisek et al., 1967; Larcher and Wagner, 1976; Larcher, 1980). ˇ Abnormal winters are usually damaging (Stursa et al., 1973). A sudden early onset of winter is destructive, especially to endotherms that migrate to avoid winter cold and lack of resources. Loons that cannot take off from frozen lakes in the Arctic and bird chicks that are unable to ﬂy may freeze. Molting

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adult penguins, and entire broods of penguin chicks still in down and not yet insulated for swimming, also freeze. Large sea mammals may be trapped by early sea ice and starve. Exceptionally cold winters with little snow cause widespread mortality among organisms reliant on snow protection. Non-migrating populations of livestock, inhabiting areas that remain bare of snow during average winters, may also perish during prolonged winter snowstorms (in the lower alpine region on the Tibetan Plateau: Shi, 1985; Schaller, 1993), and then the human pastoralists depending on them starve. In January 1997, snowstorms and temperatures of −36ºC caused the deaths of nearly a million livestock animals in the winter grazing regions of western China (Anonymous, 1997). Caribou (Rangifer tarandus), wolves, and foxes starve during winters with deep and long snow cover or a thick ice crust, which prevent them from reaching their usual food source or migrating (Scheffer, 1951; Vibe, 1967; Klein, 1968). Exceptionally long winters may kill off hibernators, which run out of moisture and energy supplies. During unusually warm winter periods (Gilbert and McKenna Neuman, 1988), the energy of some hibernators may be depleted; when melted snow and ground refreeze afterwards, unprotected organisms may be damaged. Hypoxia in cold-region ecosystems The partial pressure of atmospheric gases decreases with increasing elevation, while their proportions remain the same as at sea level. Hypoxia has a noticeable effect on lowland endotherms (Grover, 1974). Cattle cannot live indeﬁnitely at elevations above about 3000 m, and humans cannot live indeﬁnitely above 5500 m (Cern´ık and Sekyra, 1969; Franz, 1979). This diminishes human disturbances at high elevations such as mining, religious observances, and expeditionary tourism. Unlike the adverse effects of cold, the effects of hypoxia cannot be mitigated by behavioral adjustments, apart from descending to lower elevations or using an artiﬁcial source of oxygen. At high elevations, the cardiac and work outputs of lowland visitors are low (Frisancho, 1975). Native highlanders develop a relatively low blood pressure and large chests, hearts, and lung volume (Penaloza et al., 1963), but their reproduction and development are slowed. The efﬁciency of ﬁres and engines also decreases, by 3% per 300 m. Low atmospheric pressure and low partial pressure of oxygen contribute to an altitudinal

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ceiling of about 3600 m for most small single-engine aircraft; superchargers are effective on helicopters used for rescues, construction, and supply. Cooking times increase, as the temperature at which water boils drops with decreasing atmospheric pressure (by 1ºC per 300 m), leading to a shift to foods which do not require cooking (air-dried meat, and roasted barley ﬂour or tsampa in Tibet, freeze-dried potatoes or chu˜no in the Andes: Price, 1981). Hypoxia can occur even in polar regions. During long winters, plants in cold regions may be encased in ice; they are then exposed to hypoxia or anoxia, of which some common arctic taxa are highly tolerant (Crawford et al., 1994). Water and precipitation as disturbance factors In cold regions, water is potentially available in liquid form when temperatures rise above 0ºC. Water mediates most heat transfers, turning heat ﬂuctuations into disturbances; water volume increases when it freezes into ice, pushing away the surrounding materials, which collapse back when the ice thaws again. A heat subsidy may be counteracted by an accompanying drought. Drops in temperature during summer are ampliﬁed, and rises in temperature during winter are counteracted by freeze-desiccation, which may cause plant die-back and death (Larcher, 1957; Sakai and Larcher, 1987; Kullman, 1993). Summer waterlogging followed by freezing kills higher organisms in low-lying areas. Precipitation mostly decreases with increasing latitude and elevation. At 8000 m, the water-vapor content of the atmosphere decreases to <1% of that at sea level. Orographic uplift leads to heavy precipitation, erosion, and many streams on steep windward slopes. On lee slopes and adjacent lowlands there may be rain-shadow deserts. Seasonal storms in desert areas, or monsoons in the Himalaya, may be responsible for most of the annual precipitation, erosion, and mass movement. Unusually high precipitation events (McCann and Cogley, 1972; Ryan, 1993; Greenland, 1995) may lead to an above-average number of landslides (Selkirk et al., 1988, 1990), avalanches, ﬂoods and other forms of mass movement, which may reach the lowlands below, especially when the mountain slopes are deforested. Storm surges remove material from beaches (Harper et al., 1988). Many cold-region terrestrial organisms have adaptations to drought and wind-caused loss of heat and water vapor tension above their surfaces. In Antarctica,

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Fig. 3.3. Soil moisture in the root zones of Deschampsia antarctica (solid line) and Colobanthus quitensis (dashed line) and precipitation on Stepping Stone Island in Arthur Harbor, Antarctic Peninsula. Two periods of soil drought developed after extended periods without precipitation in January and February 1984 on the Deschampsia site. The soil drought was associated with a die-back of the monitored Deschampsia tussock and other vascular plants on the island (Kom´arkov´a, 1984).

large expanses of rocks and soil are bare of macroorganisms, even lichens, on account of drought. High soil salinity and winds limit the moisture availability in continental cold deserts, where some organisms live in environments where temperature and moisture ﬂuctuations are diminished and bright light is absent – in lakes (Light and Heywood, 1973, 1975; Parker et al., 1981) and inside porous rocks (Friedmann, 1982; Nienow and Friedmann, 1993). Invertebrates are often able to resist desiccation (Pickup, 1988; Worland and Block, 1986), which is common in normally wet habitats and shallow pools (Hawes et al., 1993). Many lower organisms rapidly resume normal function after being severely desiccated (Scherer et al., 1984; Peat and Potts, 1987). Poorly weathered and shallow soils with low content of organic matter have only low water-holding capacity, which diminishes with increasing latitude and elevation; life increasingly depends on meltwater, and must be able to withstand wide ﬂuctuations in its availability (Siple, 1938; Llano, 1962; Longton, 1988). During an exceptionally long, warm, dry Antarctic summer (Morrison, 1990), vascular plants died back after two periods of soil drought at 64ºS (Fig. 3.3); no vascular plants died back in comparable habitats at 62ºS (Kom´arkov´a, 1984). During average summers, water stress for established plants is minor even in polar deserts (K¨orner, 1994), where during unusually moist and warm seasons seedlings abound in places which are

thermally advantageous (Gold and Bliss, 1995; Gold, 1998). Water cannot move below rock or the surface of ground ice, resulting in waterlogged soils on ﬂat ground, and creep and downward ﬂow on slopes. Small mammals are excluded from waterlogged sites with a shallow depth of thaw. Polygon ponds, beaded streams, and string bogs are associated with ground ice. Streams (Fig. 3.4) and thaw lakes (Fig. 3.5), which move in a direction perpendicular to the prevailing wind, rework the ﬂat Alaskan Arctic at intervals of several thousand years (Britton, 1967; Everett, 1979; Billings and Peterson, 1980). During dry summers, meander cutbanks, pointbars which may border on sand-dune ﬁelds, and dry river bottoms are exposed to erosion by wind. Ponds and the surface peat dry out, because there is no additional input of thaw water from the wellinsulated ground ice below. Snow and ice as disturbing factors Snow and ice provide habitats which ﬂuctuate at various spatial and temporal scales. Locally, shortlived microorganisms often live within the snow or ice, sustained by nutrients from windblown materials (Kol, 1968; Palmisano and Sullivan, 1985; HowardWilliams et al., 1990). Needle-ice crystals, 1–3 cm long (Mackay and Mathews, 1974; Lawler, 1988) form nightly in moist, unconsolidated surface materials,

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Fig. 3.4. Meandering rivers redistribute loose sand deposits underlain by ground ice on the Alaskan Arctic Coastal Plain, creating point-bars. Melting of ground ice leads to natural thermokarst and collapse on a cut-bank near Atqasuk.

Fig. 3.5. Thaw lakes travel across the ﬂat Alaskan Arctic Coastal Plain in the direction perpendicular to the prevailing wind, loosening sand and creating sand-dune ﬁelds near Atqasuk. Low-center polygons and caribou trails appear on older surfaces on the right.

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Fig. 3.6. Fields of frost-shattered blocks in the foreground and talus, soliﬂuction, and rock streams on far slopes. Mountains of Nin Jin Kan Shan, at elevations between 5000 and 5600 m, near Karo La Pass east of Gyantse, Tibet. Middle to upper alpine belt.

particularly on recently disturbed surfaces such as newly deglaciated areas moistened by melt-water. This freezing and thawing promotes small-scale erosion, sorts particles, and inhibits colonization by plants (Bryan, 1946; H.M. Raup, 1951, 1965, 1981). It also shatters rocks, sometimes creating large land-forms (Fig. 3.6) (Thorn, 1979; Williams and Smith, 1989; Coutard and Francou, 1989). Lake, stream, and sea ice scour the shores in the spring, and the subsequent melting of snow and ice may cause increased erosion, ﬂooding (Marsh and Hey, 1989), and input of materials into sinks. Seasonal melting of the permafrost (which may extend to depths of hundreds of meters; Ives, 1973) typically occurs to depths of 0.5–5 m, depending on soil moisture content and insulation of the surface by organic matter. In unconsolidated materials, seasonal freezing and thawing of ice wedges (Chambers, 1966a,b, 1967; Lachenbruch, 1966), usually arranged in polygons several meters in diameter, moves and sorts particles in land-form patterns that change on the scale of hundreds of years (Everett, 1979). Loss of insulation and melting of ground ice cause ground collapse, subsidence, and mass movement (“thermokarst”: Fig. 3.7; Haag and Bliss, 1974; Babb, 1977; Lewkowicz, 1987). Ground ice aids the preservation of human food, but

it prevents burial of the dead, waste disposal, and the maintenance of safe water supplies. With increasing elevation and latitude, temperature ﬂuctuations above 0ºC decrease, then disappear. Through melting, refreezing, and/or compaction increasing toward the bottom of the snowpack, snow becomes ice and grows into glaciers that sculpt the surfaces of most cold regions. Snow erodes nivation depressions (Fig. 3.8; Thorn, 1976; Thorn and Hall, 1980; Caine, 1992a, 1995a), in which the growing season is considerably shortened (Billings and Bliss, 1959; Holway and Ward, 1963; Canaday and Fonda, 1974), and they gradually become cirques (Fig. 3.9). Glaciers quarry and grind the surrounding rocks as they move at rates from several centimeters to several meters per day. Loosened ﬁne materials may be deposited as loess (Smalley, 1966). Along with glacial streams, glaciers carry rocks and other materials downslope, depositing moraines on their sides (Fig. 3.10), and terminal moraines when they retreat. Snow and glaciers dam valleys, sometimes creating ephemeral lakes. Glaciers may surge and lakes may be sources of outburst ﬂoods (Visser, 1932; Thorarinsson, 1957; Hewitt, 1965), especially during eruptions of volcanoes covered by ice (Schoonmaker, 1998). Glaciers ﬂow into lowland forests, ﬂattening them by their weight; they may destroy villages and agricultural

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Fig. 3.7. Frost-sorted stone circles on Ardley Island, South Shetland Islands, Antarctica (62º13 S, 58º56 W). Diameter about 1.8 m.

Fig. 3.8. The upper part of “Martinelli’s snowpatch” and nivation depression on the southeast lee slope of Niwot Ridge, Front Range, Colorado Rocky Mountains. The rocky centers are free of vegetation because the snow-patch does not melt completely in most summers – only when snow-poor winters are followed by long, warm summers. The snow-patch ﬂuctuations suggest that only small, shallow snow-patches are likely to disappear soon under the current levels of climatic warming, unless the warming is accompanied by drought. In the foreground, coniferous trees and willows (Salix spp.) grow in krummholz ribbons in response to the dominant westerly winds.

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Fig. 3.9. A cirque in the Krkonose Mountains, located between the Czech Republic and Poland, in the lee of the dominant westerly winds. During forest advances towards the ﬂat plateaus above, rare cold-region plants survived in cirques, kept forest-free since the last glaciation by avalanches (Jen´ık, 1961). In the foreground, shrubs and krummholz (Pinus mugo ssp. mughus) with ﬂexible stems withstand snow creep and avalanches. Photograph by J. Ruzicka.

Fig. 3.10. A glacier and side moraines covered by fresh snow on the north side of Dhaulagiri, Nepal Himalaya. Ice from hanging glaciers scours the free-fall rock face on the right.

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Fig. 3.11. Tree krummholz (Picea engelmannii), stunted by wind desiccation and abraded by wind-driven ice crystals and sand, on the lee east slope of Niwot Ridge, Front Range of the Rocky Mountains in Colorado. Snow protects the near-ground plant parts. Small soil particles are being blown away from winter snow-free ridge-tops covered by fellﬁelds (see also Fig. 3.13).

ﬁelds on valley ﬂoors during minor glacial advances, and ecosystem zones and belts during major ones. Most cold region surfaces are scoured by glaciers at intervals between 20 000 and 50 000 years. Some surfaces in the proximity of glaciers are disturbed more often, while others in cold-region margins have not been glaciated for more than 100 000 years. Glaciers absorb and release large amounts of heat before responding to temperature changes by advancing or retreating. Sea ice surrounds Antarctica in a seasonally variable ring that extends from about 55º to 70ºS; more than 90% of this ice melts in the summer (Horner et al., 1992). The cold air and water associated with the Antarctic ice cap span almost 50º of latitude, and their thermal inertia will delay the onset of warming or cooling in Antarctica (King, 1991). The present annual accumulation of snow and ice in the Antarctic ice cap is equivalent to a lowering of the global sea-level by 6 mm, whereas the annual average rate of rise in the global sea level over the last century has been about 1 mm (Barnett, 1982, 1984; Budd, 1991). Wind disturbance Cold regions signiﬁcantly inﬂuence both global atmospheric circulation and local winds and weather.

Mountains block winds, storms, and tornadoes, cause upper-air disturbances, and inﬂuence jet streams. Latitudinally-oriented mountain ranges (European Alps) deﬂect invasions of the Arctic air. The Tibetan Plateau affects the seasonal variation of the atmospheric boundary layer and westerly circulation over Asia, causes seasonal displacement of the subtropical westerly jet stream, and also may control the inter-annual variability of the general circulation of the atmosphere and the monsoonal weather patterns of southern and eastern Asia (Flohn, 1974, 1981; Reiter and Ding, 1981). Steep temperature gradients in cold regions produce some of the strongest winds on earth. Wind speed, and the frequency of unusually strong wind-storms that redistribute surface materials, usually increase with elevation and latitude. Cool air surrounding glaciers and snow may descend with destructive force to the lowlands below ( f¨ohn, chinook). Wind exacerbates the loss of moisture and heat (“wind-chill”) from the surfaces of organisms (Siple and Passel, 1945; Warren Wilson, 1959; Rees, 1993), and most cold-region biota have some structural or behavioral adaptations which mitigate its effects. In many areas the tree-line is wind-determined. At its upper altitudinal limit, krummholz grows in the lee of boulders, protruding terrain, or in stripes oriented in the direction of the prevailing wind (Figs. 3.8 and

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

3.11). In the White Mountains of New Hampshire, on lee slopes (east-facing slopes at mid-latitudes in the Northern Hemisphere), downslope winds produce a wide krummholz belt (Reiners and Lang, 1979), and the same occurs in the Colorado Front Range (Marr, 1961), where the tree-line is lower on the lee slope than on the windward west slope by an average of 65 m (Kom´arkov´a, 1979). The redistribution of snow by prevailing winds to lee slopes produces a marked asymmetry of mountain ranges, with more numerous snow-patches, glaciers and cirques, and longer glacial valleys in the lee, e.g., Central European mountains (Jen´ık, 1961), Presidential Range, New Hampshire (L¨ove, 1965), Colorado Front Range (Kom´arkov´a, 1979) (Figs. 3.8 and 3.9). Wind redistributes sand from rivers, and moves dunes (Black, 1951), thaws lakes (Rosenfeld and Hussey, 1958; Rex, 1960; Carson and Hussey, 1962) and modiﬁes other land-forms (Figs. 3.4 and 3.5). Wind removes ﬁne mineral particles and dead organic material from summits, ridges, polar deserts, and deglaciated and other bare areas, reducing them to fellﬁelds, pavements, or sands with little available moisture, nutrients, or snow protection in winter (Fig. 3.11), and depositing the windblown particles in lee areas, sometimes distant (Warren Wilson, 1958; Teeri and Barrett, 1975; Thorn and Darmody, 1985a,b). Wind erosion is promoted by surface disturbances (Thorn, 1978). Wind also disperses plant propagules (Bonde, 1969; Rudolph, 1970; Miller and Ambrose, 1976) and birds throughout the cold regions, and deposits and redistributes nutrients and pollutants. Gases, droplets, dust, and other small particles, such as those dispersed by volcanic eruptions, can be carried around the globe (Franz, 1979; Rahn and McCaffrey, 1980; Bodhaine et al., 1981). Disturbance by mass movement The abrupt gradients of temperature and elevation of cold regions contribute to a level of mass movement much higher than that in temperate regions. The effects of gravity increase with elevation and slope angle. Vegetation, which impedes erosion, stabilizes slopes, and increases the water-holding capacity of soils, is often scanty in cold regions, the soils are shallow, and most precipitation and meltwater run off. Water, snow, and ice signiﬁcantly increase erosional rates by decreasing the friction between the falling particles and the ground.

51

Runoff and streams remove particles and solutes to lowlands, and undercut slopes (Caine, 1992b, 1995b). Rates of erosion and solute removal (Litaor, 1993) increase with water volume during high precipitation events, spring melting of snow and ice, surface microdisturbances (Friedmann and Weed, 1987), and rainy seasons. Erosional rates may reach 1 mm per year in high-relief mountain ranges with a humid climate, but in continental areas with gentle relief they are considerably lower (Caine, 1974). Some high mountain ranges (e.g., the Appalachians) have been worn down by erosion, and their populations of coldadapted organisms are probably extinct. Chances of migration are better when erosion of mountain ranges coincides with cooling of lowlands. Saturation of slope materials by water from rainstorms or snow-melt increases their weight and may cause abrupt sliding, especially when they are underlain by sloping rocks with layers of shingle. Vegetated soliﬂuction lobes, terraces and similar features creep down at speeds of several centimeters per year (Benedict, 1970), more on steep slopes, and under diurnal or seasonal frost-heave (Bunza, 1975). Scree and its accumulations continue to move at rates of up to 20 cm per year (Benedict, 1970; Price, 1991; Matsuoka and Moriwaki, 1992); see also Fig. 3.6. Rock glaciers, with steep front sides and ice cores, develop in areas with low snow-fall and a continual supply of rocks from above. Their rates of movement are up to 50 cm per year, and they are usually not vegetated. Rockfalls, mud-ﬂows, debris ﬂows, landslides, avalanches (Voellmy, 1955; Luckman, 1977, 1978), ﬂoods, lahars, and other forms of mass movement vary in velocity, extent, and severity, but often destroy forests, crops, transportation corridors, and human settlements. The incidence and intensity of disturbance increase particularly during earthquakes, volcanic eruptions, and unusually wet years, and after deforestation. In low-elevation mountains, avalanches prevented forest establishment in lee cirques during the warm postglacial period, enabling cold-region plants to survive there while the summit plateaus were forested, e.g., Central European mountains (Jen´ık, 1961), Presidential Range, New Hampshire, USA (L¨ove, 1965); see also Fig. 3.9. Volcanic eruptions create bare patches of lava and tephra reaching into the lowlands. They may cause melting of snow and ice and ﬂoods; by introduction of particles and droplets into the atmosphere they may cause climatic cooling, reducing crop production

52

worldwide; and they may affect precipitation chemistry, contributing to acid rain (Williams, 1969; Lewis and Grant, 1981; Wood and del Moral, 1987). Many such disturbances originate in cold regions, e.g., Iceland (Arnalds, 1987; Arnalds et al., 1987), Aleutian Islands, Scotia Arc (Collins, 1969); and from other volcanoes rising from plate boundaries and hot spots into coldregion altitudes. In the presence of moisture, new volcanic ash offers opportunities for habitat diversiﬁcation, organismic colonization, and geographical range expansion (Young and Kl¨ay, 1971). Fire Unlike forests or grasslands in which ﬁre is often important, most cold regions have little combustible material. Even in the margins of the cold regions with relatively high production, litter is often removed by wind, runoff, gravity, or snow creep. Most ﬁres occur in the Arctic tundra only during exceptionally dry, warm summers when accumulated peat dries out. In western and northern Alaska, 3800 km2 burned during the dry summer in 1977, but the ﬁre frequency per annum ranged from 0.4 to 12 ﬁres per 5000 km2 over a period of 23 years (Racine et al., 1987). Tundra ﬁres are usually an extension of forest ﬁres (Cochrane and Rowe, 1969; Douglas and Ballard, 1971; Wein and Bliss, 1973a); they can also be initiated by lightning or by humans. Arctic ﬁres burning only for a short time do not affect the ground. Long-burning ﬁres over ground ice increase the depth of thaw and may cause surface subsidence on slopes, particularly when the thick insulating vegetation and the surface layer of organic matter are burned; there may be a surplus of released moisture. The increase in depth of the thawed layer above permafrost may return to the original within six years, or it may continue even for 20 years, and burned areas may maintain reduced heat loss, higher temperatures and moisture, and consequently higher nutrient availability, lower albedos, and higher net radiation (Mackay, 1970, 1995; Bliss and Wein, 1972). Disturbances through effects on soil and nutrients In most cold regions, the inﬂuence of parent material is stronger than in other biomes; soils are shallow and erosion rates high, and essential or toxic elements

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

released from minerals are continually available (Fernald, 1907; Braun-Blanquet and Jenny, 1926; Franz, 1979). However, even on alpine slopes, leaching by rain containing weak carbonic acid acidiﬁes limestone soils so that their original ecosystems are invaded by acidophilous taxa and are usually impoverished (BraunBlanquet and Jenny, 1926). Calciphilous dominants and soil buffers may counteract these changes. On Subantarctic Macquarie Island, serpentinite, harzburgite, and basalt outcrops support areas of small and sparse vegetation with a high proportion of bare ground and low Ca/Mg ratio, unlike the surrounding dense vegetation on gabbro (Adamson et al., 1993). In the coldest areas, low temperatures dominate and rock weathering is slow. A high toxic salt content is common in the dry valleys of the southern Antarctic, where salts released by in situ weathering accumulate (Miotke, 1988). Sea spray accumulates on marine beaches, and the plants there have a high salt tolerance (Huntley, 1971; Gremmen, 1982), such as Poa ﬂabellata (R.I.L. Smith, 1985a; Dawson and Bliss, 1987; Bliss, 1988). Low-growing and sparse vegetation does not efﬁciently protect soil from erosion and scavenge nutrients from the air. In most areas, wind, snow, ice, and runoff export most of the ﬁne materials. The availability of nutrients is low and their turnover slow (Warren Wilson, 1957; Haag, 1974; McKendrick et al., 1978), but organisms are adapted to these conditions; plants have leaf nitrogen content higher than plants of temperate climates (Chapin et al., 1975; Chapin and Oechel, 1983; K¨orner and Larcher, 1988). Slow decomposition leads to organic matter accumulation in some terrestrial ecosystems (marshes, lowlands, tussock and turf communities; mossbanks in the Antarctic), but only some of them produce enough organic matter to support herbivorous mammals and their predators. The low productivity of tundra plants can be increased by fertilization (Chapin et al., 1975; McKendrick and Mitchell, 1978a,b). However, artiﬁcial addition of nitrogen and phosphorus causes changes in species composition and eventually in ecosystems, as grasses respond more quickly than shrubs and sedges (Jeffrey, 1971; Hegg, 1984a,b). Once the disturbance by fertilization ceases, taxa enhanced by it are again out-competed by strong competitors (usually tussock or turf-forming Cyperaceae). In many areas, nutrients supplied by the atmosphere (Wilson and House, 1965; Desideri et al., 1994; Cress et al., 1995), sea spray (Jenkin, 1975; Benninghoff and Benninghoff, 1985; Karl, 1993), or imported and

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

53

Fig. 3.12. Nests of gentoo penguins (Pygoscelis papua) at Point Thomas, King George Island, South Shetland Islands, northern Antarctic. The bare area surrounding the nest contains excrement rich in krill chitin and toxic levels of nutrients. The surrounding area with lower nutrient concentrations is colonized by algae, mainly Prasiola crispa. Farther away, nutrients are diluted enough to subsidize abnormally luxuriant growth of Deschampsia antarctica, tolerant of elevated nutrient levels. Normal Deschampsia grassland is about 20 m away from the nest. January 1984.

redistributed by strong winds, water, or animals play a greater role in the distribution and production of organisms than nutrients that are cycled or supplied locally (e.g., by nitrogen ﬁxation: Fogg and Stewart, 1968; Croome, 1973; V.R. Smith, 1985). The dependence of organisms on nutrient subsidies increases with altitude and latitude. Most nutrients that support the productivity of cold, barren ecosystems are probably blown in from more vegetated areas or from the ocean by wind (Swan, 1963, 1967; Papp, 1978; Edwards, 1988), deposited on windward slopes, and transferred to lee snow-patches and glaciers, where they accumulate. Meltwater from glaciers and snow carries high loads of remineralized inorganic nutrients, leachates, and particulate detritus (Benninghoff and Benninghoff, 1985), which are channeled to ecosystems below, and then by streams to the lowlands or washed out to sea. From the Antarctic Plateau, they are carried to the coastal ice-free areas (Allen et al., 1967; Claridge and Campbell, 1968). Nutrients are also transferred to land by nesting and breeding marine birds and mammals through their feces and urine, molted skin, fur and feathers, regurgitated shells and other material, and carcasses (Siegfried et al., 1978; Batzli, 1983; Ryan and Watkins, 1989). For

example, South American condors (Sarcorampus papa) nest in the high Andes, but migrate daily to the sea where they feed on dead ﬁsh (Lettau, 1967; Price, 1981). Natural nutrient enrichment by nitrogen and phosphorus in urine and feces near small-mammal burrows, bird posts, and nests can signiﬁcantly enhance the diversity and luxuriance of vegetation (Figs. 3.12 and 3.13). Enrichment also occurs at carcass sites, because the concentration of most nutrients (except potassium) is much greater in herbivores than in plants (McKendrick et al., 1980). Grazing by lesser snow geese increases net aboveground primary production of coastal salt marshes by 40–100% (Jefferies, 1988). In the Antarctic, penguins nest in colonies on ornithogenic soils created by long-term accumulation of droppings (Ugolini, 1972; Heine and Speir, 1989; Tatur, 1989), which they also deposit on snow-patches where they cool down after ﬁshing. Fresh penguin droppings contain about 2.55% phosphorus and a high nitrogen content, reﬂecting the high protein diet of penguins (Ugolini, 1972; Pietr et al., 1983). High levels of nutrients in rookeries are lethal to most organisms (Rudolph, 1967). Penguin droppings contain the antibiotic acrylic acid (Sieburth, 1961, 1963),

54

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Fig. 3.13. An area disturbed by pocket gophers on a ridge-top on Niwot Ridge, Front Range of the Rocky Mountains in Colorado. The animals burrow and graze below ground, increasing soil nutrients through their feces and urine. The patch is colonized by nutrient-subsidized grasses and dicotyledons, contrasting with the surrounding wind-blown fellﬁeld free from snow in winter (Fig. 3.11) with sparse cushion plants.

and their decomposition yields oxalic acid (Akiyama et al. 1986, 1988). At a distance from the rookeries, diluted nutrients provide a subsidy to vascular plants (Fig. 3.12). Burrowing petrels (Procellariidae) are the most important source of nutrients in vegetated areas on Marion Island (Burger et al., 1978; Smith, 1979; Lindeboom, 1984), where contributions from elephant seals (Mirounga leonina) affect plant and invertebrate distribution (Huntley, 1971; Smith, 1976; Gremmen, 1982). Giant petrels (Macronectes giganteus) and other scavenging and predatory seabirds redistribute the materials derived from carcasses of penguins, seals, etc., as guano, eggs, feathers, droppings, or their own carcasses (Conroy, 1972; Williams et al., 1978; Hunter, 1985). Growing populations of Antarctic fur seals (Arctocephalus gazella) on Signy Island in the northern Antarctic increase the output of nutrients from land into the ocean (Ellis-Evans 1985, 1990; Smith, 1988). Animals as disturbing factors In cold oceans, which have been harvested by humans for centuries, little is known about population dynamics, trophic level composition, energy ﬂow, and the disturbances caused to lower trophic levels by

consumers from higher trophic levels. In terrestrial cold environments, low-growing, often sparse, vegetation producing little (usually 100–800 g dry mass m−2 y−1 ) is easily overexploited, leading to die-backs of consumer populations, and is easily disturbed by burrowing and other activities. Where present, large mammals and small rodent herbivores affect both primary production and decomposition, and are more important than invertebrates or birds (Tikhomirov, 1959; McKendrick et al., 1980; Batzli et al., 1981). In the Antarctic, where native terrestrial herbivores are all invertebrates, most of the primary production is utilized by decomposers. At high elevations and latitudes, predators are supported mainly by detritivores, and herbivores may be absent. Some mammals have developed hooves that are effective for digging through soft snow (caribou and reindeer (Rangifer tarandus): Leader-Williams, 1988) or climbing rocks (Hoffmann, 1974). Animal populations in cold regions are highly variable spatially, and ﬂuctuate strongly in time, both within and between years, particularly at higher latitudes (Arctic herbivorous insects: Haukioja, 1981). Caribou herds appear to be regulated by the availability of food (Messier et al., 1988; Crˆete and Huot, 1993; Crˆete and Manseau, 1996). The fact that ﬂuctuations of distant populations of Ad´elie penguin (Pygoscelis

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

adeliae) occur simultaneously suggests that they reﬂect wide-ranging environmental disturbances affecting the whole Ross Sea (Taylor et al., 1990). Overpopulation occurs only in a few productive areas and during climatically favorable years. In cold-region margins, populations of small mammals reach excessive levels every 3–10 years (Schultz, 1968; Kalela et al., 1971; Tast and Kalela, 1971). Disturbances occur to both vegetation and herbivore populations, especially when primary production drops following such highly favorable summers. Excessive populations of lesser snow geese destroy Arctic wetlands, which erosion then changes into mudﬂats (Jefferies, 1988; Kerbes et al., 1990), and overuse by caribou caused persistent suppression of shrub birch (Betula glandulosa) in the northern Labrador Peninsula (Crˆete and Doucet, 1998). Few alternative food sources are available to predators when their food source is destroyed by climatic ﬂuctuations, which perhaps contributes to the relatively low number of terrestrial predators in cold regions. Low-level damage by grazing, browsing, and passage of animals may be typical for most of the productive ecosystems at cold-region margins. Repeated browsing by caribou and girdling by collared lemmings (Dicrostonyx torquatus) has damaged willows (Salix spp.) in tussock Arctic tundra, and microtine grazing has been shown to damage graminoids and mosses, particularly during peak population years (Hansson, 1969; Kalela and Koponen, 1971; White and Trudell, 1980). By winter grazing at the snow/ground interface, microtines damage mosses and decrease primary production during the summer in the Arctic (Batzli and Jung, 1980; Batzli, 1993); in alpine regions, intensive shredding by microtines produces patches of dead vegetation. In Alpine grasslands, grasshoppers removed between 19 and 30% of the aboveground phanerogam biomass in 1993 (Blumer and Diemer, 1996). Leaves which live for more than one growing season need to be protected by secondary compounds from herbivores and overwintering stress, and have lower concentrations of nitrogen and phosphorus than deciduous leaves, which are preferred by herbivores (Archer and Tieszen, 1980; Chapin, 1980; Chapin et al., 1980a). At least some secondary compounds are toxic to microtines (Jung et al., 1979; Batzli and Jung, 1980; Oksanen et al., 1987), to snowshoe hares (Lepus americanus) (Reichardt et al., 1990), and to some herbivorous insects (Niemel¨a et al., 1979). Retention

55

of standing dead leaves for one or two years may be a defense effective against ungulate grazers, as caribou do not eat evergreen shrubs (White and Trudell, 1980). Bitter chemicals protect some dicotyledonous herbs (Gentiana lutea, in the European Alps) and lichens. There seem to be few structural adaptations against herbivores, which apparently play a role smaller than in some other biomes. By periodically cutting plant material, disrupting the snow layer, and depositing feces, Arctic lemmings (Dicrostonyx, Lemmus) increase decomposition rates and plant production; ﬂowering may increase and root regrowth may be depressed (Schultz, 1968; Batzli and Jung, 1980; Batzli, 1981). Arctic ground squirrels (Spermophilus parryii) prefer plants with high water content, which may be more palatable (Batzli and Sobaski, 1980), and various Arctic grazers prefer ﬂowers and young plant parts, which contain more nutrients (White, 1983). Flower removal may favor vegetative reproduction. In the Colorado Rocky Mountains, pikas (Ochotona princeps) store plants rich in allelochemicals which increase the length of time for which the plants can be preserved, and decay by the time they are consumed (Dearing, 1997). Migratory herds of Arctic caribou and reindeer compact soil on permanent trails along old ridges and streambanks where it is easier to walk, and trample other vegetation (Pegau, 1970a; LeResche and Linderman, 1975; McKendrick et al., 1980). Trampling by introduced reindeer has created persistent holes in moss carpets on South Georgia (Kightley and Smith, 1976). Microtine trails and nests and dens of other animals also disturb vegetation. Reindeer and muskoxen (Ovibos moschatus) destroy lichens and cause erosion by digging for food in winter in places with shallow snow (Østbye et al., 1975; Raillard and Svoboda, 1989; Wielgolaski, 1997). Resting and molting seals compact and damage peat and vegetation, and ﬂick sand and loose soil over themselves, which causes erosion, and affects populations of invertebrates and birds along the Antarctic and Subantarctic coasts (Gillham, 1961; Croxall et al., 1984; Bonner, 1985). Burrowing animals (Hoffmann, 1974) create bare patches, which may be enriched in nutrients, but also increase erosion and lower the soil carbon content (Thorn, 1978, 1982; Hall and Williams, 1981). They include pocket gophers (Thomomys talpoides), which feed on underground plant parts in the southern Rocky Mountains (Fig. 3.13; Thorn, 1978, 1982), introduced rabbits (Oryctolagus cuniculus) in Macquarie Island

56

(Griffen, 1980; Selkirk et al., 1983; Scott, 1988), Arctic ground squirrels (Spermophilus parryii), which sometimes create round hill burrows several meters high, conspicuous in the surrounding ﬂat landscape (Mallory and Heffernan, 1987), and rock hyrax (Procavia capensis) and groove-toothed rat (Otomys orestes) on Mount Kenya (Mahaney and Boyer, 1986). On the Tibetan Plateau, there are a few burrows (at densities of 10 per ha−1 ) of pikas (Ochotona spp.) and zokors (Myospalax spp.) which graze above the ground, while the surface is covered by a mosaic of burrow patches in varying stages of recovery at a density of 100 or more per hectare (Liu et al., 1982; Xia, 1986). High burrow densities sometimes found over areas up to 1500 m2 may result in bare, wind-eroded patches in which vegetation recovery is difﬁcult (Ekvall, 1968; Xia, 1986). HUMAN DISTURBANCES

The ﬁrst use of cold regions by humans was probably on hunting and gathering trips from adjacent areas. Camps and settlements were necessary in remote, extensive regions where humans were almost exclusively dependent on local resources. Most important were the availability of animal prey and animal products used in shelter construction, clothing, and heat production, and the preservation of edible animal products over winter when prey was scarce (Tibetan Plateau: Ekvall, 1968; Arctic: McGhee, 1974). At the edge of the Siberian Arctic, settlements occurred as early as 300 000 B.P. (Waters et al., 1997). In most Arctic and alpine areas, they ﬂourished particularly during warmer and drier periods (e.g., 9000–5000 B.P. and 1200–1000 B.P.). Mountain ridges and ice prevented interchanges with neighboring populations, and unique gene pools, cultures, and languages developed. Small native populations and their limited technology damaged cold-region margins through hunting, ﬁshing, grazing of domesticated animals, mining, plant litter collection, cutting grass and ﬁrewood, and burning forests, and they modiﬁed the landscape by terraces, irrigation systems, trails, corrals, encampments, gamedrive walls, monuments, shelters, and other structures. Limited resources led to customs such as primogeniture, polygamy, polyandry, mandatory emigration, celibacy, and prohibition of marriage, and to various systems of private and community ownership, transportation, and cooperation (Husted, 1974; McGhee, 1974; Simchenko, 1976).

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Today, mountain settlements are common at altitudes of 4000 m, and up to 2.5×107 people live above 3000 m, most of them on the less densely populated cold plateaus of Asia (Tibet, Pamir–Tien Shan) and on the warmer and more densely populated plateaus of South America and Africa (mainly Bolivia, Peru and Ethiopia) (Hedberg, 1964, 1978; Coe, 1967; Holdridge, 1978). Defended by mountains, some highlanders have independent states (Switzerland), but most natives were subjugated by lowlanders (Tibet by China) or split up among different states (Arctic). On account of population and political pressures, some of these cultures are now disappearing, and some are engaged in armed conﬂict (Davidson, 1993). Some mountain agricultural landscapes, usually including extensive terraces, are deteriorating due to emigration of rural labor (Vogel, 1988). Other native populations are still supported by traditional agriculture, animal husbandry, hunting, and ﬁshing. Other subsistence or cash-generating activities include mining (up to 6000 m in the Andes), rock and gravel extraction, trading, home manufacturing, collection of medicinal plants, guiding, load carrying, and transportation. Most populations are becoming sedentary, utilize resources from lower latitudes and altitudes, and mix traditional and modern pursuits (Quigley and McBride, 1987; Usher and Wenzel, 1987; Smith and Wright, 1989). Increasing population pressures from lower altitudes in the mountains of less developed countries cause environmental overuse, a breakdown in patterns of land use, cooperation, and trading, and eventually emigration of native populations. Natives may also emigrate or settle in marginal habitats when immigrants overwhelm them, marginalizing their economy and instituting ill-advised policies (Chinese in Tibet: Goldstein and Beall, 1989; Goldstein et al., 1990; Tsering, 1992). Low-elevation land is being used for housing, military bases and installations, and other purposes, thus restricting croplands and lower pastures. Firewoodcutting, terracing, irrigation, and upper pastures are extended upward into unsuitable areas, resulting in erosion and loss of habitat, vegetation cover, and native organisms (King and Jackson, 1982; Harden, 1988; Longrigg and Rowe, 1990). The removal of mountain soils leads to increased runoff, ﬂoods, and water shortages; siltation of river channels, dams, and canals; and to decreasing the build-up of fertile soil in the lowlands (Byers, 1991; Bandyopadhyay and Gyawali, 1994).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

57

Fig. 3.14. Zones of ecosystem damage on the Kola Peninsula in the Russian Arctic. Zone I: completely destroyed ecosystems around industrial centers, especially nickel–copper smelters (Nikel’, Monchegorsk); this zone receives 25–30 tons of sulfur and 5–6 tons of nickel, copper, manganese, and zinc per square kilometer annually. Zone II: highly disturbed ecosystems receiving 3–5 tons of sulfur and 2–5 tons of metals per square kilometer annually. Zones III and IV represent intermediate stages, and zone V the initial stages of ecological damage. Undisturbed ecosystems occur only on the fringes. From Doiban et al. (1992).

In the densely populated mountains of more developed countries, where the tree-line has been depressed by centuries of wood-cutting and farming (Mikkelsen and Høeg, 1979; Tranquillini, 1979), the maintenance of attractive forest meadows needs to be supported by government aid to farmers, who often emigrate to lowlands or shift to more proﬁtable, now almost year-round, tourism and recreation, e.g., European Alps (Lukschanderl, 1983; Brugger et al., 1984; Broggi, 1985), Carpathians (Kotarba, 1992), mountains in general (Allan et al., 1988). Most labor and resources must be supplied to cold regions from lower latitudes and altitudes, and development is expensive. Warfare, pollution, and storage of military materials have affected some areas (parts of the Arctic, Kashmir, South America, the European Alps, Tibet). Throughout the Arctic, a sparse network of airﬁelds, rivers, ports, roads, and railroads supports military activities, research, and regional resource and economic development (Rempel, 1970; Walker, 1970; West, 1976). The environmentally destructive exploitation of rich mineral and fossil fuel resources in the margins of the Russian, North

American, and Scandinavian Arctic has been, and may still be, proﬁtable. The vast Russian Arctic has assumed a great strategic importance since the days of the Soviet Union. Nearly 4×106 people, mostly immigrants, live there: 40% of them work in mining districts (Kola, Vorkuta, and Noril’sk); 25% ﬁsh, or work in the military; and 15% are aboriginal people. The Kola Peninsula contains Murmansk, the largest city north of the Arctic Circle, Russia’s main naval base, and the largest military base complex in the world (Bergesen et al., 1987; Osherenko and Young, 1989; Doiban et al., 1992); see also Fig. 3.14. Russia’s northern population has continued to grow to service the extractive industries (Armstrong, 1990). Development often conﬂicts with the interests of native peoples and native organisms (Kennedy, 1988; Chance, 1990). In the Western Hemisphere, native land claims (M.E. Thomas, 1986) and environmental impact studies have postponed and even prevented some projects (National Foreign Assessment Center, 1978); this is now occurring also in other countries (Vitebsky, 1990a). Sustainable development is more difﬁcult to achieve than in temperate biomes (Pretes

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and Robinson, 1989; Lyck, 1990; Arikaynen, 1991). Some native peoples have beneﬁted from land grants and government cash settlements or support, e.g., Alaska (Morehouse, 1989; Flanders, 1989), Greenland (Lyck and Taagholt, 1987; Poole, 1990), which could, along with natural-resource trust funds administering the current income from non-renewable resources, provide ﬁnancial stability to native populations (Pretes and Robinson, 1989; Robinson et al., 1989; Knapp and Morehouse, 1991). Expeditions, research, and tourism at high latitudes and altitudes mostly occur during the summer. Antarctica supports transient populations of tourists and personnel operating research stations, ships, and aircraft (Parker et al., 1978; Parker and Angino, 1990). Landbased, year-round activities are more damaging than ship-based ones, because they further encroach upon the limited ice-free areas (Lipps, 1978; Harris, 1991; Olech, 1996). Antarctic deposits of minerals and fossil fuels are administered under the international Antarctic Treaty (Shapley, 1985; Gulland, 1987; Vicu˜na, 1988), which also bans military presence. The extraction of these resources would cause serious environmental problems (Gregory, 1990). Harvesting of organisms Humans have long utilized organisms, primarily animals, native to cold regions, and hunted many terrestrial animals to extinction during the Ice Ages. Caribou and reindeer, which are much easier to kill than polar bears (Ursus maritimus), probably supported most ancient Arctic populations, along with hares (Lepus spp.) and ptarmigans (Lagopus spp.). In the Russian Arctic, reindeer accounted for 49.7% and polar bear for 43.8% of the prey of hunters at 7800 B.P. (Makeyev et al., 1992); hunters followed seasonal reindeer routes (Simchenko, 1976). In the Rocky Mountains 5000 to 9000 years ago, game-drive walls directed bighorn sheep (Ovis canadensis) to concealed hunters. A tradition of hunting by Ute and other Native American tribes (Ives, 1942) continued after the arrival of Europeans; by the 1960s, the numbers of vicu˜na (Vicugna vicugna) in the Andes were reduced to 1% of previous levels – to a total of 10 000 individuals (Rabinovich et al., 1985). In the Alps, only a small herd of ibex (Capra ibex) survived (Grabherr, 1997). The greatest damage occurred during the 19th and early 20th centuries (Rossnes, 1991), when improved technology and communication increased

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the trade in animal products (fur, skin, meat, eggs, down, fat), and there were few measures to protect the animals. In populated areas today, the interests of native populations preserving their traditional way of life (Nuttall, 1990, 1991), the settler populations practicing subsistence, the animals and the institutions protecting them, and commercial sport hunting and ﬁshing, are often in conﬂict (Marquette and Bockstoce, 1980; Riewe, 1981; Freeman, 1985). Large-scale collection of medicinal and ornamental plants, hunting for subsistence purposes, for medicinal and ritual items, and for trophies (Murphy et al., 1990), and ﬁshing, still decimate the gene pools and may lead to extinctions. In Arctic Alaska, 50% loss of bowhead whales (Balaena mysticetus) that have been struck during the annual Inuit subsistence hunt (Marquette and Bockstoce, 1980; Mitchell and Reeves, 1980) is being reduced by radio telemetry (Follmann and Manning, 1989). On the Tibetan Plateau, rare and protected wild animals are hunted under permit; others are killed for meat, wool, musk, and other products sold commercially. In some areas, wild animals are not tolerated because they are thought to compete with livestock for food (Palber, 1992; Tsering, 1992; Schaller, 1993); a population of 70.4 pikas per ha−1 consume more than one Tibetan sheep (Ovis aries) (Pi, 1982). In some areas, animals that have been almost extinct are being reintroduced, e.g., the bearded vulture (Gypaetus barbatus) in the Alps, by a nongovernmental organization “Alp Action”. The damage is most systematic in the oceans, which in cold regions are easier to exploit than the land, despite local environmental ﬂuctuations (Sanger, 1980, 1988, 1991; Sanger and Dickinson, 1989). For centuries, Arctic peoples ﬁshed and hunted seals sleeping on the ice, and whales surfacing in narrow leads. Commercial sealing and whaling, which began off Svalbard in the 17th century, reduced most populations to 25% or less of their initial stock size (Braham, 1992). One thousand human residents of St. Lawrence Island in the Bering Sea starved in the winter of 1878–79 after commercial whalers, ﬁnding bowhead whales (Balaena mysticetus) scarce, shot thousands of Paciﬁc walrus (Odobenus rosmarus divergens for their ivory (Bockstoce, 1986; Huntington, 1992). Steller’s sea cow (Hydrodamalis gigas) of the western Bering Sea was exterminated in 1768, within 30 years of its discovery (National Foreign Assessment Center, 1978), the bowhead whale was reduced to less than

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

10% of its population of about 12 000 before the start of commercial whaling (Reeves and Leatherwood, 1985; McCartney and Savelle, 1993). Sealing, penguin harvesting (Swithinbank, 1993), and whaling have caused perhaps irreversible damage to the ecosystems of the Southern Ocean since the late eighteenth century (Hempel, 1985; Kriwoken and Williamson, 1993); some populations have been reduced to a small fraction of their original size. Today, Arctic seas account for about 10% of the current recorded ﬁsh catch of the whole world, but overﬁshing using modern technology is causing the disappearance of some ﬁsh stocks (National Foreign Assessment Center, 1978; Wielgolaski, 1990). Overﬁshing in the Antarctic is threatening krill (Euphausia superba and other species), and with them the base of the Antarctic food webs, consisting of tens to hundreds of millions of tonnes of krill in the Southern Ocean (Hempel, 1985; Kock, 1985; Jouventin and Weimerskirch, 1990). Some krill ﬁsheries may still be competing with birds and mammals, particularly during their breeding season – the more so because the stocks of many ﬁn-ﬁsh, potential alternative prey, have been depleted. Some populations of birds and mammals, and their contributions of nutrients to terrestrial ecosystems, may be decreasing also because of the collateral mortality during ﬁshing (Weimerskirch and Jouventin, 1987; Weimerskirch et al., 1987; Croxall et al., 1990). Southern giant petrels (Macronectes giganteus) suffer high mortality from ﬁshing operations and from shooting at sea (Jouventin and Weimerskirch, 1990). Crabeater seals (Lobodon carcinophagus) may have declined in the Weddell Sea as a result either of increasing competition by recovering populations of baleen whales or of commercial krill harvesting, now at 4×105 metric tons per year (Erickson and Hanson, 1990). Increased ﬁshing has also led to declines in elephant seals, southern giant petrels, wandering albatrosses (Diomedea exulans) in the Crozet Islands (Weimerskirch and Jouventin, 1987), and blackbrowed albatrosses (D. melanophris) in Kerguelen (Weimerskirch et al., 1987). Elephant seal populations have remained stable in South Georgia (McCann and Rothery, 1988). Trawling seriously damages benthos, which has low resilience (White, 1984; Dayton, 1990; Jones, 1992).

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Agriculture and animal husbandry Only some cold-region margins and relatively warm plateaus offer opportunity for limited crop production, and human residents usually depend on trading to obtain crop products. Cold and climatic ﬂuctuations make cropping unreliable, and limit the effects of fertilization and irrigation. On the Tibetan Plateau, crops can usually be cultivated only in protected valleys (Fig. 3.15) where the number of days where the temperature reaches 10ºC exceeds 50 annually (Lin and Wu, 1981; Yu and Sun, 1981). In Tibet, barley (Hordeum vulgare) grows at elevations up to 4700 m in the south, potatoes (Solanum tuberosum) and oats (Avena sativa) to 4500 m, and wheat (Triticum aestivum) to 4100 m; corn (Zea mays) grows up to 4000 m on the warmer plateaus (Cern´ık and Sekyra, 1969). Highlanders of Central and South America were the ﬁrst to grow potatoes, corn, squash (Cucurbita pepo) and beans (Phaseolus vulgaris), which perhaps made possible the expansion of other highland populations, e.g., potatoes in the Himalaya (Price, 1981). The diversity of mountain environments is reﬂected in the numerous varieties of mountain crops, which are endangered by current population and other pressures (Zimmerer, 1992). Sedentary societies in more densely populated mountainous areas (like the European Alps) combine farming at lower elevations with hay production and transhumance higher up (Price, 1981; Werner, 1981; Ellenberg, 1986). Farmers use fences to restrict domestic grazers to summer allotments of communally owned highlands. Animals spend snowy winters in barns at lower elevations eating stored feed, often hay from lower-elevation meadows that are not grazed, or from slopes too steep to pasture. In Europe, domestic animals control many alpine ecosystems and meadows cleared at the tree-line, all of which are now populated by weedy plants (Fenton, 1937, 1940). Lowering of the tree-line has increased avalanches, landslides, and erosion (Stern, 1983; Holtmeier, 1989), which are decreasing again now that some areas have been reforested (Bunza, 1984; Grabherr, 1997). Mountain farming and altitudinal movements of livestock are now declining, particularly in countries other than Switzerland, Germany, and Austria (Wielgolaski, 1975, 1978; Price, 1981; Greif and Schwackh¨ofer, 1983). Except for brief periods, the use of the Rocky Mountains by domestic animals was seldom uniform or intensive (Paulsen, 1960; Marr, 1964).

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Fig. 3.15. Crop agriculture in a lower-alpine river valley near the border with Nepal, above the village Zhong Zhang, Tibet. Terraced ﬁelds are limited to the relatively moist and protected river valley, and bordered by walls built of stones picked out of the ﬁelds.

Low-level, traditional grazing of domestic animals, which does not lead to vegetation changes, is the only sustainable agricultural use of the productive resilient margins of the cold regions. On highland plateaus and in the Arctic, both wild and domesticated yak (Bos grunniens), sheep (Ovis spp.), goats, antelopes, deer, llama (Lama glama), alpaca (Lama pacos), guanicoe (Lama guanicoe), vicu˜na, caribou and reindeer harvest the sparse vegetation over large plains unsuitable for crops (Ekvall, 1968). Herders build up the herds as insurance against weather or disease (Ekvall, 1968). Some of the animals have been interbred with their more productive, lowland domestic relatives which are unable to live at high altitude, and the few remaining wild populations are in danger of extinction and genepool changes, e.g., hybrids between yak and cattle. On the Tibetan Plateau, complex herd movements and pasture rotation prevent overuse, some pastures being reserved only for dry or cold season use or for hay. Overgrazing, erosion, and vegetation changes occur along migration routes and semi-permanent camp sites (Fig. 3.16). Long-term overgrazing of winter pastures leads to pasture degradation and loss of weight in animals (Xia, 1986; Zuo et al., 1986). Severe erosion (Fig. 3.17; King and Jackson, 1982) is usually related to increased population pressure from the lowlands, to military exclosures, or to past

attempts of the Chinese authorities to modify the native system by instituting collective ownership of animals and changing the pasture rotation (Goldstein and Beall, 1989, 1990; Goldstein et al., 1990; Reiter, 1991). This drives pastoralists up higher and puts greater pressure on summer pastures (Tsering, 1992), where weedy taxa replace the dominant sedges (Zuo and Le, 1980). From their recent maximum extent, high-altitude grasslands in Qinghai province had decreased by about one-ﬁfth (7.3 million hectares) by 1986 because of overgrazing, desertiﬁcation, and conversion to crop ﬁelds (Chinese sources quoted by Smil, 1993). In the Andean p´aramo, grazers overturn stem rosette plants by rubbing (Hofstede et al., 1995). Grazing is often combined with burning of natural vegetation, which has low productivity, turnover rate, belowground phytomass, and nutrient availability, and contains up to 80% of easily burnable standing-dead phytomass. Burns are colonized by palatable grasses, which replace stem rosettes not consumed by cattle (Grubb, 1970; Williamson et al., 1986; Lægaard, 1992). Reliance on animal dung for heating and cooking removes fertilizer from pastures. In the Andes, dung of llamas and cattle is used for fuel, and sheep dung for fertilizer; a family requires 25 sheep and 75 llamas for maintenance (Price, 1981). Seasonal movements with animals are not necessary on warmer plateaus. In desert mountains,

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

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Fig. 3.16. Tibetan nomadic pastoralists in the Qilian Shan (Gangkar Chogley Namgyal) on the northern edge of the Tibetan Plateau. Most nomadic pastoralists have this home base near winter pastures where some family members stay, while others take yaks and sheep to higher-altitude summer pastures (up to 4800 m in the Himalaya), a few to several hundred kilometers distant. View toward Mt. Kan She Ka, Gangkar Chogley Namgyal (Nan Shan). The skyline ridge reaches an altitude of 5000 m, and the snow-line is about 4700 m. The lower alpine belt, covered primarily by Dasiphora fruticosa, has an upper boundary at about 4000 m. Most meadows are dominated by Kobresia humilis. Weedy vegetation in the foreground is controlled by dung fertilization and trampling by yaks.

Fig. 3.17. Overgrazed and trampled hillside along the road between Golmud and Lhasa, Tibet. Vegetation is probably dominated by the sedge Kobresia pygmaea. Possibly irreversible damage is seen on a steep slope where drought and erosion may prevent recovery even if overutilization ceases.

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they may follow the growth of vegetation emerging in storm-tracks. Centuries of reindeer browsing have diminished the area of Subarctic pine and birch forests in Eurasia (M¨uller-Wille, 1974; Emanuelsson, 1987). Nomadic pastoralism is disappearing in more developed countries, as pasturage and access to it shrink as a result of development. Reindeer herding by about 1600 Saami is declining in the Kola Peninsula (Kol’skiy Poluostrov; Doiban et al., 1992), and Saami in northern Europe seek employment in farming, handicraft, and other ﬁelds (National Foreign Assessment Center, 1978). Collectivization has diminished the family life of reindeer herders of Yakutia (Vitebsky, 1989), and reindeer milking has been declining in the Russian Arctic (Fondahl, 1989). Winter grazing by reindeer may cause damage (Helle and Aspi, 1983), especially to the lichens that are their main or sole winter food (Pulliainen, 1971; Virtala, 1992); lichen overgrazing is common (Kautto et al., 1986; Klein, 1987). In the European Arctic, reindeer cause damage by trampling, particularly along fencelines. Transportation and installations Transportation of humans (roads, railroads, tunnels, airports, landing strips, parking lots, ski lifts), materials and energy (oil or water pipelines and powerlines), and associated development (construction, housing, oil ﬁelds, mines, tourism) are all common disturbances at the margins of cold-region ecosystems. Many of these are accompanied by noise, pollution, and solid waste. In the mountains, water reservoirs permanently inundate valleys; they destroy habitats for plants and animals, change stream ﬂow (Messier et al., 1989), attract tourism and, like transportation corridors, restrict animal migration routes (Hanson, 1981; Smith and Cameron, 1985; Curatolo and Murphy, 1986). Drainage changes cause ponding, ice jamming, and ﬂooding that may lead to changes in land-forms and ecosystems (Thorhallsdottir, 1993). Power stations at lower elevations divert water from mountain streams (Wielgolaski, 1978). In Antarctica, penguin colonies have been destroyed and their predator skuas disturbed in order to free space for airﬁelds (T. Thomas, 1986; Culik et al., 1990; Hemmings, 1990) and research stations (Wilson et al., 1990). Vehicles and solid waste take over habitats of Ad´elie penguins and Antarctic skuas, which decline (Johnston, 1971; Hemmings, 1990). In mountains, buildings and transportation

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corridors must be protected by fences, walls, roofs, and other structures from avalanches, rockfalls, and landslides. Slope bases may have to be reinforced by concrete and metal supports, and bare surfaces covered and treated chemically to prevent erosion and collapse of the structures the slopes support (Schiechtl, 1988; Stvan et al., 1991). In areas with ground ice, alterations to the insulating vegetation and surface organic layer, and heat input from power plants, buildings, and pipelines, lead to alterations of ground heat balance, melting of ground ice, thermal erosion, reduced albedo, additional heat input from the atmosphere, subsidence of the ground and collapse of conventionally-built structures supported by it. Meltwater accumulates in depressions, creating new ponds (thermokarst). These processes continue in some places for decades (Bliss and Wein, 1972; Price et al., 1974; Rickard and Brown, 1974). To avoid such damage, support structures for bridges, and pipelines for water supply, fuel, and sanitary purposes, are insulated and heated, or placed in heatdissipating, refrigerated structures, to keep the contents from freezing, or from thawing the ground ice around them (Hwang, 1976). After construction, continuing reclamation may be needed to prevent further spreading of terrain instability and environmental degradation (Wishart, 1988). Dirt and gravel roads must be elevated and graded continually, as paved surfaces quickly deteriorate; they are sources of dust and of spreading disturbances (National Foreign Assessment Center, 1978; Klinger et al., 1983; Hayley, 1988). In some environments, even one passage of an animal, a human, or a vehicle, whether during summer, or in the winter on roads and runways built on snow and ice, is sufﬁcient to cause damage; most of the damage occurs during the ﬁrst few passes (Ahlstrand and Racine, 1993). Both damage to dry surfaces resulting from trafﬁc, and compression of the wet organicsoil surface layer in marshes, may be permanent (Hernandez, 1973). In polar deserts, less-weathered, angular soil material is brought to the surface, and weathered stones are overturned, leaving impressions (Campbell et al., 1993, 1998). Repeated compression on paths (Pounder, 1985) and vehicle trails, even by off-road vehicles (Slaughter et al., 1990; Forbes, 1992), crush and compact vegetation and the organic surfacesoil layer; both are destroyed by surface removal and subsurface compression. Results may include increased runoff, erosion, mass movement, reduction in taxonomic richness and permanent changes in the

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

ecosystems (Challinor and Gersper, 1975; Gersper and Challinor, 1975; Bryan, 1977). Large-scale research using heavy equipment can be devastating, especially in marshes and polar deserts (Parker and Howard, 1977; Campbell et al., 1993). Extraction of minerals and fossil fuel Crystal collecting was one of the early human activities in cold regions. The effects of early mining were limited; in some areas, such as Tibet, it was banned for religious reasons. Primitive surface or shaft mining (Tibet, Bolivia, Rocky Mountains) or panning for gold (Tibet, Alaska) are still being carried out by natives, immigrants, or small-scale prospectors. Alpine mine spoils, roads, and solid waste are common in the Colorado Rocky Mountains, where mining for lead, silver, gold, tungsten, and other metals, and logging for mine supports, took place during the late 19th century. Some other mountains contain richly mineralized rocks, but large-scale mining operations have mostly been limited to accessible lower altitudes and to the Arctic. The Arctic is an important source of minerals and fossil fuels. The Kiruna mine in Swedish Lappland, which has been worked since the late 19th century, produces more than 25×106 tonnes of iron ore annually. Norway has a northern iron-ore mine (Kirken¨as, about 70ºN). Both Norway and Russia mine coal in Svalbard; the Norwegian mining is currently decreasing. Lead, zinc, and other metals (Sinding and Poole, 1991) are being mined in Greenland. Arctic sources account for 2% of Canadian mineral output. Estimates of Alaska’s coal reserves range as high as 9×1012 tons. Oil exploration there started about 60 years ago (Reed, 1958; Gryc, 1985), and a small natural gas ﬁeld near Barrow has been producing local fuel since 1955. Alaskan oil production rose sharply in 1977 with the completion of the trans-Alaska pipeline from the Prudhoe Bay ﬁeld. Most Arctic continental shelves may bear oil and gas (Rudkin, 1974; National Foreign Assessment Center, 1978; Atlas, 1985). In the Russian Arctic, the Kola Peninsula contains the world’s largest apatite deposit (2×109 tons), used for phosphate fertilizer. Two major metallurgical centers use copper–nickel ores, and there are many other mineral deposits and metal byproducts. Iron ore reserves may reach 109 tons. The Pechora Basin produces 107 tons of coal annually. Numerous mines, smelters, and reﬁneries near the mouth of the Yenisey

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River yield more than half of the Russian production of nickel and a substantial part of its copper. Eastern Siberia supplies more than 90% of the Russian output of diamonds and tin, and the Far East supplies coal and mercury. Most of Russia’s gas and some oil reserves are in the Arctic, some offshore (National Foreign Assessment Center, 1978; Vitebsky, 1990b; Doiban et al., 1992). Mining leads to ecosystem destruction, accumulation of unweathered and potentially toxic overburden or spoils, and pollution by acids and heavy metals (Cooke, Chapter 14, this volume), which can be spread in mine efﬂuents, leached out by water, or scattered by wind. In the Kola Peninsula, irreparable environmental damage surrounds mining and smelting operations, advancing at a rate of 2 km per year (Doiban et al., 1992; Rees and Kapitsa, 1994); see also Fig. 3.14. Oil and gas ﬁelds, offshore drilling platforms, pipelines, tankers, and other oil and gas transport and processing are sources of oil spills, oil and gas ﬁres, and other forms of oil pollution. Back-up systems for containing oil from leaking storage tanks are required (Mackay, 1985). Drilling muds and concrete help to prevent freezing and thawing, and drill casing has to be insulated to prevent ground collapse. Underwater structures are endangered by bottom scouring from ice ridges and icebergs unless well heads are countersunk in the sea-ﬂoor and protected by silos (National Foreign Assessment Center, 1978). Solid waste Solid waste is scattered even through some of the most remote cold regions, because its removal is expensive and seldom carried out, and its distintegration is slow in the cold conditions. Current solid waste is similar to that at lower latitudes and altitudes, except that wood is also an exotic material. On land, it commonly includes gravel roads and pads, machinery, supports, pilings, and concrete abandoned at mining and drilling sites, barrels, camp remains, food remnants, coal, plastic, metals, and glass. Solid waste covers vegetation, killing it along with other organisms, and buries the soil. It creates new habitats (Gregory, 1987, 1990), and may change ecosystems in its vicinity by shading the ground, accumulating moisture and snow (Kershaw, 1991). Solid waste may disintegrate into, or leak, exotic or polluting chemicals. At Antarctic stations, sites for disposal of solid waste are often a source of such chemicals (Harris, 1991; Rodgers et al., 1991,

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1992), as may be sites in the Arctic where solid waste is buried (Schindler, 1983). Sites for the permanent disposal of nuclear-fuel waste must be able to withstand continental glaciations (Sheppard et al., 1995). Plastic and other garbage ﬂoating in cold-region surface ocean waters, from both distant and local sources, is often washed ashore (Johnston, 1971; Gregory et al., 1984; Merrell, 1980, 1984), and it may serve as a means of dispersal for invaders such as rats (Gregory, 1990). Seals and other marine animals die from being entangled in ﬁshing nets, packing bands, twine, and other debris (Bonner and McCann, 1982; Ryan, 1987; Croxall et al., 1990), or from being poisoned by toxic chemicals from plastics (Ryan et al., 1988). Pribilof Island fur seals (Callorhinus ursinus) suffer a 5% mortality rate from entanglement in plastic debris (Reed et al., 1989). Birds may die by striking aerial arrays in ﬂight, or from choking on humandiscarded objects. On glaciers, solid waste is being slowly incorporated into ice. Aluminum survey markers drilled into the surface of a blue iceﬁeld were still standing 34 years later (Brunk and Staiger, 1986; Swithinbank, 1993). Pollution Most airborne pollutants in cold regions are anthropogenic (Boutron and Patterson, 1987), carried from lower latitudes and altitudes by variable winds and ocean currents (Holdgate and Wace, 1961; Llano, 1967; Nriagu, 1978). Winds and currents also reexport some allochthonous pollutants, and export some autochthonous materials, which are relativelly scarce. Natural autochthonous pollutants include volcano emissions, e.g., Mount Erebus, Deception Island (Siegel et al., 1980; Bargagli et al., 1993), aerosol particles produced by oxidation of reduced sulfur gases produced by marine microorganisms (Gras, 1992), and bromine compounds produced by ice algae following the loss of surface ozone (Oltmans and Komhyr, 1976; Berg et al., 1984; Barrie et al., 1988), which affects most cold regions (Fehsenfeld et al., 1983; Jaffe, 1991; Oltmans, 1991, 1993). Sulfur compounds released by marine algae to the atmosphere can be oxidized there to sulfur dioxide and other compounds, forming aerosol particles serving as cloud condensation nuclei (Gibson et al., 1989; Curran and Burton, 1992). High diversity of cold regions is reﬂected both in the amounts of incoming pollutants, which may be modiﬁed by local wind patterns, and in the

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degree of pollutant effects, which may be modiﬁed by local conditions such as bedrock chemistry, insolation, distribution of organisms, or environmental and population ﬂuctuations (Dunbar, 1985). Orographic uplift or scavenging may deliver relatively large amounts of some pollutants, such as acid rain, to windward mountain slopes (Parrish et al., 1986a–c; Oltmans and Levy, 1994), while downwind areas may be unaffected. Pollutant concentrations are usually lower in the more remote south than in the north (Bennington et al., 1975; Risebrough, 1977; Schneider et al., 1985), where pollutants, mainly from Europe and Asia, create haze which persists between November and May (Rahn and McCaffrey, 1980; Rahn, 1981; Rasmussen et al., 1983). Surface snow in Greenland contains about 200 times more lead than prehistoric Greenland ice (Murozumi et al., 1969; Ng and Patterson, 1981), while the Antarctic snow contains only 2–3 times more lead than prehistoric Antarctic ice (Boutron and Patterson, 1983). Trace metals that occur in cold-region environments and organisms are usually correlated with wind, current, and migration patterns (Bohn and McElroy, 1976; Bohn and Fallis, 1978; Boutron and Lorius, 1979). Mercury bioaccumulates in polar bears (Eaton and Farant, 1982; Norstrom et al., 1988; Dietz et al., 1990) and their prey (Smith and Armstrong, 1975, 1978). Mercury content in lichens of Victoria Land was surprisingly high (Bargagli et al., 1993). Antarctic seawater contains high levels of lead (Lee et al., 1990), which is, along with mercury, also present in native peoples in the Arctic (Charlebois, 1978; Grandjean, 1989), as well as other organisms. Radioactive fallout from atmospheric explosions of nuclear weapons (from 1945 until the 1970s) and nuclear accidents (the Chernobyl power plant in 1986) is accumulated by slow-growing organisms of the cold regions, especially by fruticose lichens and cushion plants (Rickard et al., 1965; Hanson, 1967, 1971, 1982). Cs137 increases by a factor of two for each higher trophic level; its concentration in human consumers of reindeer meat, which may be over 100 times higher than in other people (Hanson, 1967, 1968; Bird, 1968), could be responsible for high cancer rates among northern natives in Russia (Vitebsky, 1990c). The areas closer to the sources are more contaminated than distant areas (Baeza et al., 1994) and extreme latitudes (Hardy et al., 1968; HutchisonBenson et al., 1985; Taylor et al., 1985).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

The atmospheric content of nitrous oxide, halocarbons (Hall et al., 1990), anthropogenic gaseous chlorine, other trace gases (Rasmussen et al., 1981; Khalil and Rasmussen, 1985–1987) and air pollutants (Molski et al., 1981) have been rising both at the South Pole and other locations. Adsorption of organic vapors of high molecular weight to particulate matter in the atmosphere is enhanced by low temperatures (Bidleman, 1988; Bidleman et al., 1989; Focardi et al., 1991). Dichloro-diphenyl-trichloroethane (DDT) and other organochlorine pesticides and their derivatives are present in air, water, snow, ice, and sediments, and some are preferentially accumulated in the lipids in organisms. Concentrations of some have been declining, probably in response to the discontinuation of their use in some countries (Sladen et al., 1966a; George and Frear, 1966; Brewerton, 1969). Exudates from brown algae may alter the bioavailability of polychlorinated biphenyls (PCBs) or other compounds to which they bind (Lara et al., 1989). Organic runoff from animal husbandry and warm water from power plants may lead to eutrophication of water reservoirs (Doiban et al., 1992). Sewage can contaminate limited water supplies, especially in areas with ground ice or shallow weathering. Chemical fertilizers and insecticides used in river valleys on the Tibetan Plateau lower soil and water quality (Goldstein and Beall, 1989; Tsering, 1992). Pollution and chemical changes in environment and organisms are also caused by emissions from cars, boats, ships, generators and aircraft, concrete batching plants (Adamson and Seppelt, 1990; Roser et al., 1992; Adamson et al., 1994), waste water from research stations (Risebrough et al., 1990; Harris, 1991; Railsback, 1992), municipal waste (Bourgoin and Risk, 1987; Haertling, 1989) and exotic bacteria, fungi, and viruses (Toyoda et al., 1985, 1986). Dust from roads, cement, or long-range transport causes earlier snowmelt, changes soil chemistry, moisture, and nutrient mineralization, reduces primary production, and causes changes in the composition of communities (Bodhaine et al., 1981; Spatt and Miller, 1981; Walker and Everett, 1987). Lead in batteries, anti-freeze compounds, or other wastes may poison animals such as skuas (Catharacta spp.), sheathbills (Chionis spp.), and polar bears, which then spread the toxins in their carcasses (Johnston, 1971; Lunn and Stirling, 1985; Amstrup, 1989). In Arctic Russia, mining and military wastes include acids, heavy metals (mainly nickel, copper, manganese, zinc), radioactive materials, oil, and organochlorines. In

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the Kola Peninsula, annual wastes include 410×106 m3 of efﬂuents, 6.5×105 tons of atmospheric pollutants (mainly SO2 ), and 55×106 tons of mining solids. Concentrations of heavy metals in the top soil are increasing, lakes are mineralized, and the Murmansk harbor is contaminated by oil and phenols. Heavily contaminated areas, in which vegetation has been completely destroyed, are spreading, and pollution is reaching to Finland and Norway (Fig. 3.14). Between 1980 and 1990, total emissions of sulfur dioxide were reduced by 18%, mainly through technological modernization (Freydin, 1972; Strelkov and Freydin, 1973; Rees and Kapitsa, 1994). Increases in ultraviolet-B radiation A thinning of the ultraviolet-absorbing stratospheric ozone layer, caused by chlorine pollutants in the atmosphere originating primarily in the Northern Hemisphere, is most serious in the polar regions despite their relatively large total stratospheric column of ozone during the periods of maximum radiation. The lack of denitriﬁcation in the northern polar vortex currently limits the loss of ozone over the Arctic, but future cooling of the lower stratosphere could increase it (Santee et al., 1995). Some volcanic and other aerosols enhancing ozone depletion (Hofmann et al., 1992; Hofmann and Oltmans, 1993; Tsitas and Yung, 1996) originate in cold regions. The largest ozone “hole” (up to 60% loss) develops in Antarctica in the spring, because the catalytic reactions of anthropogenic pollutants on the surface of ice crystals in polar stratospheric clouds are promoted in this coldest region of the atmosphere. Long days, lack of heavy cloud cover, and high elevation contribute to high monthly totals of ultraviolet radiation reaching the surface at the South Pole during midsummer, despite the high angle of incidence and truncation of the shorter, biologically most active wavelengths at the poles. A small percentage of UV-B may even reach depths of 60 m in clear ocean water, and penetrate clear sea ice, but not snow, which protects some organisms during the greatest development of the ozone “hole”. During that time in open water, nutrients may be low at the surface, and herbivores may be at depths where ultraviolet irradiation is low. Ozone destruction is thought likely to intensify and spread even as the emissions of pollutants decrease, because their transfer into the stratosphere takes decades (Caldwell et al., 1980; Baker-Blocker et al., 1984; Farman et al., 1985).

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Increased levels of UV-B radiation may cause mutagenesis, an increase in photoinhibition and photooxidative destruction of chlorophyll, depression of physiological processes such as photosynthesis, other detrimental changes (such as reduction in mobility: Ekelund, 1992), and may result in death. There may be 4% or more reduction in primary production and nutrient ﬂux as a result of enhanced UV-B radiation. This radiation is damaging during spring bloom when organisms with little photoadaptation approach the ocean surface to take advantage of higher light levels. Many aquatic herbivores and other organisms with little protection (for their eyes and other organs) are sensitive to UV-B radiation and to the potentially toxic products of its reactions with pollutant chemicals. However, increased photooxidation of toxic chemicals could stimulate aquatic productivity. A shift toward dominance by protected native organisms or immigrants could cause changes at higher trophic levels (Calkins and Thordardottir, 1980; Adamson et al., 1988a; El-Sayed et al., 1990). Irradiation by UV-B could also affect the terrestrial alga Prasiola crispa, which is poorly protected (Adamson and Adamson, 1992). Ultraviolet radiation decreases the cell-mediated cutaneous immune response in humans in Antarctica (Williams et al., 1986; Muller et al., 1988; Roy et al., 1994). Increase in carbon dioxide and climatic change The current climatic warming and lengthening of the growing season have been linked to an atmospheric increase in the concentration of trace gases (carbon dioxide, methane, nitrous oxide, etc.) originating at lower latitudes (Barnola et al., 1987; Etheridge et al., 1988; Raynaud et al., 1988). This increase may be enhancing natural climatic cycles related to ocean circulation (Bond et al., 1997; Broecker, 1997), and causing increased solar irradiance, decreased volcanic activity, and internal climate feedbacks (Overpeck et al., 1997). A decrease in ozone in the lower stratosphere and in tropospheric aerosols may slow down the warming. Continuing warming may be accompanied by changes in air and water circulation and by the movement of latitudinal atmospheric and oceanic boundaries poleward. The changes in oceanic circulation might, however, cause a sudden cooling (Kerr, 1998). Temperature increases will be enhanced at high latitudes (up to 5ºC in summer and 10ºC in winter) through

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positive feedback, to which increased evaporation and cloudiness may contribute. Stratospheric temperatures will decrease, and the temperature of warm air at the surface will rise, as it will be conﬁned by the stable cold polar air above. The amount of heat lost to space will decrease, because sea ice, which reﬂects most of radiation insulates the ocean surface and prevents evaporation, will decrease in extent; subsequently, higher temperatures of ocean water will reduce the deep ocean sink of carbon dioxide (Manabe and Wetherald, 1975, 1980; Kelly et al., 1982). Increases in the average surface temperatures and a net decrease in the mean extent of sea ice (0.23 degrees of latitude per decade between 1973 and 1988) have been observed at most Antarctic and Southern Ocean stations (Raper et al. (1984)), Heard Island (Allison and Keage, 1986; Jones et al., 1986), Macquarie Island (Adamson et al., 1988b; Gloersen and Campbell, 1988; Jones, 1988, 1990) and in the Arctic (Garﬁnkel and Brubaker, 1980; Chapman and Walsh, 1993; Oechel et al., 1993). Satellite observations indicate a decrease in the areal extent of sea ice by 2.9±0.4 per cent per decade in the Arctic, and an increase by 1.3±0.2 per cent per decade in the Antarctic (Cavalieri et al., 1997). A cooling trend could be occurring in other areas, e.g., in the Rocky Mountains in Colorado, by 0.07ºC per year (Barry, 1973; Greenland, 1989), and in Fennoscandia (Kullman, 1993). The effects of warming are evident in the melting and breakup of polar ice shelves (Orheim, 1988; Zakharov, 1988; Doake and Vaughan, 1991), in the retreat and thinning of many glaciers, e.g., in the Alps, the Arctic (Calkin, 1988; Koerner and Fisher, 1990), Central Asia, New Zealand, the Subantarctic, northern and West Antarctica (Rignot, 1998) and Patagonia (Aniya et al., 1997), and in the warming and melting of Arctic and alpine permafrost (Lachenbruch and Marshall, 1986; Kane et al., 1991b; Haeberli et al., 1993). The reduction in global glacier area due to retreat is calculated as 6–8×103 km2 , and the mass balance has generally been negative from 1961 to 1990 (Dyurgerov and Meier, 1997a,b). Newly deglaciated surfaces may be subject to widespread erosion (National Foreign Assessment Center, 1978; Koster, 1991; Ballantyne and Benn, 1994). Increased runoff during melting may increase the frequency of landslides and failures of moraines and ice-dams (Mahaney, 1986), and contribute to the spreading of pollutants (Kane et al., 1992; Sheppard et al., 1995). Isostatic rebound of the underlying land masses, following melting of

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

the Antarctic and Greenland ice sheets, could result in the elevation of low-lying coasts and islands in polar regions; but worldwide most coastal land masses would be drowned, as the rise in sea level due to thermal expansion would be substantial (Frei et al., 1988). The current warming trend may be decreasing the amplitude, frequency, and duration of precipitation events and storms in some areas and during some seasons (Manabe and Wetherald, 1986, 1987). In such areas, lakes, streams, snow-patches and glaciers are receding, and droughts and ﬁres may become more frequent. Average precipitation is decreasing on the western part of the Tibetan Plateau, where the desertiﬁcation trend is aggravated by the logging of the adjacent forests (Reiter, 1991). A slow decrease of soil moisture has been established in the Rocky Mountains in Colorado (Taylor and Seastedt, 1994). In other areas, precipitation and stream volume may be increasing due to greater evaporation, evapotranspiration, and a higher rate of water cycling, e.g., in the Antarctic (Morgan et al., 1991), and in oceanic areas of northern Europe (Hagen, 1996). Mean annual temperature is decreasing and annual precipitation is increasing in the Rocky Mountains in Colorado (Barry, 1973; Greenland, 1989), where the increase in precipitation (14 mm yr−1 ) explains about half of the 200% increase in annual wet deposition of nitrate over the last decade (Williams et al., 1996). In some areas, snow may turn to rain. Some glaciers, e.g., in Norway, Greenland (Zwally, 1989; Zwally et al., 1989), the Arctic, Iceland and the Antarctic (Morgan et al., 1991; Peel, 1992) have been growing during the last decade, probably through increased winter snowfall (Hagen, 1996). Increasing precipitation over the Antarctic, partly as a result of evaporation from areas previously covered by ice, could lead to lowering of the sea level (Robin, 1986; Budd, 1988, 1991; Gloersen and Campbell, 1988). Acid rain Acid rain, caused by the solution of NOx and SOx emissions from power plants and cars in liquid water in the atmosphere, could lead to signiﬁcant changes in the nutrient and base content of shallow, poorly developed soils in the cold region (Lewis and Grant, 1980; Lewis, 1982; Haselwandter et al., 1983) and in the composition of the ecosystems that these soils support. So far, the effects of acid rain are not seen in remote cold regions, and are slight even in some cold regions near industrial

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centers. In mountains that have high rates of erosion and solution, a signiﬁcant proportion of the imported nutrients may be exported; in other areas, they may not be utilized due to drought. In remote headwater basins of the Urumqi River in the Tien Shan (China), the solute composition of surface water and snow is dominated by dissolution of rocks with rapid weathering kinetics, such as calcite and chromite (Williams et al., 1995a). In the Sierra Nevada in California, mineralization and nitriﬁcation produce more inorganic nitrogen than is received in wet and dry deposition (Williams et al., 1995b). Acid rain could, within a decade, lead to a widespread acidiﬁcation of surface water in the alpine zone of the Rocky Mountains in Colorado (Caine, 1995b), although increased snowfall may increase rates of decomposition, mineralization, and nitrate export in streams (Williams et al., 1998). The annual wet deposition of nitrate increased by 200% over the last decade and only half of this amount is explained by the increase in precipitation. The increase in precipitation and decreased energy ﬂux to the snowpack between 1994 and 1995 resulted in a 4- to 5-fold increase in the magnitude of solute release from the snowpack in the form of an ionic pulse, causing episodic acidiﬁcation in headwater catchments (Williams et al., 1996). The retention of nitrogen is nearly 1 mg m−2 per day, or 2.5– 3 kg ha−1 of new nitrogen annually, which approaches 50% of the net annual mineralization of nitrogen (Sievering et al., 1992, 1996). The annual inorganic loading in wet deposition of about 4 kg/ha/yr is about twice that on the Paciﬁc coast and similar to the northeast, leading to N saturation as a result of anthropogenically ﬁxed N in wetfall and dryfall (Williams et al., 1998). Hydrocarbon spills Cold slows down the rates of oil evaporation, dispersion, and oxidization. At high latitudes, severe weather makes the behavior of marine oil spills less predictable, and the use of chemical dispersants (by aerial application) or burning less predictable and effective. Oil spills may persist longer than in temperate or tropical ecosystems, and their fate and effects may be more diverse. Oil sticks to sea ice, and may become encapsulated in it, followed by release the following spring (Atlas and Bartha, 1972a,b; Haines and Atlas, 1982). Crude oils are usually less toxic at 15ºC than at 0ºC or 10ºC (Hsiao, 1978).

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Oil kills aquatic organisms, inhibits primary production, and changes the composition of communities in ponds and lakes (Miller et al., 1978; Federle et al., 1979). It also may reduce heterotrophic activity over a period, especially in sediments [for 18 months according to Grifﬁths et al. (1981a,b, 1982a,b)], depress cellulase activity (Linkins et al., 1978), and lower the rates of nitrogen ﬁxation (Grifﬁths et al., 1982b; Atlas, 1985). The acute lethal toxicity of oil dispersions and water-soluble fractions for zooplankton in waters of both the cold and the cold-temperate region is indicated by values of the 4-day LC50 ranging between 0.05 and 9.4 mg °−1 . Isopods seem to be more resistant to oil than amphipods and decapods, which often swim erratically before they die. Molluscs, other invertebrates, and ﬁsh become sluggish and disoriented, and suffer reduced growth, development and reproduction, and death (Percy, 1976, 1977; Busdosh and Atlas, 1977). Fish can avoid many oil spills, but natural stresses (ﬂuctuations of salinity, temperature, and food abundance, disease, parasites) increase their sensitivity to toxic compounds. Low concentrations of hydrocarbons in ﬁsh tissues make them unpalatable to humans (Rice, 1985). Oil damages bird embryos, and sticks to the skin, fur, and feathers of marine endotherms, adversely affecting thermoregulation and metabolism and causing loss of fur and feathers. Animals insulated by blubber are resistant to heat loss indiced by oil coatings. Animals attempting to clean their fur or plumage and those eating contaminated food will ingest oil. This may result in reproductive impairment, stunted growth, metabolic stress, anemia, and increased mortality (Kooyman et al., 1977; Peakall et al., 1980, 1981). Oil tends to be concentrated in ice leads and breathing holes, and wind piles it up against ice edges (Ayers et al., 1974; Percy and Wells, 1984). These places are frequented by marine mammals, which may be coated by the oil, inhale volatile hydrocarbons, or ingest contaminated food. Oil matted in baleen may reduce the ﬁltering efﬁciency of baleen whales, especially in cold water, where it is more viscous. Oil-covered seals have difﬁculty in swimming and may die of exhaustion, their breathing oriﬁces may be plugged by oil, and eye lesions may occur. Pathological tissue changes and death due to renal failure occurred in polar bears exposed to oil (Engelhardt, 1985). After the grounding of MV Bah´ıa Para´ıso at Anvers Island in Antarctica on 28 January 1989, adult skuas, penguins, and other birds were covered with oil, ﬂedglings were deserted or poisoned by oil, and

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intertidal fauna and ﬂora were damaged; other fuel spills have also damaged terrestrial ecosystems in the Antarctic and Subantarctic (Barinaga and Lindley, 1989; Karl, 1989, 1992). In 1977, the largest blowout in an offshore drilling ﬁeld at Ekoﬁsk in the North Sea (56º38 N, 3º12 E) spread 23 000 tons of crude oil over an area of 40 000 km2 , killing many seabirds but causing little damage to other marine organisms (Wielgolaski, 1989). On land, natural oil seeps and substances spilled during extraction, processing, or use of crude oil and its products stick to soil particles and kill or damage organisms, usually through their toxicity (Wein and Bliss, 1973b; Everett, 1978; Walker et al., 1978). Human disturbance through organismic invasions Within cold regions, the success of alien taxa and the complexity of successful taxa rapidly decrease with increasing latitude and altitude. Beyond cold-region margins, there are no successful exotic invaders, and the organisms present (mostly lower organisms such as lichens, but also some sedges) are common to different cold regions, especially the dominants among them. This relatively high resistance to invasions resembles that in tropical ecosystems, forests in particular, where high resistance to invasions may be due to both biotic and abiotic factors and does not seem to be related to taxonomic diversity (Rejm´anek, 1996a). Cold regions seem to be protected by their harsh environment and by the lack of taxa adapted to similar conditions elsewhere. Their protection may be mostly environmental, because they have few invaders despite their low taxonomic diversity and the low competitiveness of their few residents. At cold region margins, cosmopolitan weeds may occur on exotic materials (along gravel roads) but, unlike in temperate biomes, there are no cosmopolitan weed ecosystems maintained by trampling or other disturbances. Other alien taxa may colonize disturbances. Introduced plants may be common in animal-disturbed sites, e.g., Cerastium fontanum in rabbit-disturbed sites on Macquarie Island (Scott and Kirkpatrick, 1994); or outcompete or destroy native taxa, but again only at the margins of the cold regions (Corte, 1961; Longton, 1966; Walton and Smith, 1973). Most invasive taxa that colonize overgrazed areas and other disturbances are natives of cold regions (Druzhinina and Zharkova, 1979; Matveyeva, 1979, 1988). Such invasive taxa are characterized by rapid and efﬁcient reproduction,

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

dispersal, growth, development, and population growth (Rejm´anek, 1996b; Rejm´anek and Richardson, 1996). Organisms with certain strategies (Grime, 1979; Grime et al., 1988) or trophic or growth-form types probably never reached, or evolved in, isolated relatively warm Subantarctic islands, leaving an opening for alien organisms. Some of these organisms could have been spontaneous colonizers, but most of them probably were transported by humans (whalers, sealers, unsuccessful settlers, research personnel) or on ﬂoating solid waste, or escaped from shipwrecks. Microorganisms, plants, invertebrates, and vertebrates have become established in almost all island groups. Flightless invertebrates and plants, which had evolved in the absence of large predators and herbivores, were decimated by vigorous invaders on many islands (Costin and Moore, 1960; Wace, 1960; Holdgate and Wace, 1961). On South Georgia and the Iles Crozet, invaders constitute a high proportion of the ﬂora (Selkirk et al., 1990), and on some islands they have changed the predator–prey relationships and ecosystem processes (Tr´ehen et al., 1990; Chevrier et al., 1997). Examples of well-established destructive mammalian invaders include cats (Felis catus) and house mice (Mus musculus (house mouse)) on Marion Island (van Aarde, 1979, 1986; van Rensburg and Bester, 1988a), rabbits (Oryctolagus cuniculus) and other animals on Kerguelen (Dreux, 1974; Lesel and Derenne, 1975; Chapuis, 1988), rabbits (Costin and Moore, 1960; Scott, 1988) and house mice (Pye, 1993) on Macquarie Island, and rats (Rattus norvegicus) (Pye and Bonner, 1980) and reindeer (Rangifer tarandus) on South Georgia (Lindsay, 1973; Leader-Williams et al., 1981, 1987; Vogel et al., 1984). Following their introduction on Arctic islands, reindeer greatly reduced the biomass of lichens, their important winter food in areas free of snow or where snow is shallow in winter (Scheffer, 1951; Ouellet et al., 1993). Some invader populations decline eventually, primarily through the reduction of their food source (Scheffer, 1951; Klein, 1968; Caughley, 1970). In most cases, ecosystem recovery depends on artiﬁcial removal of the invader (Johnston, 1973; Watkins and Cooper, 1986; Leader-Williams et al., 1989) – for instance, of rabbits by myxomatosis (Brothers et al., 1982; Chapuis and Bouss`es, 1989; Chapuis et al., 1994), and cats by panleucopaenia and hunting (Van Rensburg, 1986; Van Rensburg et al., 1987; Van Rensburg and Bester, 1988b).

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Tourism and recreation Cold regions provide spectacular scenery, largely undisturbed nature, year-round snow for sports, recreation, and expeditionary travel, and include important religious sites. Visitors are becoming more numerous as leisure time, disposable income, and mobility increase, but mass tourism does not reach the remote parts that are difﬁcult and expensive to explore. Tourism and recreation have a capacity to provide revenues for the protection of popular sites from trampling, pollution, and other threats, for research, and for the potentially expensive protection of visitors, e.g., in the Antarctic International World Park (Potter, 1970; Schweickart, 1988). Many national parks and wilderness areas in the United States, and governments in countries where there are popular destinations for expeditions, issue permits to avoid overuse, ban roads and vehicles, provide public transportation and lodging outside parks, and pave the most popular trails (Wielgolaski, 1971, 1978). In more developed countries, sports facilities, overcrowding, high-rise buildings, noise, pollution, roads, and cars are overwhelming the valleys and are spreading to the slopes above (Tsuyuzaki, 1994). Bulldozing of ski runs may cause the most severe impact above the treeline (Cernusca, 1977; Schauer, 1981; Klug-P¨umpel, 1988). Damage by visitors includes litter, erosion along trails, trampling (Willard and Marr, 1970; Grabherr, 1982; Gibson, 1984), camps and campﬁres, wood gathering, snowmobile damage (Greller et al., 1974), effects on wildlife (Cederna and Lovari, 1983; Hamr, 1988), and species introduction. Tens of thousands of visitors come through some national parks in the United States every summer day, e.g., Rocky Mountain National Park (Marr and Willard, 1970), even if 95% of them visit only 2% of the total area (Ives, 1974). Similar problems are now arising in popular, accessible areas in less-developed countries, where tourists burden local food chains. In Antarctica and parts of Svalbard, tourism is less destructive than research because most of it is shipbased, but polar tourism and other visits are increasing (Reich, 1980; England, 1982; Harris, 1991). Between the late 1950s and 1992, about 39 000 tourists have visited Antarctica (Reich, 1980; Enzenbacher, 1992, 1993) and trampling is becoming a problem (Scott and Kirkpatrick, 1994). Airborne tourism has a potential for reducing impact (Boswall, 1986; Swithinbank, 1988), because fuel residues are dispersed, inland operations

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seldom encounter wildlife, and waste can be brought out, replacing spent fuel (Swithinbank, 1993). Visitor guidelines for polar regions have been developed (National Science Foundation, 1990; Stonehouse, 1990; Hall and Johnston, 1995), although they are not always followed (Davis, 1995).

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habits, e.g., bears (Ursus spp.) (Follmann and Hechtel, 1990). Seals occur in leads associated with off-shore drilling rigs, and polar bears are attracted to them when sea ice is completely frozen; this increases the likelihood of conﬂicts between bears and humans (Stirling, 1988). On land, bears are attracted to garbage dumps, where they feed (Lunn and Stirling, 1985).

Contact with wild animals Cold-region animals are disturbed by noise, by vehicles, and by human presence. Paciﬁc black brant geese (Branta bernicla nigricans) never habituate to helicopter noise and seek safety in water; they may be prevented from completing their molt by frequent overﬂights (Miller et al., 1994). The heart rate of adult penguins rose 15.5% near humans, and up to 350% near helicopters, which caused panic reactions from even 1500 m away (Thompson, 1977; Culik et al., 1990; Wilson et al., 1991; Swithinbank, 1993). Helicopters also caused their body temperature to rise (Sladen et al., 1966b). Human visits to penguin colonies are not sufﬁciently regular and frequent to lead to habituation (Culik et al., 1990), and the disruption can lead to declines in breeding success (Thompson, 1977; Wilson et al., 1990; Young, 1990) and populations size (T. Thomas, 1986; Jouventin and Weimerskirch, 1990). During a circumnavigation of Macquarie Island at an altitude of 250 m by a C-130 Hercules aircraft in 1990, 7000 king penguins (Aptenodytes patagonicus) stampeded and suffocated (Rounsevell and Binns, 1991). At Marion Island, overpasses by this aircraft caused king penguins to panic and run inland (Cooper et al., 1994). Aggressive and predatory skuas may be killed because they interfere with research or station operations, or with agriculture or rabbits on southern temperate and Subantarctic islands (Hemmings, 1990). Sealing and whaling may have helped to maintain a larger skua population on South Georgia than the island could support otherwise (L¨onnberg, 1906; Stonehouse, 1956; Hemmings, 1990). Skuas feed on station wastes, defend dumps, destroy petrels dazzled by station lights, may overwinter instead of ﬂying north, may be tolerant of people (Young, 1990), and their populations may increase (Trivelpiece and Volkman, 1982; Hemmings, 1990). Food wastes and pet animals may pass avian and other diseases and parasites to wild populations (Parmelee et al., 1979; Harris, 1991). At the edges of cold regions, food wastes may support invasions of rats (Rattus spp.) and gulls (Larus spp.). Human activities may lead to changes in animal movements and feeding

ECOSYSTEM RESPONSES TO DISTURBANCE

Steep environmental gradients, environmental ﬂuctuations, and high levels of stress and disturbance produce greater diversity of ecosystems in cold regions than within most temperate biomes (Fig. 3.18). It could even be argued that cold regions encompass coldclimate variations of many other biomes (scrublands, grasslands, deserts), which have more in common with their temperate variants than with most extensive cold-region ecosystems (snow and ice, rocks). The level of surface disturbance reﬂects the level of cold stress, the availability of water, gravity, and movement-prone or water-accumulating surﬁcial materials. Most disturbances break up large patches into smaller ones, increasing patch heterogeneity. The new patches support a greater variety of environmental and organism combinations, but even exotic disturbances only exceptionally produce exotic ecosystems. Areas with more frequent surface disturbances have greater ecosystem diversity than those less frequently disturbed (Kom´arkov´a, 1993); see Fig. 3.19. Damage to populations and ecosystems increases and the probability of ecosystem recovery decreases with increasing intensity, scope, duration, diversity, predictability, and frequency of disturbances. The rate and direction of recovery are dependent on the characteristics of the disturbance and the responding unit (Vitousek et al., 1981). Some disturbances may short-circuit or eliminate certain ecosystem pathways, and others may accelerate certain processes (Lugo and Snedaker, 1974). This disturbance-speciﬁcity of ecosystem response is important where ecosystems are in equilibrium with particular disturbances, especially where the probability of disturbance increases with the time since its last occurrence (Holling, 1981; Vitousek et al., 1981). Recovery does not occur if disturbance intensity exceeds certain thresholds (D.M. Raup, 1981; Auerback, 1981; Westman, 1986). Such thresholds decrease as altitude and latitude increase, and the time needed

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

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Fig. 3.18. Cold-region landscapes ranked by disturbance frequency and patch size. At high latitudes and altitudes, year-round temperatures below 0ºC limit surface disturbances by ice and ecosystem diversity in areas with little water, such as polar deserts, most of which are deﬂated by wind. Snow and glaciers cover and abrade areas at high latitude and altitude with enough precipitation. Snow and ice ecosystems do not undergo freeze/thaw cycles as they do at lower altitudes or latitudes. Large energy transfers occur at middle altitudes and latitudes, where temperature ﬂuctuations include frequent passes through 0ºC and changes of state of water. Wet, relatively warm mountains with steep, long slopes, soft bedrock, and high runoff have probably the highest disturbance levels. Flat, waterlogged plains with loose surface materials and large amounts of ground ice and seasonal freeze/thaw cycles are also frequently disturbed, as are ice-scoured coasts and now deglaciated mountains with free-fall rock faces. Dry, gently sloping mountains and dry plateaus are much less disturbed. The least disturbed are the well-vegetated areas in cold-region margins with shrub, tussock, and turf ecosystems.

for recovery increases. They are speciﬁc for different ecosystem types (Fig. 3.20). Because organisms seem to have relatively weak control of cold-region ecosystems, these thresholds may be located at lower levels of disturbance intensity in cold than in temperate regions. Resistance and resilience Specialized structures and biochemistry enable most organisms resident in the cold regions to tolerate freezing, pollution, and other disturbances which affect functions more than structures. However, short growing seasons, cold stress, and low availability of nutrients lead to slow growth, metabolism, and maturation (Clarke, 1980; K¨orner and Larcher, 1988; K¨orner, 1994 ), and reproduction often fails, as result of environmental stress and ﬂuctuations along extreme

physical gradients. Most resident organisms, therefore, are poor at resisting disturbances and repair damage slowly, and avoidance of stress and disturbance is perhaps more important in cold regions than in many other biomes. Ecosystems, most of which also have poorly developed soils, are also less resistant and resilient (“fragility”; Webber and Ives, 1978) to disturbance than ecosystems in most other biomes. However, when they contain uniquely adapted organisms, they eventually recover to the pre-disturbance state, unless an environmental change intervenes during their prolonged recovery. Resistance and resilience increase with increasing inputs of heat, moisture, and nutrients, and the most resistant and most resilient ecosystems occur in the relatively warm, moist marginal areas, where the environment is relatively favorable to plant production, reproduction, and growth, and supports relatively large

72

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Fig. 3.19. In areas of approximately similar sizes and vascular plant diversities, Braun-Blanquet types of mature vegetation unaffected by human disturbance are more numerous and deﬁned at lower similarity levels in the more frequently disturbed Arctic Coastal Plain at Atqasuk, Alaska (A) than in the less frequently disturbed glaciated mountains of the Front Range of the Rocky Mountains in Colorado (B). From Kom´arkov´a (1993).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

73

Fig. 3.20. The relationship between resistance, resilience, and recovery in some cold-region ecosystems.

amounts of living organic matter and well-developed soils. The terrestrial ecosystems most resistant to mechanical disturbance (mesic scrub, tussocks, turfs) have a relatively large amount of tough, sclerenchymatous phytomass, silica-rich or woody, forming a more or less closed cover of slow growing plants which are relatively strong competitors. Thick, welldeveloped soil, hard rocks, ﬂat surfaces, and consolidated, structured surface materials increase resistance. Belowground storage, stem ﬂexibility, low growth form, and summer hardening diminish the injuries caused by aboveground mechanical disturbance, but their effectiveness is limited; after 150 passes of offroad vehicles in an Alaskan shrub tussock tundra, there was little difference in the degree of compression between any of the shrub taxa (Ahlstrand and Racine, 1993). Resistance seems to be unrelated to the number of taxa; it mainly reﬂects their properties. The relatively low resilience of these ecosystems reﬂects the slow rates of growth and organic-matter accumulation by the dominants, and the numerous small, slow-growing, drought- and freeze-tolerant organisms, such as lichens, that live between them. Resilience may be inversely related to the number of taxa and to ecosystem complexity. Frequently disturbed, simple ecosystems (marshes, shores, coasts, sand dunes) have low resistance but

high resilience, even at relatively high latitudes and altitudes. Usually, the few vascular plant taxa present are composed of soft tissues, reproduce vegetatively, and disperse, establish, and grow rapidly, some from belowground or near-ground storage reserves. They have wide ecological ranges (Kom´arkov´a, 1979), high nitrogen storage per unit leaf area (K¨orner, 1989), and highly resilient growth forms (Z¨ottl, 1951; Jefferies, 1988; Kom´arkov´a and McKendrick, 1988), enabling them to respond to disturbances rapidly even under cold stress. Landforms and soils are loose and easily colonized. Such ecosystems recover to almost their original state in a relatively short time. In Arctic Alaska, the large amount of water in the marshy landscape and frequent surface disturbances are perhaps the reasons for a high proportion of rhizomatous plants adapted for rapid vegetative reproduction, colonization, and growth (Tables 3.1A and 3.1B; Kom´arkov´a and McKendrick, 1988). Change and recovery The rate and direction of recovery are dependent on the characteristics of the disturbance and of the responding unit (Vitousek et al., 1981). To a greater degree than in temperate biomes, recovery after disturbance in cold regions is controlled by steep environmental gradients,

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

74 Table 3.1A Proportions of vegetation growth forms in three Arctic locations 1,2 Site

Growth form 3

Latitude evergreen shrubs

rhizomatous graminoids

caespitose graminoids

prostrate forbs

0

0

23.4

5.5

33.4

6.8

30.7

deciduous large dwarf shrubs shrubs

erect forbs cushion others

Edgeøya 4

78º05 N

Barrow 5

71º18 N

1.7

4.5

1.7

46.4

7.2

7.8

0.2

30.3

Atqasuk

70º29 N

14.6

15.0

7.8

31.1

6.9

2.9

0.8

17.4

0

1 Rhizomatous growth forms seem to have greater importance in the more frequently disturbed landscapes than in the less frequently disturbed ones. 2 After Kom´ arkov´a and McKendrick (1988). For each growth form, the importance of vegatation cover is expressed as a percentage of the total for all vascular plants at that site. 3 The deﬁnitions of growth forms approximate those in Barkman’s system (Barkman, 1983, 1987). 4 Data from Barkman (1987). 5 Data from Webber et al. (1980).

Table 3.1B Proportions of vegetation growth forms in different latitudinal belts 1,2 Site

Growth form 3

Latitude deciduous shrubs

evergreen shrubs

rhizomatous graminoids

caespitose graminoids

prostrate forbs

erect forbs cushion others

Northern Arctic Edgeøya, Spitsbergen

78º05 N

0

0

23.4

5.5

33.4

6.8

30.7

Devon Island, Canada

75º33 N

5.0

7.8

46.1

11.1

9.4

11.8

8.1

Tareya, Russia

73º30 N

7.0

2.3

28.5

21.2

2.0

27.0

11.7

Barrow, Alaska

71º18 N

9.8

2.2

48.3

9.4

6.0

0.3

24.0

Atqasuk, Alaska

70º29 N

29.6

7.8

31.1

6.9

2.9

0.8

17.4

Disko Island, Greenland

69º15 N

28.4

2.3

2.6

16.2

6.0

12.6

28.6

68º22 N

9.3

22.4

35.1

12.0

0

0

21.1

54º12 N

1.1

35.5

11.3

35.0

2.2

0.4

13.5

54–55ºS

31.8

0

16.8

49.6

1.8

0

0

0

0

0

91.1

0

8.9

0

Southern Arctic

Subarctic Abisko, Sweden Temperate Glenamoy, Ireland Northern Antarctic South Georgia Southern Antarctic Signy Island

60ºS

1 Rhizomatous growth forms seem to have greater importance in the more frequently disturbed landscapes than in the less frequently disturbed ones. 2 Data from French (1981). 3 The deﬁnitions of growth forms approximate those in Barkman’s system (Barkman, 1983, 1987).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

frequent environmental ﬂuctuations, and cold stress. The direction of recovery and the structure of successional pathways are determined by the environment, including the disturbed landform, and not by the intensity or type of the original or newly triggered disturbance (Kom´arkov´a, 1983b, 1989), unless that disturbance is persistent and indigenous to cold regions (Kom´arkov´a, 1983a, 1989). Recovery rates and sequences are best documented where the age of the disturbance and the original landform are known. The recovery from natural disturbances, the age of which is usually not known, may involve slow changes in landform and climate as well. In cold regions, ecosystems can recover from surface disturbances of only low to medium frequency. The precise medium frequency of disturbance that keeps ecosystems permanently in successional stages by allowing them to recover only partially, seems to be rare or nonexistent. While slow recovery probably continues and complexity increases until the next surface disturbance, the similarity of most younger ecosystems to the older ones (determined by the taxonomic composition of the vegetation), after a few decades to thousands of years of recovery, falls within the limits of their variability. Because surface disturbances are more frequent, and there is probably a constant adjustment to changing climate and disturbances (Krajina, 1933; Dahl, 1956; Kom´arkov´a, 1979), coldregion ecosystems may reﬂect a slow environmental change or a trend (climatic warming, increase in UV-B radiation) more closely than temperate ecosystems. However, this appears unlikely. Despite their simpler structure, the rate of change and recovery of most coldregion ecosystems is probably slower than that of their temperate counterparts. Above certain levels of disturbance frequency, recovery does not occur and the original ecosystem shifts to one controlled by frequent or persistent disturbance, such as ground-material movement, which is tolerated only by organisms adapted to it, e.g., scree (Schr¨oter, 1926). This is also true for temperate ecosystems (Tilman, 1996), but the proportion of such ecosystems may be higher in cold regions. In the presence of frequent avalanches, the shift from vegetation dominated by trees to vegetation dominated by shrubs occurs where the average interval between avalanches is less than 15 to 20 years. At that frequency, recovery does not produce trees capable of competing with shrubs that have ﬂexible stems and

75

can reach maturity even where avalanches are frequent (Johnson, 1987). Ecosystem response to surface disturbance Recovery rates reﬂect the high diversity of environments, ecosystems, and disturbances in cold regions. They are generally slower than in temperate climates. They decrease with increasing latitude (Figs. 3.21– 3.23; Forbes, 1992), altitude, severity of climate, patch size (Kom´arkov´a, 1989; Forbes, 1992), intensity of disturbance, and importance of successional taxa and stages, and with decreasing availability of moisture (Fig. 3.24). They are considerably slower in continental than in oceanic areas, e.g., northern Sweden (Stork, 1963), the Alps (Jochimsen, 1963), the Alaska Range (Viereck, 1966), the Wrangell Mountains, Alaska (Scott, 1974). Elevated sites in mesic areas, dry hygrophobic organic matter, dry and unweathered or coarse mineral materials, e.g., such as volcanic ash (R.I.L. Smith, 1984a,b; H.G. Smith, 1985), and warm dry slopes may still be bare decades after the disturbance even when propagules are available. On Subantarctic Macquarie Island, landslides revegetate rapidly (Scott, 1983) and large grass tussocks develop in 50 years, but the windiest sites may still be bare even after this period of time (Ashton, 1965). The life-span of human and other natural disturbances is considerably changed by the natural disturbance factors and by cold stress. Persistent natural disturbances (wind erosion, movement of material under gravity, freeze–thaw effects) obliterate tracks or disperse pollutants in some habitats (sand dunes, talus slopes) within days or weeks. In ﬂat, waterlogged landscapes with ground ice, human disturbances may persist for centuries (on polygon topography) or thousands of years (on river terraces or next to thaw lakes: Everett, 1979), depending on the original landform and the type of disturbance. Most will recover before a new disturbance occurs. In cold continental and dry areas, where most surfaces are disturbed by glaciers at intervals of 20 000 years, only ecosystems on deposits 3000–5000 years old or more have complexity comparable to that of ecosystems undisturbed since the last glaciation, e.g., Colorado Rocky Mountains (Kom´arkov´a, 1979, 1993). Probably only in such areas may the dominants, zonal dominants in particular, be in a disequilibrium with the current climate, and new surface disturbances may bring about the establishment of different zonal ecosystems reﬂecting

76

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Fig. 3.21. Simple regression of percentage cover of vegetation in plots recovering from surface disturbance and their controls of natural undisturbed vegetation along the latitudinal gradient of the Antarctic Peninsula. The recovery rates and vegetation cover decrease toward the south. Some recovery occurs on human-disturbed surfaces after 30–50 years at 62ºS (Fig. 3.22) and at 64ºS (Fig. 3.23); none was observed at 68ºS. Complete development of complex plant communities has not been observed anywhere along the latitudinal gradient. Kom´arkov´a, unpublished data.

the changed climate. In polar deserts, only small surﬁcial changes occur during tens of thousands of years, and disturbances may be preserved by persistent cold. Ecosystem response to soil compression In the moist margins of the polar regions, mosses and surface organic matter do not rebound, and subsidence persists even after a passage by a single vehicle on

partially disturbed winter trails for several years (Adam and Hernandez, 1977; Felix and Raynolds, 1989a,b) – even as long as 30 years (Kom´arkov´a, 1989; Forbes, 1992, Emers et al., 1995); the dissimilarity to the original vegetation may increase during the ﬁrst ﬁve years after the disturbance (Felix et al., 1992). The effect of a single passage by a human or vehicle in summer may persist for decades – even for more than ﬁfty years (Bellamy et al., 1971; Abele et al., 1984; Wielgolaski, 1997). The injury to long-lived

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

77

Fig. 3.22. Recovery at the United Kingdom Base G (Admiralty Bay), east side of Keller Peninsula (62º05 S, 58º26 W), King George Island, South Shetland Islands, northern Antarctic. An almost complete cover of rock lichens, primarily Caloplaca, on a concrete surface that probably dates to the time when the base was constructed (1948–1950). Trampling ceased in 1961 when the base was abandoned (Hattersley-Smith, 1991).

Fig. 3.23. Recovery at the Argentinian Primavera Station, Cierva Point (64º09 S, 60º57 W), Danco Coast, Antarctic Peninsula. Part of the station was established in 1954 and a new, frequently occupied part opened in 1977 (Hattersley-Smith, 1991). Colobanthus quitensis and Deschampsia antarctica are colonizing concrete and wastes in the older part of the station. The largest cushions and tussocks are probably at least 15 years old. January, 1984.

78

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Fig. 3.24. Comparison of the rates of recovery on disturbed surfaces of different landforms in a more or less mesic cold region, such as the Arctic Coastal Plain in Alaska, the Colorado Rocky Mountains, or the northeastern edge of the Tibetan Plateau. Recovery slows down with decreasing availability of moisture and as more taxa become established. Kom´arkov´a, unpublished data.

plants in multiple-pass trails is cumulative (Ahlstrand and Racine, 1993). Vegetation and the surface organic matter are compressed much more in marshes than in uplands (Bliss and Wein, 1972), but the melting of ground ice and thermokarst effects may change uplands into marshes. These disturbances favor herbs, including exotics (Poa annua on Macquarie Island), reduce cover by bryophytes and lichens (Bell and Bliss, 1973; Hernandez, 1973; Bayﬁeld et al., 1981) and at high intensities may eliminate woody plants (Bliss and Wein, 1972; Chapin and Shaver, 1981). Recovery is rapid in wet, depressed areas with late-lying snow (Hernandez, 1973; Moskalenko, 1980, 1983), where the saturated compressed ﬁbrous organic material may regain its original structure rapidly (Abele et al., 1984). Lichen recovery may be delayed (Bliss and Wein, 1972; Forbes, 1992; Roxburgh et al., 1988), especially where moisture accumulates in compressed areas (Kershaw, 1983; Herbein and Neal, 1990; Truett

and Kertell, 1992). Where habitats do not change, lichens do recover even from substantial damage (Kershaw and Kershaw, 1987; Adamson et al., 1990; Harper and Kershaw, 1996). In partially disturbed sites, soil temperature and thaw depth increase, releasing organic matter, seeds, and nutrients which increase plant productivity and biomass (Bliss and Wein, 1972; Chapin and Shaver, 1981). On the other hand, biomass may decrease, along with moisture and nutrient availability, when organic matter is removed (Bliss and Wein, 1972; Bishop and Chapin, 1989). In the absence of freeze/thaw cycles in the arid Antarctic, tracks and other slight disturbances were visible after 33 years; sparse vegetation developed only in moist spots (Campbell and Claridge, 1982; Campbell et al., 1993). In the dry Rocky Mountains of Colorado, colonization of alpine disturbances was not complete after 10 years (Kiener, 1939) or 26 years (Griggs, 1956); sites damaged during 25 seasons of visitor use

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

may require hundreds of years to recover. However, the recovery of vegetation was almost complete in four years following a one-year trampling period (Willard and Marr, 1971; Bell and Bliss, 1973). The vegetation may recover rapidly from light damage in mesic habitats, but from serious damage only in wet habitats (Willard and Marr, 1970). In the northern Arctic, 18– 20 years after vehicle disturbance, recovery was almost complete only in the wettest meadows (Forbes, 1992). In Norway, after ﬁve years of trampling, damage was more pronounced in a lichen heath and a wet meadow than in a relatively nutrient-rich dry meadow, where the recovery was faster (Table 3.2; Wielgolaski, 1997). In Subantarctic South Georgia, no regeneration has been observed in areas devastated by the increasing populations of fur seals (Bonner, 1985).

79

Table 3.2 Study of succession over 22 years following trampling 1 Vegetation

1973

1976

1981

1986

1991

1995

Lichens

0.08

0.12

0.19

0.48

0.80

0.90

Bryophytes

0

12.50

9.60

7.68

2.82

2.25

Monocotyledons

0.18

0.48

0.14

0.20

0.57

0.03

Forbs

−

−

−

−

−

−

Woody plants

0.13

0.20

0.32

0.39

0.43

0.46

Lichens

0.05

0.23

0.84

2.30

2.11

2.70

Bryophytes

0.11

0.91

1.05

1.45

3.78

1.60

Monocotyledons

0.36

0.58

0.75

0.86

0.85

0.64

Forbs

0.46

0.65

0.79

0.71

0.73

0.65

Woody plants

0.13

0.31

0.47

0.49

0.57

0.66

Lichens

−

−

−

−

−

−

Bryophytes

0.27

1.06

1.27

1.46

1.40

1.59

Monocotyledons

0.12

0.39

0.57

0.60

0.54

0.50

Lichen heath

Dry meadow

Wet meadow

Ecosystem response to removal of surface soil After removal of vegetation and partial or total removal of soil organic matter, recovery usually begins with slow colonization by the original dominants, by the dominants of landforms similar to those newly created by the disturbance (e.g., gravel roads are colonized by species from river gravel bars), or by taxa inhabiting microdisturbances of adjacent mature, otherwise undisturbed vegetation. In dry sites, the removal of dead vegetation and the surface organic soil speeds up recovery by increasing moisture availability; in mesic sites, the propagules contained in this layer aid recovery, and removal of this material will cause a delay (Van Cleve, 1977; McGraw, 1980; Chambers et al., 1990). Vascular plants are usually the ﬁrst colonizers, sprouting from live propagules in the remaining soil, new seeds, and rhizomes or roots (S¨oyrinki, 1938, 1939; Hernandez, 1973); lichens may also colonize, reducing needle-ice disturbance and facilitating seedling establishment (Bell and Bliss, 1980; Grulke and Bliss, 1988). Burned surfaces, and bare soil compacted by trampling (Table 3.2), are usually colonized by bryophytes; bryophyte recovery may be delayed in sites where soil has been removed (Harper and Kershaw, 1996). Mycorrhizal fungi, which may be reduced or lost by disturbance, colonize rapidly (Allen et al., 1984, 1987).

Forbs

0.31

0.31

0.73

0.39

0.37

0.47

Woody plants

0.33

0.13

0.38

0.47

0.53

0.74

1 The data are ratios of cover inside and outside trails used during the International Biological Program (1969–1973) at Hardangervidda, southern Norway. Values on six line transects were averaged. Data from Wielgolaski (1997).

Vegetative reproduction predominates among colonizing taxa in the northern Arctic, where the seed supply is low, plant growth is particularly slow, and where artiﬁcial revegetation by native taxa, useful elsewhere in cold regions (Gams, 1940; Younkin, 1973; Brown et al., 1978), may not speed up recovery (Bliss and Grulke, 1988; Forbes, 1992). In the margins of cold regions, artiﬁcial revegetation by non-native taxa may delay the establishment of native taxa, as planted grasses and colonizing bryophytes reduce moisture, space, and nutrients available to native propagules, and possibly disperse inhibitory chemicals (Younkin and Martens, 1987; Densmore, 1992). Colonization is followed by a gradual increase of the number of organisms and vegetation cover (Shaver et al., 1983; R.I.L. Smith, 1985b,c, 1987, 1993, 1995). This process appears to be simpler than in temperate regions. Successional taxa are limited to natives of cold regions, where they are present, and the number of successional stages to one or two.

80

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Fig. 3.25. Ordination by detrended correspondence analysis of data for cumulative percentage cover of vascular plants at the Fish Creek Test Well 1 (70º18 36 N, 151º52 40 W) drilled in 1949 on the Arctic Coastal Plain, Alaska. Mature plant communities unaffected by human disturbance are paired in the outlined landform–vegetation categories (solid line) with communities on surfaces recovering from human disturbance. A group of undisturbed communities in the center of the ordination space (dashed line) indicates that disturbance is one of the principal factors ordering the communities on the ordination planes. Moisture, duration of snow cover, and depth of thaw are other important factors ordering them. The levels of similarity at which the composite undisturbed communities and communities that had recovered for thirty years were clustered decreased in the following order: marsh (82%), lowland (52%), snow-patch (26%), ridge (15%) and upland (14%). From Kom´arkov´a (1983a).

Marshes may form or almost recover in a few decades, even following blading1 , as long as the dominants are rapid vegetative colonizers, usually with rhizomes (Figs. 3.25 and 3.26; Kom´arkov´a, 1983a, 1989). Even marshes recover slowly, and support successional communities, when the original dominants are relatively strong competitors but poor colonizers. Zonal uplands are usually dominated by such taxa (Eriophorum vaginatum ssp. spissum in most of Arctic Alaska, Kobresia pygmaea on the Tibetan Plateau, Kobresia myosuroides in the Rocky Mountains of Colorado), which spread by seeds (Gartner et al., 1983, 1986) produced after a maturation delay (Grime et al., 1988) and form tightly closed turf or tussocks where organic matter accumulates (Fig. 3.27). However, Deschampsia antarctica, the tussock-forming zonal dominant along the Antarctic Peninsula, is an 1

efﬁcient colonizer, spreading by seeds and tussock fragments into disturbed areas. Eriophorum vaginatum ssp. spissum is a successful colonizer in the southern, relatively warm parts of its geographical area (Hopkins and Sigafoos, 1951; Chapin and Chapin, 1980). Zonal uplands, ridges dominated by dwarf shrubs and lichens, and snow patches dominated by dwarf shrubs all support successional communities of grasses and other weedy colonizers which occur in microdisturbances of undisturbed vegetation. Such successional communities develop within 2–15 years after the disturbance (Fig. 3.28; Johnson, 1969; Hernandez, 1972, 1973). Seedlings and small patches of the original dominants are usually present after 25 years and after 70 years the similarity of recovering vegetation to the controls is about 25–35% (Tibetan Plateau: Kom´arkov´a, unpublished data). After 36 years, a

“Blading” means scraping of the surface vegetation and some soil by a metal blade dragged behind a vehicle.

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

81

Fig. 3.26. A drainage ditch at the Fish Creek Test Well 1 (70º18 36 N, 151º52 40 W) drilled in 1949 on the Arctic Coastal Plain, Alaska. After the site was abandoned the same year, the ditch probably deepened through melting of ground ice and thermokarst effects. About 30 years after the disturbance, recovering marsh vegetation has 75% similarity to an undisturbed marsh. The rhizomatous dominants Carex aquatilis ssp. stans and Eriophorum angustifolium ssp. subarcticum recover directly, without a successional stage (Kom´arkov´a, 1983a, 1989).

Fig. 3.27. Mature zonal upland unaffected by human disturbance dominated by Eriophorum vaginatum ssp. spissum near the Fish Creek Test Well 1 (70º18 36 N, 151º52 40 W) site on the Arctic Coastal Plain, Alaska.

82

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A

Fig. 3.28. Recovering zonal upland dominated by rhizomatous successional grasses including Arctagrostis arundinacea and Poa arctica on a bladed surface of the Fish Creek Test Well 1 (70º18 36 N, 151º52 40 W) drilled in 1949 on the Arctic Coastal Plain, Alaska. About 30 years after the disturbance, the upland vegetation has 28% similarity to the undisturbed tussock, whose dominant, Eriophorum vaginatum ssp. spissum, is present only in small quantities. Barely disintegrated and vegetated solid waste remained on the site in 1977 (Kom´arkov´a, 1983a, 1989).

spaded plot in Alaskan tundra was only 40% vegetated with Dryas and moss (Pegau, 1970b). When the disturbance ceases in animal-controlled areas (burrows, trails, carcasses, corral, camp, and shelter sites), the original dominants eventually outcompete the nutrientsubsidized grasses and dicotyledons dependent on persistent disturbance, and the overall phytomass decreases (Rikhari et al., 1993). On the Tibetan Plateau in the Qinghai province, a failed Chinese attempt to establish grain ﬁelds covering several square kilometers in the lower-alpine region at an elevation of about 3200 m, on zonal surfaces about 150 m above oat ﬁelds in protected valley bottoms, was followed by aeolian erosion and subsidence; the recovery of vegetation cover was almost complete after 25 years, but the original dominants were present only in small quantities (Kom´arkov´a, unpublished data). Successional dicotyledons (Ajania tenuifolia) may inhibit the return of the original dominants through the release of toxic substances. In artiﬁcial grasslands, productive rhizomatous C4 grasses (Shi, 1985; Xia, 1986) are within several years outcompeted by weedy dicotyledons, and eventually by sedges (Zhang et al., 1986a,b). Extensive artiﬁcial grasslands could result in

the depletion of nutrients, loss of soil, reduction in moisture-holding capacity, and desertiﬁcation. Willow dominants of zonal surfaces in relatively warm and humid Arctic regions also occur in successional stages in areas where they do not dominate mature undisturbed vegetation. They spread from individuals or parts that survived the disturbance, but also from germinating seeds (Bliss and Wein, 1972; Hernandez, 1972, 1973). In areas where they dominate mature vegetation, woody plants may slowly recover from mechanical disturbance (Bliss and Wein, 1972; Babb and Bliss, 1974; Rickard and Brown, 1974). Soil effects in ecosystem responses The type of soil inﬂuences the process of recovery, even when plant propagules stored in the soil are discounted. On the Tibetan Plateau, burrowing by small mammals (zokors, pikas, pocket gophers) is shallow, and the nutrient-rich soil with high water-holding capacity is colonized by vascular plants 1–2 years after the disturbance. The mounds of subsurface material next to deep marmot (Marmota spp.) burrows contain unweathered ﬁne particles with small stones, and are not colonized for several years after the disturbance

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

ceases, even after the mounds are ﬂattened by erosion (Kom´arkov´a, unpublished data). While partial disturbances usually damage vegetation more than soil, recovery of vegetation may lag behind the recovery of soil properties also for other reasons. However, when soils become warmer and ground ice melts during or after the disturbance, only the recovery of the insulating vegetation cover leads to the recovery of pre-disturbance soil properties and the depth of the ground-ice table (Haag and Bliss, 1974; Gersper and Challinor, 1975; Harper and Kershaw, 1996, 1997). In Arctic Alaska, soil properties did not return to the levels of the undisturbed controls in six years in the case of vehicle tracks (Challinor and Gersper, 1975; Gersper and Challinor, 1975), or in 30 years following blading; base saturation, nutrients, and depth of thaw were higher, and organic matter, available water, silt, clay, and cation exchange capacity were lower in the disturbed areas (Kom´arkov´a, 1989). The life span of dominants may be less important for total ecosystem recovery in cold regions than in temperate areas with more rapid soil formation. The development of vegetation cover after glacier retreat and in rocky habitats during primary succession may take hundreds of years or even longer (Frenot et al., 1995; Harper and Kershaw, 1996), and soils may take thousands of years to develop. Twentysix years after its appearance, only 1–2% of the island Surtsey near Iceland was vegetated (Fridriksson, 1987, 1989). In northwestern Canada, forty-eight years after disturbance, primary-succession soils were more coarsely textured, drier, warmer, and less acidic, and accumulated much less organic matter than undisturbed soils (Harper and Kershaw, 1997); shrub birch recovered only 10% of its cover, and a substantial part of the site remained bare (Harper and Kershaw, 1996). Soil development is also slow on glacier outwash (Viereck, 1966) and moraines distant from a nutrient source (Mahaney, 1974; Messer, 1988; Matthews, 1993). On moraines in Kerguelen, in the absence of manuring by marine birds or elephant seals, the organic content of soils was low even after 200 years, whereas leaching was rapid under the relatively high precipitation (Frenot et al., 1995). Ecosystem response to grazing Damage and recovery both depend on the intensity and frequency of grazing, and are related to the growth form and other plant grazing defenses, and to the

83

manner of plant utilization by the herbivores. In an alpine meadow in the central Himalaya, tussock grass was grazed less deeply and showed better regrowth than a more deeply grazed forb and a sedge; sheep bites were signiﬁcantly deeper than those of a horse, and horse-grazed plants showed better regrowth than sheepgrazed plants (Negi et al., 1993). Grazing generally reduces lichens and shrubs, and increases herbs. The ability of deciduous shrubs and herbs to survive defoliation by low-intensity, repeated herbivory depends on nutrient reserves built up in favorable years (Archer and Tieszen, 1980; Chapin, 1980; Henry and Gunn, 1991). In graminoids, these reserves for replacement growth after grazing, die-back or mechanical damage such as abrasion by snow and ice are usually located below ground. Unlike shrubs, they also have meristems protected by long-lived, non-green, stiff stem bases, and recover rapidly once grazing or other low-intensity mechanical disturbance ceases (Babb and Bliss, 1974; Archer and Tieszen, 1980; Chapin, 1980). Clipping signiﬁcantly affects their growth only if it is repeated during the season (Mattheis et al., 1976; Archer and Tieszen, 1980; McKendrick, 1981). When grazed equally, caespitose graminoids may not recover rapidly (Jefferies et al., 1992), although in Subantarctic South Georgia Poa ﬂabellata recovers well from disturbance and severe grazing by reindeer (R.I.L. Smith, 1985a). Some rhizomatous or stoloniferous Arctic graminoids produce new leaves continuously during the grazing season and have a greater capability of regrowth than other graminoids, in which regrowth may be delayed until the next season. This is true for Carex subspathacea and Puccinellia phryganodes, which are grazed by geese along coasts, and for some other grasses (Arctophila fulva, Dupontia ﬁsheri) (Mattheis et al., 1976; Jefferies et al., 1992). Rapid recycling of goose feces and old dead leaves provides the nutrients for regrowth (Kotanen and Jefferies, 1987; Bazely and Jefferies, 1989a,b). Herbaceous dicotyledons usually have a smaller proportion of biomass below ground than graminoids, grazing decreases their ﬂowering and plant size, and it takes them longer to reach pregrazing levels after being grazed by muskoxen (Ovibos moschatus) (Mulder and Harmsen, 1995). Under intermediate frequencies and intensities of vertebrate browsing, deciduous shrubs produce leaves with progressively lower nutrient content, and growth forms that are less likely to be regrazed, whereas

84

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Fig. 3.29. Recovery on different disturbance types in a dry, nutrient-poor fellﬁeld on Niwot Ridge, Front Range of the Rocky Mountains in Colorado, in an area of relatively continental alpine climate. Unweathered subsurface material with little water-holding capacity delays recovery. Kom´arkov´a, unpublished data.

defoliation in evergreen shrubs decreases the concentration of protective secondary metabolites in leaves and increases their food value (Tuomi et al., 1984; Bryant et al., 1988, 1989). Some bryophytes are favored by grazing, but vascular plants may be favored over bryophytes (Batzli, 1975; Batzli et al., 1980; Oksanen and Ranta, 1992). Lichens may be slow to recover (Klein, 1968, 1987; Pegau, 1968). Twenty-two years following a crash in the reindeer population on an island in the Bering Sea, lichens had recovered, but only to 10% of the standing crop of lichens on an adjacent island with no history of grazing (Klein, 1987). To support migrating animals in the long term, the return intervals must be long enough, sometimes years, for vegetation to recover between visits (Andreyev, 1988). After being introduced to Alaska at the turn of the century (1892–1914), reindeer increased to 594 000

animals by 1936, then decreased to 20 000 animals. Most areas have recovered within 50 years after being overgrazed in the late 1930s, although overgrazing from the 1950s and 1960s is still evident on Bering Sea islands (Hanson, 1952; Collins, 1986). Intensive grazing or trampling by muskoxen creates microdisturbances invaded by weedy native colonizers, which may become dominant; overgrazed muskox pastures are rapidly recolonized once grazing ceases (McKendrick, 1981). On the Tibetan Plateau, overgrazing is damaging during spring and fall when the capacity for regrowth is low (Wang et al., 1986). On the relatively humid northeastern edge of the Tibetan Plateau, the original dominants have a low cover, and weedy taxa dominate both grazing exclosures (Zhou and Zhang, 1986) and 25-yr-old upland surface disturbances (Kom´arkov´a, unpublished data).

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Fig. 3.30. Recovering hydrocarbon spill on a bladed surface of the Fish Creek Test Well 1 (70º18 36 N, 151º52 40 W), drilled in 1949 on the Arctic Coastal Plain, Alaska. About 30 years after the disturbance, most of the spill is not vegetated; areas were the spill intensity is lower have 14% similarity to the control mesic surface. Solid waste has disintegrated little. The similarity of the vegetation recovering on solid waste to the corresponding vegetation unaffected by human disturbance is about 18%, but more on drier landforms (Kom´arkov´a, 1983a, 1989).

Ecosystem effects of solid waste In cold environments, disintegration of solid waste and pollutants and the release of toxic chemicals occur much more slowly than in most temperate environments. Potentially high photooxidation rates in some cold regions do not seem to contribute to higher rates of disintegration. Disintegration is faster in terrestrial environments (beaches), with relatively high temperature, humidity, salt spray, and ultraviolet radiation (denaturing plastic), than in the ocean (e.g., Ryan, 1987). Recovery rates from pollution are slower than from surface disturbances (Figs. 3.28, 3.29), but they vary with the same factors, primarily with the degree of cold and the availability of moisture. At high latitudes and altitudes, solid waste and pollutants persist and often accumulate, and the ecosystems may not recover unless the pollutant or solid waste is removed (Schindler, 1983). Surfaces of solid waste are abraded, but most remain bare of macro-organisms. Even paper in cold, relatively dry environments (south of the Antarctic Peninsula) and soft materials absorbing moisture (rope, tarpaulins, wood) in more humid cold regions (northern Alaska) persist for decades (Figs. 3.30 and 3.31) (Kom´arkov´a,

1983a, 1989). Colonization is delayed on unweathered solid-waste materials with low water-holding capacity (concrete, asphalt), especially on unweathered mining waste which may contain acids and toxic materials (Fig. 3.29), and on surfaces affected by chemicals (oil, paint, rust) and away from a substrate which can conduct moisture and supply abrading particles. Revegetation of solid waste is more rapid when it is covered by windblown mineral and organic particles that absorb moisture. The communities colonizing solid waste are similar to those on the surrounding landform. Solid waste of natural exotic materials (wood, gravel, concrete) may support adventive cold-region taxa in areas without wood and rocks. In its vicinity, solid waste creates disturbances such as shading, reduced wind, reﬂected heat, and prolonged duration of snow cover. Bryophyte-dominated communities in such environments are outcompeted by grasses or sedges soon after shading is removed. Ecosystem effects of pollution Many slow-growing organisms in the cold regions accumulate pollutants (Focardi et al., 1992), and some of the accumulations may reach harmful levels

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Fig. 3.31. Solid waste, surface disturbance, and diesel fuel spills near Chaplinski hot springs on the Ul’khum River, Chukotski Peninsula, Russian Arctic. Plants colonizing surface disturbances, both here and elsewhere in cold regions, include mostly rhizomatous apophyte grasses (Arctagrostis arundinacea, Leymus interior), adventive grasses (Elytrigia repens) and ruderal plants (Artemisia tilesii). Photograph by A.E. Katenin.

Fig. 3.32. Unweathered subsurface mining waste on Niwot Ridge, Front Range of the Rocky Mountains in Colorado is not completely vegetated even after perhaps 70 to 100 years, despite being located in an area with late-lying snow and relatively high vegetation cover. The dominants of the recovering vegetation and the vegetation unaffected by human disturbance are the same (Trifolium parryi ssp. parryi, Acomastylis rossii ssp. turbinata).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

(Bargagli et al., 1993). Yet the dominance of cold and environmental ﬂuctuations combine to limit the effect of low-level pollutants. Arctic marine invertebrates (e.g., oyster larvae, crustaceans, and annelid worms) are tolerant of high concentrations of zinc and lead, high salinity, and other extremes (Jones, 1975; Chapman and Brinkhurst, 1984; Lewis and Horning, 1991). Because the solubility of metal salts and the rate of movement of water and solutes across cell membranes decrease with temperature, metal toxicity may also be reduced (MacInnes and Calabrese, 1978; Chapman and McPherson, 1993). Cold and long winters, when soils are frozen and the ocean is covered by ice, and summer cold temperatures, delay the breakdown of pollutants and their entry into food chains, so that pollutants may be especially persistent in cold regions. Slow rates of production, decomposition, and chemical reactions account for the low self-puriﬁcation and buffer capacity of cold ground and water. The rates of chemical reactions and pollutant breakdown decrease with increasing latitude and altitude, and in dry areas, where growth of organisms and turnover rates are slower (Horowitz et al., 1978; Hutchison-Benson et al., 1985; Adamson and Seppelt, 1990). Although the levels of contamination are less than at lower latitudes, Cs137 deposited over the last decades does not turn over readily in Arctic ecosystems, and they retain elevated levels for many years without further additions – longer than temperate ecosystems. The persistence of Cs137 in ecosystems increases with increasing latitude (Svoboda and Taylor, 1979; Hutchison-Benson et al., 1985; Taylor et al., 1985). Ecosystem effects of ultraviolet-B radiation The present impact of enhanced UV-B radiation is relatively small, as compared to the photoinhibition resulting from UV-B and UV-A radiation under normal conditions of stratospheric ozone. Up to thresholds for UV irradiance that occur at certain water depths, some Antarctic phytoplankton can adequately respond by synthesis of UV screening compounds (mycosporine-like amino acid pigments) and by repair mechanisms. It does not appear that such thresholds have been exceeded. The degree of increased protective pigmentation may be determined by the balance between increasing protection and decreasing effectiveness of photosynthesis (El-Sayed

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et al., 1990; Karentz et al., 1991; Holm-Hansen et al., 1993). When exposed to supplementary UV-B radiation in the ﬁeld, signiﬁcant accumulation of UV-absorbing leaf pigments occurred in vascular plants from higher latitudes, even if it was insufﬁcient to protect them completely; most alpine plants were apparently protected without additional pigments (Barnes et al., 1987). Synthesis of photoprotective ﬂavonoids in the moss Bryum argenteum is positively correlated with the levels of UV-B irradiation and negatively correlated with levels of ozone in the Antarctic stratosphere (Markham et al., 1990). Protective pigments did not change in cyanobacterial mats in the Antarctic under four UV regimes in the ﬁeld, but pigmented cells might have migrated to the surface of the mat (Quesada et al., 1998). Ecosystem responses to increase in carbon dioxide and climatic change In cold regions, the effects of higher atmospheric concentration of carbon dioxide alone will be limited. Levels of carbon dioxide in dense moss and lichen carpets, turfs, in moist areas where decomposition is high, and under spring snow may be many times higher than in the surrounding air, even with the projected increase in carbon dioxide (Peterson and Billings, 1975; Billings et al., 1984; Oberbauer et al., 1986a). Long-term experimental exposure to elevated atmospheric concentration of carbon dioxide increased the root:shoot ratio or tiller production in some taxa (Oberbauer et al., 1986b; Tissue and Oechel, 1987), but it did not have a fertilizing effect except under higher temperatures (Oechel et al., 1994). Small alpine plants ﬁx more carbon dioxide per unit area than lowland plants, possibly because of low temperatures and a greater length in the diffusion path in their thicker leaves; even under moderate climatic warming and enhanced atmospheric nitrogen deposition, they do not show positive phytomass responses to the enrichment of atmospheric carbon dioxide (K¨orner and Larcher, 1988; K¨orner, 1994; Sch¨appi and K¨orner, 1996). Most Arctic plants may have inherent physiological limitations to their ability to use additional carbon dioxide (Oechel and Billings, 1992). Nardus stricta showed a reduced photosynthetic and ﬂowering capacity near cold springs in Iceland giving off carbon dioxide (Oechel and Vourlitis, 1996). The current initial response to climatic warming is

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slight; but, as in the past, persistent warming will cause changes in the extent of sea ice and in atmospheric and oceanic circulation, leading to ecosystem shifts upward and poleward. If the amplitude of change rapidly exceeds the current marginal temperatures of mountain and boreal forests, then trees will quickly outcompete the low-growing cold-region vegetation in areas with enough moisture. Ecosystems and organisms that do not have anywhere to retreat to will disappear (Tast, 1991; Grabherr et al., 1994), and coldadapted taxa that cannot acclimatize will be eliminated (Marchant, 1992). In some areas, the migration could be haphazard, as some slow-growing plants may not be able to track the rate of environmental change (Grabherr et al., 1978; K¨orner, 1994). Some treeless areas will become isolated, and trees will not be able to colonize rock outcrops and marshes. However, most ecosystems will probably migrate as units, and the present dominants will maintain the role determined by their competitiveness in new, equivalent environments. It seems unlikely that migrating ecosystems will be inﬂuenced by non-inherent herbivores, ﬁres, or insect outbreaks. Changes in taxonomic composition will be reﬂected in changed demand for nutrients and water, decomposition, herbivory, and predation (Smith and Steenkamp, 1990; Jefferies et al., 1992; Seppelt, 1992). Infectious diseases and microorganisms which are rare in cold regions today will spread from lower latitudes and elevations, and people will have to change their practices – for instance, in agriculture and food preservation. Crop agriculture will be possible in more habitats than today. Some ecosystem changes are taking place already, although the replacement of dominants, zonal dominants in particular, is rare if it occurs at all. In the recent past, forest advanced only in a few places, and it is probably not advancing today. In Scandinavia, it expanded mainly through recruitment between 1950 and 1964 (Kullman, 1979, 1981, 1988, 1993); in the Rocky Mountains in Colorado, tree invasion of forest openings at the tree-line probably dates from the same, warmer and wetter period (Hessl and Baker, 1997). In the European Alps, taxonomic richness is increasing on summits, and nival plants are migrating upward at rates of 0–4 m per decade – rates far lower than the hypothetical movement of temperature isolines (Hofer, 1992; Gottfried et al., 1994; Grabherr et al., 1994, 1995). On Hopen Island in the Svalbard Archipelago, during the past 110 years, the proportion of vascular taxa with seed dispersal has increased at

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the expense of taxa with only vegetative reproduction (Skye, 1989). The number of vascular plant taxa is positively correlated with summer temperatures both there and in Arctic Canada (Rannie, 1986). On the Antarctic Peninsula and the Subantarctic islands, glacial retreat has exposed new areas for colonization, but only dominants or colonizers from adjacent areas are moving in (R.I.L. Smith, 1982; Hayward, 1983; Allison and Keage, 1986). However, the reported spread of Colobanthus quitensis to Gamage Point in Anvers Island, Antarctic Peninsula (Grobe et al., 1997) apparently concerns an extensive experimental garden where Colobanthus seeds were sown out and to which mature Colobanthus plants with seeds were transplanted in 1983–1984 (Kom´arkov´a, unpublished data). Some deglaciated areas cannot be colonized because snow and ice edges are moving away from ecosystems previously supplied by melt-water. Along the Antarctic Peninsula, some mossbanks are not recovering from decades-old disturbances, because their supply of moisture is drying out as a result of recession of glaciers and snow-patches (Kom´arkov´a, 1983b) caused by climatic warming in that area (Morrison, 1990; Stark, 1994). On the drying Tibetan Plateau, recovery may be slowing down, and may not occur, especially in arid central and western Tibet. Some invading taxa seem to be slowly outcompeting minor resident taxa, especially those that have been weakened by climatic warming, drought, or pollution; others are colonizing microdisturbances created by animals or by freeze–thaw processes between dominant plants. Lichens are overgrowing bryophyte dominants, e.g., on the Antarctic Peninsula (Kom´arkov´a, 1983b); where the moisture supply has decreased, fungal infections of the bryophytes are spreading (Longton, 1973). Shifts in atmospheric and oceanic circulation may be responsible for a steeply increasing proportion of exotics in the pollen rain on Marion Island, and easier invasion by exotic taxa (Scott and van Zinderen Bakker, 1985; Smith and Steenkamp, 1990). Other recent invasions include a terrestrial isopod, possibly introduced by human transport, on Macquarie Island (van Klinken and Green, 1992), additional vascular plants on Heard Island (Scott, 1989, 1990), algae (Broady and Smith, 1994), and bipolar lichen taxa (Smith and Øvstedal, 1994; Øvstedal and Gremmen, 1995). The current success of cosmopolitan weeds at cold-region margins, e.g., Poa annua on King George Island, South Shetland Islands (Olech, 1996), and of vigorous invaders on Marion Island in the

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Subantarctic – Agrostis stolonifera, Sagina apetala, herbivorous aphids and the diamondback moth Plutella xylostella (Crafford and Chown, 1987, 1990; Bergstrom and Smith, 1990; Chown and Smith, 1993) may be attributed to climatic warming (V.R. Smith and Steenkamp, 1990). There is little growth response to increasing temperatures among current cold-region plants, as a result of genetic constraints, low availability of nutrients (primarily nitrogen), or drought (Oechel and Billings, 1992; Havstr¨om et al., 1993; Wookey et al., 1993). The temperature optima of many alpine and arctic plants for uptake of carbon dioxide and for respiratory losses are similar to those of lowland plants (Tieszen, 1973; Tieszen et al., 1980; K¨orner, 1982). The acclimation to higher temperatures of dark respiration by leaves and of the photoperiodic control of development are different in different taxa (K¨orner and Diemer, 1987; K¨orner, 1994; Larigauderie and K¨orner, 1995), and the importance of photoperiod increases with increasing latitude and altitude (Heide, 1985, 1992; Prock and K¨orner, 1996). Temperature-opportunistic taxa may beneﬁt from warming, especially if more water and nutrients become available, whereas taxa strongly controlled by photoperiod will not (Prock and K¨orner, 1996). Elevated night temperatures may affect tissue development (K¨orner and Pelaez MenendezRiedl, 1989; K¨orner, 1992). Until recently, some wet cold-region margins have been accumulating soil carbon, mostly in the form of dead organic matter slowly being incorporated in insulated, rising ground ice (Coyne and Kelley, 1978; Chapin et al., 1980b; Billings et al., 1982). Probably because of increasing temperatures and consequently decreasing site moisture, tussock tundra has become a source of carbon dioxide (Grulke et al., 1990; Oechel and Billings, 1992; Oechel et al., 1993, 1994), while wet marshes continue to be a source of methane (Sebacher et al., 1986; Whalen and Reeburgh, 1990a,b). Migration of carbon-accumulating ecosystems to higher latitudes and altitudes where little carbon is accumulated today, and the movement of forests into the current tundra areas, would signiﬁcantly increase the removal of carbon dioxide from the atmosphere (Pastor and Post, 1988; Melillo et al., 1990; Oechel and Billings, 1992). In areas which remain treeless, warming temperatures and increases in precipitation, thaw depth, and the activity of anaerobic microorganisms could lead to increased emission of methane and

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renewed accumulation of atmospheric carbon dioxide in dead organic matter; if warming were moderate, the upper limit of permafrost could rise (Billings et al., 1982, 1983, 1984; Oberbauer et al., 1986a). Alternatively, if peat in waterlogged marshes dries out, their methane emissions would decrease, but positive feedback loops for carbon dioxide could decrease the thickness of the organic layer (M. Smith, 1990; Waelbroeck, 1993), carbon currently stored in dead soil organic matter could be depleted, and climatic warming could intensify (Billings et al., 1982; Whalen and Reeburgh, 1988, 1990a,b). Reduction of sea ice associated with a climatic warming would diminish the sea ice ecosystems and the nutrients in the Antarctic Bottom Water and their transport north (Marchant, 1992). The southern seals, penguins, and other sea birds could beneﬁt from closer feeding and breeding areas, but the foraging areas of northern animals may already be more remote (Croxall et al., 1988; Smith and Steenkamp, 1990; Budd, 1991). Some feeding areas and population distributions of marine birds and mammals may also be affected by the El Ni˜no Southern Oscillation (Vergani and Stanganelli, 1990). Areas abandoned by penguin colonies during cool intervals (Stonehouse, 1970; Baroni and Orombelli, 1994; Emslie, 1995) could be reoccupied. Early melting and breakup of sea ice may result in the loss of the entire year’s brood of emperor penguins (Aptenodytes forsteri) through the break-up of colonies or drowning of chicks (Spindler 1994; Spindler and Dieckmann, 1994). Other penguins may suffer from too much heat, lack of snow patches on which they cool down after ﬁshing, or greater snowfall (Kaiser, 1997). Ecosystem responses to hydrocarbon spills The numbers of cold-environment bacteria and other organisms (fungi, yeasts) utilizing hydrocarbons are similar to more temperate areas, but their numbers usually rise more slowly after a spill than in the temperate zone, especially in the absence of sufﬁcient nutrients, oxygen, or moisture. The number of microbes usually decreases, because only hydrocarbon metabolizers can tolerate some of the toxic compounds present (Cundell and Traxler, 1974; Horowitz and Atlas, 1977; Sexstone and Atlas, 1977a,b). Bacterial communities may appear unchanged after 14 days of contamination, following an initial enhanced growth (Delille and Siron, 1993). Planktonic primary production slowly recovers but

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the taxonomic composition of the plankton changes, with an increase in relatively oil-resistant algae (Miller et al., 1978; Federle et al., 1979; Atlas, 1985). Heterotrophic activity may recover rapidly, especially in the water column (Jordan et al., 1978; Grifﬁths et al., 1981a,b). Copepods and other aquatic organisms may be tolerant of high levels of dispersed crude oils, but suffer internal and external contamination for weeks. Hydrocarbon concentrations in ﬁlter-feeders are usually lower than in deposit-feeders, which continue to take in oil from the sediment even a year after a spill (Wells and Percy, 1985). Oil-coated seals were clean after intervals ranging from one day to one month. Seals were efﬁcient in excreting hydrocarbons from their bodies, and showed little pathological damage whether exposure was by immersion or by oil ingestion (Geraci and Smith, 1976; Engelhardt, 1985). On land, hydrocarbon spills are usually more damaging in dry sites than in wet ones, where they may disperse quickly; but they may persist for decades, as hydrocarbon-degrading microorganisms may be limited by available nitrogen and phosphorus (Atlas, 1977). Hydrocarbon spills remained bare even after 30 years in Arctic Alaska (Kom´arkov´a, 1983a, 1989), and under a 28-yr-old spill of diesel fuel there was still a signiﬁcant depression of permafrost and a toxic component in the soil. Reﬁned petroleum products are considered more toxic to plants than crude oil, and their spills are colonized more slowly (Plice, 1948; Cundell and Traxler, 1974; Linkins and Antibus, 1978). The recovering communities are similar to those on the surrounding landform. Recovery of oil spills may be accelerated by bioremediation in fertilized soils (McKendrick and Mitchell, 1978a,b; McKendrick, 1987; Kerry, 1993). Ecosystem effects of ﬁre Because cold-region vegetation is seldom completely destroyed by ﬁre and high root phytomass may increase resilience (Payton and Mark, 1979), recovery following ﬁre may be relatively rapid (Hall et al., 1980; Fetcher et al., 1984). In the wet Arctic, Carex spp. and Eriophorum vaginatum tussocks often survive, and their recovery may take only two years (Wein and Bliss, 1973a). Rapidly growing herbs and grasses which occur as minor components in the original vegetation may increase their cover when the dominants are destroyed. Bryophytes colonizing burned surfaces may persist for decades (Johnson and Viereck, 1983;

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Tvorogov, 1988), or reach maximum cover 5–6 years after ﬁre and then decline as the vascular overstory develops (Racine et al., 1987). Lichen recovery is much delayed (Wein and Bliss, 1973a; Wein, 1975; Hall et al., 1978). The proportions of the original taxa did not recover in Eriophorum vaginatum tussock tundra 13 years after ﬁre (Fetcher et al., 1984). In New Zealand, recovery of burned Chionochloa tussocks may take 14 to 15 years (Payton and Mark, 1979). Empetrum evergreen dwarf-shrub and lichen vegetation on Amchitka Island in the Aleutians may take 20 years to recover (Shacklette et al., 1969). After a ﬁre in South American p´aramo, dicotyledons with stem rosettes increased growth, but had a higher mortality rate (Verweij and Kok, 1992). Meristems survived burning, but trampling and consumption by cattle fragmented burned grass tussocks. Under increased frequency and intensity of burning and grazing, tussocks are replaced by shorter, more grazing-tolerant grasses and herbs, and the productivity, turnover, and belowground biomass increase. Further disturbance decreases biomass by up to 66%, and causes poor recovery of stem rosettes and large patches of bare ground (Cleef, 1981; Schmidt and Verweij, 1992; Verweij and Budde, 1992). Increased growth, ﬂowering, and production after ﬁre have been attributed to higher surface and soil temperatures, a surplus of moisture, increased nutrient availability as a result of increased mineralization and microbial activity in the warmer soils (Rowley, 1970; Wein and Bliss, 1973a; Weber, 1975), litter removal, and mobilization of reserves (Williams and Meurk, 1977). The nutrients in ash are generally not considered signiﬁcant (O’Connor and Powell, 1963; Mark, 1965; Wein and Bliss, 1973a). The recovery of insulating vegetation leads to a decrease in the depth of thaw and ground uplift (Mackay, 1995). Fires may rejuvenate ecosystems in cold regions (Wein and Bliss, 1973a), but there appear to be no ecosystems dependent on ﬁre. Changes in animal populations International and national agreements, and moratoria on commercial exploitation of cold-region animals, have helped to arrest or reverse the past steep declines of many populations of Greenland fur seal (Callorhinus ursinus), walrus, bowhead whales, and gray whales (Eschrictius robustus) (National Foreign Assessment Center, 1978). In Arctic Canada, populations of muskoxen (Ovibos moschatus), reduced by hunting for

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

meat, hides, and furs to a few hundred individuals by 1917, have recovered to over 1.3×104 animals. However, commercial exploitation continues, and some populations are delayed in returning to previous levels. Populations ﬂuctuating with climate (fur seals, penguins) may be more resilient to harvesting than whales. Populations of cold-region animals are dependent on ﬁnite food resources, so that declines in one species may lead to increases in others. For example, harvesting of whales apparently favored penguins and fur seals. The Antarctic biomass of baleen whales (Mysticeta), excluding the minke whale (Balaenoptera acutorostrata), decreased from 22×106 metric tons in the early 1920s to 2×106 metric tons in the mid-1960s. They previously consumed 100–150×106 metric tons of krill; this biomass then became available to other foragers (Mackintosh, 1970). Macaroni penguins (Eudyptes chrysolophus) on the Kerguelen Islands, and chinstrap penguins (Pygoscelis antarctica) and Ad´elie penguins (Pygoscelis adeliae) in Ad´elie Land and the Ross Sea region may be increasing as a result of a reduction in the numbers of whales (Sladen, 1964; Conroy, 1975; Croxall et al., 1984, 1988). Some populations of Ad´elie penguins have been increasing, even in human-disturbed rookeries, since 1982 – more than 15 years after the collapse of baleen whale stocks. This could be related to greater availability of winter food, caused by the reduction in sea ice which reﬂects the recent warming trend (Adamson et al., 1988b) in the Ross Sea area (Taylor et al., 1990), or to a combination of factors related to winter or spring sea ice, whale decline, krill availability, and breeding success (Wilson, 1990). This may also be true for chinstrap penguins (Fraser et al., 1992; Emslie, 1995), king penguins on Macquarie Island (Rounsevell and Copson, 1982), and Antarctic fur seals (Laws, 1977; Payne, 1977). Under the current climatic warming, the amplitude of ﬂuctuations in populations of marine birds and mammals could diminish, and some populations could reach higher numbers in one part of their area than the carrying capacity of another part, leading to disturbances. Since the 1930s, from a few tens of animals (Payne, 1977), fur seal populations in South Georgia and the South Shetland Islands have increased to 9×105 animals (Bonner, 1985; Bengtson et al., 1990). They import more nutrients than the Antarctic terrestrial ecosystems can absorb, destroy them by trampling and lethal concentrations of nutrients

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(Bonner, 1985; Smith, 1988), and cause eutrophication of lakes (Ellis-Evans, 1990; Hawes, 1990). Fur seals have eradicated and damaged large expanses of Poa ﬂabellata tussock grassland on South Georgia (Bonner, 1985) and Deschampsia antarctica meadows on Bird Island (Payne, 1977). Destruction of tussocks by lactating females feeding their pups deprives birds such as burrowing petrels of breeding habitat and exposes them to predation by Antarctic skuas (Catharacta antarctica) (Bonner, 1985). Young wandering albatrosses (Diomedea exulans) are disturbed by seals; their reproductive success has become lower, and the population is declining (Croxall et al., 1984, 1990).

CONCLUSIONS

Natural disturbances The gradient of surface heat, gradually decreasing beyond the marginal conditions for life, characterizes all cold regions. High altitudes and latitudes covered by ice, snow, and barren rock are dominated by the deep cold; temperatures above 0ºC are rare or nonexistent, and there are no resident higher organisms. At lower altitudes and toward the equator, and on the margins of the cold regions, a remarkable spatial and temporal diversity of natural stresses and disturbances is produced by ﬂuctuations along steep environmental gradients, probably both in the oceans (light, nutrients, salinity) and on land (temperature, water, nutrients, duration of snow cover, wind, ice, gravity, animals). Natural stresses and disturbances are more frequent, variable, and intense than in most other biomes, and the steep environmental gradients control both their outcome and that of human-related disturbances to a greater degree. Ecosystem characteristics that contribute to the continuation of the original disturbance, or enable it to trigger new, different disturbances, are also more common in cold regions than in most other biomes. Near-surface ground ice, wind and moving water, freeze–thaw activity, unconsolidated surface deposits, and steep slopes all lower both ecosystem resistance and resilience, and disturbance-triggered processes (melting of ground ice, ground collapse, increased movement of material, changes in moisture and snow regimes) may eventually cause the replacement of the original landforms and/or ecosystems by new ones. Organic matter and pollutants accumulate in certain

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sinks, but in most cold regions there is greater export and import of materials by more powerful disturbance agents acting over longer distances and greater altitudes than in most other biomes. There is probably also more interconnectedness between cold regions, especially the small alpine ones, and the respective surrounding areas, than between other biomes and their neighbors. Most cold-region ecosystems have fewer organisms and lower complexity and productivity than their temperate counterparts. Interruption of ongoing evolutionary processes which would lead to adaptations to coldregion stresses and disturbances may dominate over an increase in evolutionary rates related to frequency of disturbances, and thus contribute to the organismic paucity of cold regions, in addition to their lack of resources. Most cold-region adaptations serve multiple purposes, e.g., a marine animal with a hard outer calciﬁed shell has protection from abrasion, from being eaten, and from ultraviolet radiation (Karentz, 1994). Many cold-region organisms also have mechanisms by which to avoid severe disturbance and stress. Geographical limits of taxa may be determined by their tolerance of summer environmental ﬂuctuations. If the harmful ﬂuctuations become wider or more frequent, or move toward one or the other extreme, the geographical area of some organisms may diminish. Diversity of ecosystems is low in areas where deep cold predominates. In areas with ground ice, where surface disturbances recur at intervals between one thousand and several thousand years, ecosystems may be more diverse than those disturbed less often. Disturbances probably control the ecosystems that they disturb if they occur at least once every 20 years, or more often, depending on the characteristics of the disturbance and ecosystem. Only organisms capable of rapid growth, dispersal, and reproduction dominate in frequently disturbed areas. During recovery, successional stages are absent in stressed (the coldest) or frequently disturbed areas. They are usually present in zonal habitats at the margins of the cold regions. Most cold region ecosystems have relatively little resistance and resilience towards disturbance. More resistant ecosystems are usually less resilient, and more resilient ecosystems are less resistant. Low resilience stems from slow growth, failure of reproduction and establishment, and interruptions of the recovery process by new disturbances. Recovery rates are probably slower than in most other biomes, and some recovering ecosystems may not develop the maximum possible

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diversity of microtopography and complexity by the time of the next surface disturbance. This may happen when the interval between disturbances is between several decades and few thousand years long. Recovery may lead to ecosystems which differ from the original because of an intervening climate change; this probably occurs commonly in relatively stable cold regions where ecosystems are destroyed at intervals longer than several thousand years, and recovery may take even longer. Human disturbances Small native human populations live and practice agriculture in the cold-region margins. Their impact is limited, even if most have extensive contacts with lower latitudes and altitudes, the source of people and materials causing greatest disturbance. This inﬂux and the disturbances are increasing, with growing human population, in most cold regions. In remote cold-region areas, human disturbances of relatively low intensity and extent occur only infrequently. In more accessible parts, they recur seasonally (tourism, research) or are persistent (construction, extraction of minerals and fossil fuel), but still affect only a small proportion of the total area. In contrast to lower latitudes and altitudes, human-controlled ecosystems are few, and limited in extent. Ecosystems controlled by domestic grazers are widespread in some marginal alpine regions. Suggestions as proposed by Laws (1984) and Gudmundsson (1986), that the secondary production of cold regions used by humans can increase, must probably be discounted, because the estimates of productivity and stocking rates used (Arnalds and Rittenhouse, 1986; Zhou et al., 1986) do not take into account many factors. The carrying capacities for people and livestock [2–3 Tibetan sheep per hectare (Pi and Zhao, 1986); 80–100 vicu˜nas per square kilometer (Rabinovich et al., 1985); see also estimates by Virtala (1992) for reindeer] may be diminishing due to overgrazing and other disturbances, especially in areas where climatic warming and drying trends coincide. Climatic cooling and grazing overuse probably led to a steady decrease of the primary productivity and the vegetated area in Iceland over the last 1100 years, and to the demise of medieval Norse settlements in Greenland (Fridriksson, 1972, 1986; Thorsteinsson, 1986; Whitaker, 1990).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

Stresses and disturbances related to human activity increase the naturally high levels of stress and disturbance, especially for cold-region organisms at their geographical and tolerance limits. The effects are more obvious than in most ecosystems of warmer climates, partly because organisms and ecosystems have little resilience and resistance to disturbance. The importance of where and when the disturbance occurs is perhaps greater in cold regions than in other biomes; the growing season is short, and the productive areas limited in size. Human disturbances which free space are tolerated in habitats with similar natural disturbances (trampling by humans in animaltrampled habitats, blading on ice-scoured shores), unless the new disturbances are more intensive, longlasting, or frequent (overgrazing). Some destructive human disturbances are equivalent to natural ones (rock blasting and rockfall). Some types of pollution initially at a low level may affect polar regions more than elsewhere because of stable polar air, the earth’s rotation, the magnetic poles, or the large masses of ice located there. They could cause a gradual die-off of taxa, and facilitate invasions by taxa from the surrounding areas. Simple cold-region ecosystems could be overwhelmed by smaller amounts of pollutants than might be tolerated in temperate ecosystems, but they are protected by slow chemical reactions, slow turnover, slow accumulation in sinks, long distances for pollutant dispersal, re-export of the pollutants by disturbance agents (wind, water, gravity, animals), and by organismic tolerance of environmental ﬂuctuations. Existing adaptations may not enable the organisms to respond adequately to an increase in UV-B radiation that could destabilize polar food webs by decreasing primary production, but subtle ecosystem responses are more likely (Vincent and Roy, 1993). Rapidly reproducing organisms in exposed habitats may develop the minimum repair capacity needed to survive (Calkins and Thordardottir, 1980; Calkins, 1982; Frederick and Lubin, 1994). The impact of infrequent high-intensity point disturbances over limited areas (e.g., mining in parts of the Russian Arctic) are probably irreversible. Possibly greater irreversible damage has been done by partial non-point disturbances scattered over large areas (excessive harvesting of whales, overgrazing of high-altitude plateaus). Climatic warming or cooling related to the release of “greenhouse gases”, increase in ultraviolet radiation related to ozone layer destruction, and acid rain could also cause similar damage, but

93

their demonstrated effects are limited. Cold-region organisms (fur seals, some vascular plants, lichens) and some invaders seem to be advancing slowly toward the centers of some cold regions. Rapid ecosystem response to the current climatic warming or cooling can probably be expected only where precipitation, meltwater, or ground ice conditions change rapidly. Indirectly through responses in the cold regions, organisms and ecosystems in the areas from which the stresses originated could also be affected. The most likely long-term effects of a climatic warming will be that the organisms now resident in the cold regions, and adapted to cold, will be outcompeted by more productive taxa invading from the adjacent temperate areas. There are few if any obligatory psychrophiles; but, if the rate of warming is rapid, the organisms best adapted to cold, with adaptations that cannot be changed rapidly (insulation, hibernation habits, and some migration patterns), may be the ﬁrst ones to become extinct. The ﬁrst ecosystems to disappear could be small alpine areas surrounded by advancing forests, and the Antarctic terrestrial ecosystems inhabiting the narrow strip of land between the oceans and glaciers; this land could be drowned before the thick glaciers retreat. Marine organisms in the Antarctic, where the Antarctic continent is in the cold center, and land organisms in the Arctic, where the Arctic Ocean occupies the same position, have few opportunities for long-term retreat. Subtle experimental manipulations of environmental factors may reveal the initial responses of organisms and ecosystems to long-term, gradual environmental change. However, long or intense experiments will only lead to changes in taxonomic composition and eventually will produce ecosystems similar to natural ecosystems controlled by the same environmental factors that are being manipulated. Such ecosystems can usually be found in the vicinity of the manipulated plots – ecosystems in neighboring warmer areas (climatic warming), on acid substrates (acid rain), on bird posts or in small mammal burrows (fertilization), or in snow-patches (prolonged duration of snow cover). Cold-region organisms and ecosystems should be protected from further human disturbances that may make them more vulnerable to slow environmental changes. Alpine regions, including the high plateaus (Tibetan Plateau), should be protected as nature sanctuaries or National Parks. Protection in the Southern

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Ocean Whale Sanctuary should be extended to all animals, because krill harvesting and ﬁshing put both marine and terrestrial food webs in the Antarctic out of natural balance. A marine World Park in the north, both terrestrial and marine World Parks in the south (Herber, 1991, 1992), and national Arctic parks including undisturbed lands, should be established. Cold-region ecosystems should also be protected from excessive increases in the populations of indigenous animals, e.g., fur seals (Bonner, 1985), and from exotic invader organisms.

ACKNOWLEDGMENTS

The authors thank Drs. D.W. Goodall and L.R. Walker for their kind improvements of the manuscript. V.K. would like to thank all researchers who have sent reprints and bibliographies, and Drs. M.P. Hainard and M.H. Cl´emen¸con for the use of the facilities at the University of Lausanne. Her work was partly aided by U.S. National Science Foundation grants DPP-8201047, DPP-8611827, DPP-7825748, and OPP-7512949, and by a grant from the Committee on Scholarly Communication with the People’s Republic of China, and the U.S. National Academy of Sciences.

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Worland, M.R. and Block, W., 1986. Survival and water loss in some Antarctic arthropods. J. Insect Physiol., 32: 579–584. Xia, W.P., 1986. Brief introduction of the International Symposium of Alpine Meadow Ecosystem. In: The Proceedings of the International Symposium of Alpine Meadow Ecosystem. Northwest Plateau Institute of Biology, Academia Sinica, Xining, pp. 1–9. Young, E.C., 1990. Long-term stability and human impact in Antarctic skuas and Ad´elie penguins. In: R.K. Kerry and G. Hempel (Editors), Antarctic Ecosystems. Ecological Change and Conservation. Proceedings of the 5th SCAR Symposium on Antarctic Biology, Hobart, Australia, 1988. Springer, Berlin, pp. 231–236. Young, S.B. and Kl¨ay, J-R., 1971. Bryophytes in the 1969 crater of Deception Island, Antarctica: an apparent case of rapid longdistance dispersal. Ohio J. Sci., 71: 358–362. Younkin, W.E., 1973. Autecology studies of native species potentially useful for revegetation, Tuktoyaktuk region, N.W.T. In: L.C. Bliss and R.W. Wein (Editors), Botanical Studies of Natural and ManModiﬁed Habitats in the Eastern Mackenzie Delta Region and the Arctic Islands. ALUR Report 72–73–14. Department of Indian Affairs and Northern Development, Ottawa, Canada, pp. 45–76. Younkin, W.E. and Martens, H.E., 1987. Long-term success of seeded species and their inﬂuence on native species invasion at abandoned rig site A-01, Caribou Hills, N.W.T., Canada. Arct. Alp. Res., 19(4): 566–571. Yu, X.G. and Sun, S.Z., 1981. The upper limit of agriculture in Xizang and its factor analysis. In: D.S. Liu (Editor), Geological and Ecological Studies of Qinghai-Xizang Plateau. Vol. 2. Environment and Ecology of Qinghai-Xizang Plateau. Proceedings of the Symposium on Qinghai-Xizang (Tibet) Plateau, Beijing, China. Gordon and Breach, New York, pp. 2033–2044. Zakharov, V.G., 1988. Fluctuations in ice shelves and outlet glaciers in Antarctica. Polar Geogr. Geol., 12: 297–311. Zhang, B.C., Zhen, R.D., Gu, L.H. and Bai, X.F., 1986a. Study on the allelopathy in the natural degenerated process of manmade grassland on alpine meadow. I. The correlation of the degenerated speed of man-made Elymus nutans grassland with the invasion of Ajania tenuifolia. In: The Proceedings of the International Symposium of Alpine Meadow Ecosystem. Northwest Plateau Institute of Biology, Academia Sinica, Xining, p. 56. Zhang, B.C., Bai, X.F., Gu, L.H. and Zhen, R.D., 1986b. Study on the allelopathy in the natural degenerated process of manmade grassland on alpine meadow. II. The investigation of intraspeciﬁc and interspeciﬁc alleloinhibitory phenomena during the germinating course of Elymus nutans and Ajania tenuifolia seeds. In: The Proceedings of the International Symposium of Alpine Meadow Ecosystem. Northwest Plateau Institute of Biology, Academia Sinica, Xining, p. 57. Zhou, X.M. and Zhang, S.L., 1986. Change state of community structure and biomass in fenced area of Kobresia humilis meadow. Acta biologica plateau sinica 1986 (5): 1–6. (in Chinese). Zhou, X.M., Pi, N.L., Zhao, X.Q., Zhang, S.L. and Zhao, D.H., 1986. The preliminary study on optimum stocking rate in alpine meadow. Acta Biol. Plateau Sin., 1986(5): 21–34 (in Chinese). Zimmerer, K.S., 1992. The loss and maintenance of native crops in mountain agriculture. GeoJournal. Z¨ottl, H., 1951. Die Vegetationsentwicklung auf Felsschutt in der alpinen and subalpinen Stufe des Wettersteingebirges. Jahrb. Ver. Schutz Alpenpﬂanzen Tiere, 16: 10–74.

122 Zuo, K.C. and Le, Y.Z., 1980. Formation and fertility of the alpine meadow soil in Qinghai Province. Acta Pedol. Sin., 17 (4): 308–318 (in Chinese). Zuo, K.C., Le, Y.Z., Zhang, J.X., Wang, Z.M., Zhao, B.L. and Bao, X.K., 1986. The cycling of nitrogen and phosphorus between the soil and plant nutrient pools in an alpine meadow ecosystem. In: The Proceedings of the International Symposium of

´ ´ and F.E. WIELGOLASKI V. KOMARKOV A Alpine Meadow Ecosystem. Northwest Plateau Institute of Biology, Academia Sinica, Xining, pp. 59–61. Zwally, H.J., 1989. Growth of Greenland ice sheet: interpretation. Science, 246: 1589–1591. Zwally, H.J., Brenner, A.C., Major, J.A., Bindschadler, R.A. and Marsh, J.G., 1989. Growth of the Greenland Ice Sheet: measurement. Science, 246: 1587–1589.

Chapter 4

ECOLOGICAL EFFECTS OF EROSION David PIMENTEL and Celia HARVEY

INTRODUCTION

CAUSES OF EROSION

Erosion of soil from land areas is widespread, and adversely affects all natural and human-managed ecosystems, including agriculture and forestry. Soil erosion ranks as one of the most serious environmental problems in the world because its effects are pervasive and long-lasting. Although soil erosion has occurred throughout history, it has intensiﬁed as growing human populations and their diverse activities have increased and intruded into natural ecosystems. Each year, approximately 75×109 tons of fertile soil are eroded worldwide from agricultural systems (Myers, 1993), and several billion additional tons are lost from other land areas. Erosion of soil degrades natural, agricultural, and forest ecosystems by reducing both the productivity of soils and the diversity of associated animals and microbes. Together, these losses diminish biodiversity and ecosystem stability (Pimentel et al., 1995a). In an effort to offset the deleterious effects of erosion on crop production irrigation is extensively used, and large quantities of fertilizers and pesticides are applied. Not only are these inputs dependent on fossil energy, but they can also harm human health and the environment (Pimentel et al., 1995a). When agricultural land is eroded to the point at which it is no longer productive, it is abandoned. Then forests, grassland, and wetlands are cleared to provide the extra agricultural land needed (Myers, 1989). Indeed, erosion – indirectly – is the major cause of the deforestation now taking place throughout the world. This chapter focuses on global dimensions of soil erosion and the impact erosion has on both natural and managed terrestrial ecosystems. Attention is given to both temperate and tropical ecosystems.

Erosion occurs when water or wind energy strikes exposed soil. Raindrops hit exposed soil with great energy, and launch soil particles with the water into the air. Raindrop splash and sheet erosion remove a thin ﬁlm of soil from the land surface, and are the dominant forms of erosion (Allison, 1973; Foster et al., 1985). The impacts of both are intensiﬁed on sloping land, where more than half of the soil contained in the splashes is carried downhill (Pimentel et al., 1995a). Wind also dislodges soil particles and carries them off the land. Airborne soil particles are often transported thousands of kilometers. For instance, soil particles eroded from African ecosystems are blown as far west as Brazil and Florida (Simons, 1992), and Chinese soil eroded during spring plowing is found deposited in distant Hawaii (Parrington et al., 1983). Land areas covered by living and dead plant biomass experience reduced soil erosion, because raindrop and wind energy is dissipated by the canopy or litter layer. In Missouri, for example, barren land lost soil 123 times faster than land covered with sod (dense grass), which lost soil at less than 0.1 t ha−1 yr−1 (United States Forest Service, 1936). In Utah and Montana, as the amount of ground cover decreased from 100% to less than 20%, erosion rates increased by a factor of approximately 180 (Fig. 4.1). Loss of vegetative cover is especially widespread in developing countries because population densities are high, agricultural practices frequently are poor, and serious fuel shortages exist for cooking and heating. For example, about 60% of crop residues in China, and 90% in Bangladesh, are stripped from the land and burned as fuel (Wen, 1993). In areas where fuelwood and other biomass is scarce, even the roots of grasses

123

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David PIMENTEL and Celia HARVEY

Erosion (t/ha)

20

10

0 0

20

40

60

80

100

120

% Ground Cover Fig. 4.1. Soil erosion rates related to percentage of ground cover in Utah and Montana (after Trimble and Mendel, 1995).

and shrubs are collected and burned (McLauglin, 1993). These practices leave the soil surface barren and exposed to rainfall and wind energy. Marginal land, often with steep slopes, is being converted from forests to agricultural use, not only to meet the food crop needs of a rapidly growing world population but also to replace the eroded, unproductive land (Lal and Stewart, 1990). Erosion rates on slopes are high. In Nigeria, ﬁelds of cassava (Manihot utilissima) on ~12% slopes lost 221 t ha−1 yr−1 , compared with a loss of 3 t ha−1 yr−1 on relatively ﬂat land (<1% slope) (Aina et al., 1977). In the Philippines, where over 58% of the land has slopes greater than 11%, and Jamaica, where 52% of the land has slopes greater than 20%, soil erosion rates are as high as 400 t ha−1 yr−1 (Lal and Stewart, 1990). For forested areas, a minimum of 60% tree cover is necessary to prevent soil erosion (Singh and Kaur, 1989). The Himalayan region of India was once heavily forested. Now only 35% of the region is covered with forests, and soil erosion and landslides are common problems. Hawley and Dymond (1988) reported that the area damaged by landslides in the Himalaya was increased from an average of 14% in regions where trees remain to cover the slopes to 70% in deforested areas. Although world agricultural production accounts for about three-quarters of the soil erosion worldwide, signiﬁcant erosion also occurs in other human-managed ecosystems (El-Swaify et al., 1985; Lal and Stewart, 1990). The construction of roadways, parking lots, and buildings illustrates this problem. Fortunately, erosion associated with construction is relatively brief,

generally lasting only while the construction disturbs the land surface. Once the disturbed land surface has grass or other vegetation established, erosion decreases (International Erosion Control Association, 1991). Natural areas also erode along stream banks and on steep slopes, where erosion occurs from the powerful action of moving water. On steep slopes (>30%), a stream cut through adjacent land can cause signiﬁcant loss of soil (Alonso and Combs, 1989). Even on only a 2% slope, stream banks erode, especially during heavy rains and ﬂooding. Cattle in and around streams further increase streambank erosion. For example, in Wisconsin a stream area inhabited by cattle lost about 60 t of soil along each kilometer of stream length each year (Trimble, 1994; Trimble and Mendel, 1995). Portions of the stream without cattle lost only trace amounts of soil. Erosion also accompanies landslides and earthquakes (Bruijnzeel, 1990). Overall, the erosion impact from earthquakes is fairly minimal because these events are relatively rare worldwide. Landslides are more frequent, and the damage from them more widespread than that of earthquakes. However, in the Caribbean basin region landslides are reported to affect less than 3% of the forested landscape per century (Walker et al., 1996). The incidence of landslides is often increased because of human activities such as the construction of roads and buildings or the removal of forests (Walker et al., 1996). The texture, depth and structure of a given soil, the amount of vegetative cover, the slope of the land, and the extent of land manipulation, all inﬂuence the soil’s susceptibility to erosion. Soils with medium to ﬁne texture, low organic matter content, and weak structural development are most easily eroded. Typically these soils have low inﬁltration rates or capacities and, therefore, are subject to high rates of water runoff, the eroded soil being carried away in the water ﬂow (Foster et al., 1985).

EXTENT OF SOIL EROSION

Erosion of soil particles from the land has always taken place on earth. Erosion in natural ecosystems is relatively slow, but its impact throughout the terrestrial system over millions of years has been signiﬁcant. Worldwide, the rate of natural erosion ranges from 1 to 5 t ha−1 yr−1 for mountainous regions with minimal vegetative cover, and from 0.001 to 2 t ha−1 yr−1 for

ECOLOGICAL EFFECTS OF EROSION

relatively ﬂat lands with vegetative cover. Even at relatively low erosion rates, the extent of soil movement over millions of years can be signiﬁcant and widespread (Lal and Stewart, 1990). Often, eroded soil accumulates in valleys, forming vast alluvial plains. For example, over a period of 100 years an erosion rate of 2 t ha−1 yr−1 on 10 ha deposits soil equivalent to a soil depth of 15 cm over 1 ha of land. Agricultural land Nearly one-half of the land surface of the world is devoted to agriculture. Of this, about one-third (1.5×109 ha) is planted to crops and two-thirds is pastureland (United States Department of Agriculture, 1993). Of the two, cropland is more susceptible to erosion than pastureland because it is tilled repeatedly, exposing the soil to wind and water erosion, and often left bare between plantings for several months of the year. Erosion on agricultural land is estimated to be 75 times greater than that occurring in natural forests (Myers, 1993). On some croplands, especially in developing countries, up to 100 to 200 t ha−1 yr−1 of soil are eroded by rainfall and wind, separately or in combination. In extreme circumstances, erosion may exceed 400 t ha−1 yr−1 (Hurni, 1985; Lal and Stewart, 1990; Huang, 1996). Currently, about 80% of the world’s agricultural land suffers moderate to severe erosion, whereas only 10% has relatively slight erosion (Pimentel, 1993; Speth, 1994; Lal, 1994). Worldwide, erosion on cropland averages about 30 t ha−1 yr−1 , ranging from 0.5 to 500 t ha−1 yr−1 (Pimentel et al., 1995a). As a result of erosion, during the last 40 years about 30% of the world’s arable land has become unproductive, thereby causing it to be abandoned for agricultural use (World Resources Institute, 1994). The area of arable land now under cultivation for crop production (nearly 1.5×109 ha) is about equal to the amount of arable land (2×109 ha) that has been abandoned by humans since farming began (Lal, 1990, 1994). The abandoned land, once biologically productive, now produces little biomass and has lost most of its biodiversity (Pimentel et al., 1992; Heywood, 1995). Erosion rates on cropland in the United States and Europe average about 13 t ha−1 yr−1 (Barrow, 1991; United States Department of Agriculture, 1994). Even these relatively low rates, however, greatly exceed the average rate of natural soil formation, which ranges from 0.5 to 1 t ha−1 yr−1 (Troeh and Thompson, 1993;

125

Lal, 1994; Pimentel et al., 1995a). Speciﬁcally, more than 90% of United States cropland now is losing soil above the sustainable replacement level (Hudson, 1982; Lal, 1984). Similar rates of soil erosion also occur on irrigated lands. In the United States, erosion is severe in some of the most productive agricultural ecosystems. For instance, one-half of the fertile topsoil of Iowa has been lost during the last 150 years of farming because of erosion (Risser, 1981; Klee, 1991). Unfortunately, high rates of erosion continue in Iowa at a rate of about 30 t ha−1 yr−1 , because of the rolling hills and the type of agriculture practiced (United States Department of Agriculture, 1989). Similarly, 40% of the productive soils on the Palouse, a type of grassland in the northwestern United States, has been lost in the past 100 years of cultivation. Both of these regions support a high-input agricultural system, generally one with monocultural plantings. In addition, if the ﬁelds are left unplanted during the fall and winter months, wind and rainfall will speed erosion. Worldwide, soil erosion rates are highest in agroecosystems located in Asia, Africa, and South America, averaging 30 to 40 t ha−1 yr−1 . In developing countries, soil erosion is particularly severe on small farms because they often occupy marginal lands, where soil productivity is poor and the topography is hilly. In addition, many poor farmers raise row crops, such as corn (Zea mays), which leave the soil highly susceptible to erosion (Southgate and Whitaker, 1992). For example, in the Sierra region of northeastern Ecuador, 60% of the cropland recently had to be abandoned because inappropriate agricultural practices had caused severe soil erosion and degradation (Southgate and Whitaker, 1992). Similar problems are evident in the Amazonian region of South America, especially where forested areas are being cleared to provide more land for crops and livestock. Pasture land In contrast to the average soil loss of 13 t ha−1 yr−1 from cropland in the United States, pastures lose about 6 t ha−1 yr−1 (United States Department of Agriculture, 1994). However, erosion rates intensify whenever overgrazing occurs, as is now occurring on more than half of the world’s pasture land (World Resources Institute, 1994). In many developing countries, heavy grazing by sheep and goats has removed most of the vegetative cover, exposing the soil to the energy of

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David PIMENTEL and Celia HARVEY

rainfall and wind. Even in the United States, about 54% of the pasture land, including Federal land, is now overgrazed and subject to high erosion rates (Hood and Morgan, 1972; Byington, 1986).

reduced fertility of the soil (Dregne, 1992). Such major reductions in food-crop yields are particularly serious at this time in history when the growing human population continues to require increased quantities of food.

Forest land

Water

In stable forest ecosystems, where soil is protected from rain and wind energy, erosion rates are relatively low, ranging from only 0.004 to 0.05 t ha−1 yr−1 (Bennett, 1939; Roose, 1988; Lal, 1994). In forests, tree leaves and branches intercept and diminish rain and wind energy, while leaf and branch litter on the forest ﬂoor provide additional protection. However, this changes when forests are cleared for crop production or for pasture. For example, in Ecuador, the Ministry of Agriculture and Livestock reported that 84% of the soils in the hilly, forested, northeastern part of the country should never have been cleared for pastures because of the high erosion potential of the soils and their limited fertility (Southgate and Whitaker, 1992).

Water is a primary limiting factor for productivity in most terrestrial ecosystems, because all vegetation requires enormous quantities of water for growth and reproduction (National Soil Erosion–Soil Production Research Planning Committee, 1981; Follett and Stewart, 1985; Falkenmark, 1989). For example, one hectare of corn or wheat (Triticum aestivum) transpires more than 4×106 liters of water each growing season (Leyton, 1983), and loses an additional 2×106 liters by evaporation from the soil (Waldren, 1983; Donahue et al., 1990). When erosion occurs, the amount of water runoff signiﬁcantly increases, and with less water entering the soil, less is available to support the growing vegetation (Table 4.1). Moderately eroded soils absorb from 10 to 300 mm less water annually from rainfall than uneroded soils. This represents a 7% to 44% decrease in the water available to growing vegetation (Wendt and Burwell, 1985; Wendt et al., 1986; Murphee and McGregor, 1991). A diminished absorption rate of 20% to 30% of rainfall represents signiﬁcant water shortages for plants, including crops (Elwell, 1985). Lal (1976) reported that erosion reduced inﬁltration in some tropical soils by up to 93%. In general, when water availability in soil is reduced by 20% to 40%, plant productivity is reduced by 10% to 25% depending also on total rainfall, soil type, and slope, as well as other factors. Such major reductions in plant biomass also reduce the soil biota, as well as the overall biodiversity within the ecosystem (Heywood, 1995).

EFFECTS OF EROSION ON PRODUCTIVITY

Erosion reduces the overall productivity of terrestrial ecosystems in several ways. First in order of importance, erosion increases water runoff, thereby decreasing water inﬁltration and water storage in the soil (Troeh et al., 1991; Pimentel et al., 1995a). Also, organic matter and essential plant nutrients are lost and soil depth is reduced in the erosion process. These changes reduce biological activity and biodiversity in the soil (Troeh et al., 1991; Pimentel et al., 1995a). Because these factors interact with one another, it is almost impossible to separate the speciﬁc impacts of one factor from another. For example, the loss of soil organic matter increases water runoff, which reduces water-storage capacity. This diminishes nutrient levels in the soil, and also reduces the natural biomass and the biodiversity of the entire ecosystem. Overall, the cumulative effects of erosion directly diminish plant productivity. For example, erosion reduced corn productivity by 12% to 21% in Kentucky, up to 24% in Illinois and Indiana, 25% to 65% in the southern Piedmont of Georgia, and 21% in Michigan (Frye et al., 1982; Olson and Nizeyimana, 1988; Mokma and Sietz, 1992). In the Philippines, erosion caused declines of as much as 80% in corn production over a period of 15 years because of the

Nutrients When soil is eroded, such basic plant nutrients as nitrogen, phosphorus, potassium, and calcium also are lost. The soil removed by erosion typically contains about three times more nutrients than the soil left behind on the eroded land (Lal, 1980; Young, 1989). One ton of fertile topsoil may contain 1 to 6 kg of nitrogen, 1 to 3 kg of phosphorus, and 2 to 30 kg of potassium, whereas a soil on eroded land frequently has nitrogen levels of only 0.1 to 0.5 kg

ECOLOGICAL EFFECTS OF EROSION

127

Table 4.1 Water runoff rates for conservation plantings of corn (Zea mays) versus conventional plantings of corn 1 Treatment

Water runoff (cm depth)

Corn stover mulch left in place 3

0.06

Corn without stover

1.30

Corn stover plus rye mulch 4 Corn residue burned Corn grown in manure mulch 5 Corn without manure mulch

3.90

Increased yield 2 (t ha−1 )

References

1.24

0.34

Ketcheson and Onderdonk (1973)

13.50

3.40

Klausner et al. (1974)

4.10

1.10

Musgrave and Neal (1937)

2.50

0.60

Ketcheson (1977)

7.00

1.80

Spomer et al. (1976)

7.20

1.80

Schuman et al. (1973)

1.50

0.40

McIsaac and Mitchell (1992)

17.40 9.00 13.10

Corn grown after an oats–hay–hay rotation 6

0.58

Conventional continuous corn

3.08

Corn planted in no-till sod 7

3.70

Conventional corn

Conserved water (cm)

10.70

Corn grown on level terraces 8

0.94

Corn planted on contour

8.14

Corn grown on reduced till 9

2.10

Conventional corn

3.60

1

These results were all obtained in the Corn Belt of the United States and adjacent areas of Canada, under an average annual rainfall of about 1000 mm. The control plots in each case were subject to the normal procedures followed in the Corn Belt – the crop is grown year after year on the same land; on sloping land, rows are planted along the contours; crop residues are plowed under each year; and the soil is fully cultivated between successive crops, with herbicide applications for weed control and inter-row cultivation during the growth of the crop. 2 Increased yield based on the method of Troeh et al. (1991), to ﬁnd equivalent yields with varying amounts of moisture. 3 Water runoff rates were measured with and without corn stover mulch on sloping land ranging from 7% to 9%. The experiments were carried out in Guelph, Ontario. 4 Water runoff rates were measured with corn stover mulch plus rye-grass (Lolium perenne) mulch in position, and when the corn stover was burned. Soil was moderately to somewhat poorly drained. Experiments carried out in Aurora, New York. 5 Water runoff rates were measured with corn grown with a manure mulch, and corn grown without a manure mulch. Experiments conducted on Marshall silt-loam soil in Clarina, Iowa. 6 Water runoff rates were measured with corn grown after an oats (Avena sativa)–hay–hay rotation, and with corn grown continuously, in the conventional way. Experiments were conducted on a Guelph loam soil in Guelph, Ontario. 7 Water runoff rates were measured with corn grown in sod as a no-till system, and with corn grown conventionally. Experiments were conducted on western Iowa loess soil in Ames, Iowa. 8 Water runoff rates were measured with corn grown on ﬂat terraces, and corn grown on the contour; slopes ranged from 2 to 18%. Experiments were conducted on Monona Ida and Napier silt loams in Lincoln, Nebraska. 9 Water runoff rates were measured with corn grown on reduced till and with conventionally grown corn. Experiments were conducted on a Catlin silt loam soil in Champaign, Illinois.

per ton (Alexander, 1977; Troeh et al., 1991). Plant productivity is signiﬁcantly reduced when soil nutrient levels are at this low level. Soil organic matter Both wind and water erosion selectively remove the ﬁne organic particles in the soil, leaving behind larger particles and stones. Fertile soils frequently contain

about 100 t ha−1 of organic matter, or 4% of total soil mass (Follett et al., 1987; Young, 1990). Because most of the organic matter is close to the soil surface in the form of decaying leaves and stems, erosion of the topsoil causes a signiﬁcant decrease in soil organic matter. Several studies have demonstrated that the soil removed by either wind or water erosion is 1.3 to 5 times richer in organic matter than the soil left behind (Allison, 1973; Barrows and Kilmer, 1963).

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Soil organic matter facilitates the formation of soil aggregates, and increases soil porosity. In this way it improves soil structure, which in turn facilitates water inﬁltration, and ultimately overall productivity (Chaney and Swift, 1984; Langdale et al., 1992). In addition, organic matter aids cation exchange, enhances root growth, and increases the populations of important soil organisms (Allison, 1973). About 95% of the soil nitrogen and 25% to 50% of the phosphorus is contained in organic matter (Allison, 1973). Once the organic-matter layer is depleted, the productivity of the ecosystem – as measured by cropplant yields – declines, both because of the degraded soil structure and the loss of nutrients contained in the organic matter. For example, in a Michigan study the reduction of soil organic matter from 4.3% to 1.7% lowered the potential productivity for corn by 25% (Lucas et al., 1977). When soil nutrients are depleted by erosion, plant growth is stunted and productivity declines (Pimentel et al., 1995a). Soils that suffer severe erosion may produce crop yields 15% to 50% lower than uneroded soils (Follett and Stewart, 1985; Olson and Nizeyimana, 1988; Schertz et al., 1989; Langdale et al., 1992). In addition to low crop yields, the total biomass of the biota and overall biodiversity of the ecosystem is substantially reduced in eroded soils (Heywood, 1995).

David PIMENTEL and Celia HARVEY

and the percolation of water into the soil. These changes then increase the rate of water runoff, further intensifying soil erosion. With reduced soil organic matter, reduced silt, and reduced soil water availability, the quality of the soil as a seedbed is reduced. If soil erosion continues, eventually the soil becomes a coarse sand with many rocks or primarily a clay type soil with little productivity. Soil loss Where does the eroded soil go? Earlier we mentioned the enormous amounts of soil that are blown off the African continent and are found in Florida and South America. Actually, relatively little reaches Florida and South America; most of the soil ends up in the Atlantic Ocean. The same applies to the soil that is blown from China to Hawaii. Most of this wind-eroded soil ends up in the Paciﬁc Ocean. Not all soil that is removed by the wind ends up in the oceans. While a small amount is deposited on cropland, most is deposited in forests and pasture lands of the region where the erosion takes place. Some soil eroded by rainfall is washed down slopes and does not leave the farm. However, most of the eroded soil (more than 60%) ends up in streams, lakes, and reservoirs after it is loosened by rain drops and moves down slopes (Pimentel et al., 1995a).

Soil depth When erosion reduces total topsoil depth – especially along with changes in soil organic matter, water availability, and nutrients – the number of species present in the soil is signiﬁcantly reduced (Pimentel et al., 1995a). When total soil depth is reduced from 30 cm to less than 1 cm, plant root space is minimal and, concurrently, valuable soil organisms nearly disappear. Soil structure Soil structure is important for plant growth because it provides the following valuable attributes: (i) suitable aeration; (ii) the ability of water to percolate into the soil; (iii) resistance to erosion; (iv) formation of a productive seedbed. Erosion reduces the quality of these soil characteristics. When erosion occurs, the pores and holes in the soil are ﬁlled with ﬁne soil particles, which reduces both aeration of the soil

IRRIGATION, WATER, AND SOIL EROSION

Both water and energy resources are expended in the irrigation of arid land in order to make it productive. Approximately 16% of the world’s cropland is irrigated (World Resources Institute, 1992), and about 33% of the world’s food is produced on this irrigated land (Postel, 1992). Worldwide, the amount of land under irrigation is slowly expanding, even though salinization, waterlogging, and siltation are decreasing the productivity of some irrigated lands (D. Haith, pers. comm., 1994). Despite a small annual increase in irrigated areas, the area per person of the population has been declining since 1978 (Postel, 1992). For example, the area per person of irrigated land in the United States declined 8% between 1978 and 1988 (United States Department of Agriculture, 1993). Because crops require large quantities of water for their growth, whether from irrigation or rainfall, it is

ECOLOGICAL EFFECTS OF EROSION

129

vital that as much water as possible percolate into the soil instead of running off. Soil erosion often limits the amount of water that is available for crop use (Lal and Stewart, 1990). When raindrops hit exposed soil they have an explosive effect, launching soil particles into the air. If the water does not percolate into the soil, it runs off and carries soil with it. More than half of the soil contained in the splashes is carried downhill on land with a slope greater than 1% (Foster et al., 1985). In most ﬁelds, raindrop splash and sheet erosion are the dominant forms of erosion (Foster et al., 1985). As expected, loss of rain water and/or irrigation water severely reduces crop productivity. The amounts of water required to produce a crop range from about 500 to 2000 ° kg−1 (Table 4.2). Conserving water and minimizing runoff and erosion beneﬁt agricultural production and especially help to reduce the need for irrigation. Irrigation is costly in terms of water, energy, and money. It should be emphasized again that some irrigation practices themselves can cause serious erosion and waste water (Pimentel et al., 1997). Table 4.2 Estimated liters of water required to produce 1 kg of agricultural products (Pimentel et al., 1997) Product

° kg−1

Potatoes (Solanum tuberosum)

500

Wheat (Triticum aestivum)

900

Alfalfa (Medicago sativa)

900

Sorghum (Sorghum vulgare)

1110

Corn (Zea mays)

1400

Rice (Oryza sativa)

1912

Soybeans (Soja max)

2000

Broiler fowls (Gallus domesticus)

3500

Beef (Bos taurus)

100 000

SOIL EROSION CONTROVERSY

Unfortunately, there are few quantitative data concerning erosion rates and the extent of erosion in agriculture and forestry. Earlier it was mentioned that about 80% of the world’s agricultural land suffers from moderate to severe erosion (Speth, 1994). Oldeman et al. (1990) have provided an estimate for moderate to severe erosion similar to that of Speth (1994); however, there

is a great difference in what is deﬁned as severe by these scientists. Speth’s deﬁnition of severely eroded land includes land that remains productive, although level of productivity may be low. Oldeman et al. deﬁned severely eroded land as land that is totally unproductive and should be abandoned. Clearly, these are major differences in the estimates of what constitutes moderate to severe erosion. The conclusion of another worldwide survey of the soil-erosion problem by soil scientists representing many different nations generally agrees with Speth’s assessment (Pimentel, 1993). Crosson (1985), an economist, estimated that nutrients worth about $500×106 are lost annually in the United States. However, Troeh et al. (1991) – soil scientists – estimated that the total annual loss of nutrients was worth $20×109 . Thus, the estimate by Troeh et al. was 40 times larger than that of Crosson. Crosson (1995) estimated that nutrients and all other costs of erosion at that time totaled only $100 to $120×106 yr−1 . Crosson gave no explanation as to why he signiﬁcantly reduced his estimates. A related estimate by Pimentel et al. (1995a) of $27×109 for nutrients and all other on-farm costs was about 250 times higher than Crosson’s revised estimate. The major reason for the differences between Crosson (1995) and Pimentel et al. (1995b) is that Crosson relied on models to develop his results, whereas Pimentel et al. (1995b) used ﬁeld data produced by soil scientists. Follett and Stewart (1985) have highlighted this type of controversy and the differences in results and conclusions between modellers and soil scientists. We believe that models are important, but that they cannot be used as a substitute for ﬁeld data from experiments (Pimentel et al., 1995b). BIOMASS AND BIODIVERSITY

By diminishing soil organic matter and overall soil quality, erosion reduces biomass productivity in ecosystems. Ultimately, this has a profound effect on the diversity of plants, animals, microbes and other forms of life present in the ecosystem. The biological diversity existing in any ecosystem is directly related to the amount of living and nonliving organic matter present in that ecosystem (Wright, 1983, 1990). Numerous positive correlations, from low to optimum levels of biomass abundance, have been established between biomass and species diversity (Elton, 1927; Odum, 1978; Sugden and Rands, 1990; M. Giampietro, 1991, pers. commun.). Vegetation is the

130

David PIMENTEL and Celia HARVEY

Table 4.3 Biomass of various groups of organisms in a pasture at Ithaca (New York State) Organisms

Biomass (kg ha−1 fresh weight)

Reference

Plants

20 000

Fungi

4000

Walter (1985) Richards (1974)

Bacteria

3000

Richards (1974)

Arthropods

1000

Walter (1985)

Annelids

1320

Richards (1974)

Protozoa

380

Richards (1974)

Algae

200

Walter (1985)

Nematodes

120

Walter (1985)

Mammals

1.2

Pimentel (1985) unpublished

Birds

0.3

Pimentel (1985) unpublished

main component of ecosystem biomass and provides the resources needed by animals and microbes. This relationship is illustrated by data from a temperate grassland in New York State (Table 4.3). The standing plant biomass (fresh weight) was estimated to be 20 000 kg−1 ha−1 . The biomass of fungi and bacteria totaled 7000 kg−1 ha−1 and that of arthropods and annelids totaled 2320 kg−1 ha−1 . The remaining groups of organisms totaled only about 700 kg−1 ha−1 . The smallest biomass was only 1.2 kg−1 ha−1 for mammals and 0.3 kg−1 ha−1 for birds. Living organisms are a vital component of the soil and constitute a large percentage of the soil biomass. One square meter of soil may support about 200 000 arthropods and enchytraeids, and billions of microbes (Wood, 1989; Lee and Foster, 1991). Anderson (1978) reported that a favorable temperate-forest soil with abundant organic matter supports up to 1000 species of animals in a square meter, including arthropods, nematodes, and protozoa. Soil bacteria and fungi add another 4000 to 5000 species to the biodiversity in moist, organic forest soils (Heywood, 1995). Erosion rates which are 10 to 20 times the rate of soil formation (from less than 0.5 to 1 t ha−1 yr−1 ) decrease the diversity and abundance of soil organisms (Atlavinyte, 1964; 1965), whereas agricultural practices that maintain the soil organic-matter content favor the proliferation of soil biota (Reid, 1985). For example, the simple practice of adding straw mulch to the soil surface increased soil organic matter and the size of

the biota as much as three-fold (Teotia et al., 1950). Similarly, the application of organic matter or manure increased earthworm and microorganism biomass as much as ﬁve-fold (Ricou, 1979). In the former Soviet Union, species diversity of macrofauna (mostly arthropods) increased 16% when organic manure was added to experimental wheat plots (Bohac and Pokarzhevsky, 1987). Similarly, species diversity of the macrofauna in Japan (mostly arthropods) more than doubled when organic manure was added to grassland plots (Kitazawa and Kitazawa, 1980). In the United Kingdom, although data for species diversity were not presented, the biomass of arthropods increased from 2 to 7-fold when organic matter, as manure, was added to crops of either wheat or mangold (Beta vulgaris) (Morris, 1922; Raw, 1967). Also, when organic manure was added to agricultural land in Hungary, the biomass of soil microbes increased ten-fold (Olah-Zsupos and Helmeczi, 1987). Because increased biomass in soil generally is correlated with increased biodiversity, it is logical to assume that the increase in biomass of arthropods and microbes represents an increase in biodiversity (Pimentel et al., 1992). The relationship between plant biomass and animal biodiversity was further illustrated in ﬁeld experiments with collards (Brassica oleracea acephala). The species diversity of arthropods was four times higher on experimental plots amended with organic wastes relative to the control, which was associated with higher collard biomass (Pimentel and Warneke, 1989). Ward and Lakhani (1977) reported that the number of arthropod species associated with an ecosystem containing juniper bushes (Juniperus sp.) increased four-fold when the density of bushes was increased ten-fold. Elsewhere, a strong correlation between plant biomass productivity and bird species diversity was reported; a hundred-fold increase in plant biomass productivity yielded a ten-fold increase in bird diversity (Wright, 1983, 1990). Erosion frequently has indirect effects on ecosystems that may be nearly as damaging as the direct effect of reduced plant productivity. For example, the stability and biodiversity of grasslands were reported to be signiﬁcantly reduced when plant species diversity was reduced (Tilman and Downing, 1994). As plant species diversity decreased from 25 to 5 or fewer species, the grassland became less resistant to drought. The total amount of biomass declined to less than a quarter. The species-poor sites were more susceptible to drought

ECOLOGICAL EFFECTS OF EROSION

and more time was required to recover their original productivity than the species-rich sites. Erosion can cause the loss of a keystone species, whose absence may have a cascading effect on a wide array of other species within the ecosystem. Keystone species include plants that maintain the productivity and integrity of the ecosystem: predators and parasites that control the feeding pressure of some organisms on vital plants, pollinators of various important plants, seed dispersers, and plants and animals that provide a habitat required by other essential species like biological nitrogen-ﬁxers (Paine, 1966; Cox et al., 1991; Heywood, 1995). In these ways, the regular activities within an ecosystem may be interrupted or eliminated when a keystone species is lost. The impacts of this loss can be particularly severe in agroecosystems when, for instance, pollinator populations are drastically reduced and/or eliminated. Many soil organisms perform beneﬁcial activities that improve soil quality and, ultimately, its productivity. For example, soil organisms recycle basic nutrients required by plants for their growth (Van Rhee, 1965; Pimentel et al., 1980). The tunneling and burrowing of earthworms and other soil organisms enhance productivity by increasing water inﬁltration. Earthworms can produce up to 220 tunnel openings (3 to 5 mm in diameter) per square meter, enabling water to run rapidly into the soil and increasing inﬁltration rates (Anderson, 1988). Certain soil organisms contribute to soil formation and productivity by mixing the soil, enhancing aggregate stability, and preventing soil crusting. Earthworms bring from underground to the soil surface between 10 and 500 t ha−1 yr−1 of soil (Edwards, 1981; Lavelle, 1983; Lee, 1985), while insects bring a smaller amount to the surface (Hole, 1981; Zacharias and Grube, 1984; Lockaby and Adams, 1985). The churning and mixing of the upper soil redistributes nutrients, aerates the soil, exposes soil to the weather for soil formation, and increases inﬁltration rates, thus making conditions favorable for increased soil formation and plant productivity. In arid regions like the Negev desert, Euchordrus snails also help form soil by consuming lichens and rock on which the lichens are growing (Shachak et al., 1995). This snail activity contributes about 1000 kg of soil per hectare per year, which is equal to the annual wind-borne deposits in the same area.

131 SEDIMENTS AND WIND-BLOWN SOIL PARTICLES

Beyond its direct effect on agricultural and forestry ecosystems, the impact of erosion reaches far into the surrounding environment. The major off-site problems include earth dam failures, eutrophication of waterways, siltation of harbors and channels, loss of reservoir storage, loss of wildlife habitat and disruption of stream ecology, ﬂooding of land and communities, plus increased need for treatment of water for human consumption (Gray and Leiser, 1989). The most costly off-site damage is caused by soil particles entering water systems (Lal and Stewart, 1990). Of the billions of tons of soil lost from the world’s cropland, nearly two-thirds is ﬁnally deposited in streams and rivers (United States Department of Agriculture, 1989). These sediments harm aquatic ecosystems by contaminating the water directly with soil particles, increasing turbidity, and on account of the fertilizers and pesticides they contain (Clark, 1987). In addition, siltation of reservoirs and dams reduces water-storage volume, increases their maintenance costs, and shortens their useful lifetime (Pimentel et al., 1995a). Furthermore, heavy sedimentation frequently exacerbates river and lake ﬂooding (Myers, 1993). For example, some of the ﬂooding that occurred in the midwestern United States during the summer of 1993 was caused by increased sediment deposition in the Mississippi and Missouri Rivers and their tributaries. Sedimentation raised the water level of these waterways, making them more prone to ﬂooding (Allen, 1994). Wind-eroded soil also causes signiﬁcant off-site damage and costs because soil particles propelled by strong winds act as abrasives and air pollutants. Estimates are that the sand-blasting of automobiles and buildings by soil particles in the United States causes about $8×109 in damage per year (Huszar and Piper, 1985; Soil Conservation Service, 1993; Pimentel et al., 1995a). Wind blown soil may also sand-blast crop and natural plants, in some situations killing the plants. This emphasizes the importance of vegetative cover including shelter belts to prevent or control wind erosion. A prime example of the environmental impact of wind erosion occurs in New Mexico, U.S.A., where about two-thirds of the land is used for agriculture, including grazing. In New Mexico, erosion rates on pastures often exceed 6 t ha−1 yr−1 , and frequently reach

132

as high as 100 t ha−1 yr−1 . Yearly off-site erosion costs in this state, including health and property damage, are estimated to reach $465×106 (Huszar and Piper, 1985). Assuming similar costs for other states, the offsite damage from wind erosion alone could cost nearly $10×109 annually (Pimentel et al., 1995a). Soil erosion also contributes to the global warming problem by adding carbon dioxide to the atmosphere because of the oxidation of enormous amounts of biomass carbon in the soil, similar to plowing and exposing organic matter (Phillips et al., 1993). As mentioned, a hectare of soil may contain about 100 tons of organic matter or biomass. When the forces of erosion uncover this organic matter, the carbon in it is oxidized with the release of carbon dioxide. The subsequent release of carbon dioxide into the atmosphere contributes greatly to the global-warming problem (Phillips et al., 1993). In fact, a feedback mechanism may exist since increased global warming intensiﬁes rainfall, which increases erosion (Lal, 1990). Then, with more erosion, more organic matter is exposed and is oxidized to carbon dioxide, thereby further increasing global warming. CONCLUSION

Soil erosion has become a critical environmental problem throughout the world’s terrestrial ecosystems. Erosion is a relentless, insidious process. A loss of one millimeter of soil, which can easily occur in one rain or wind-storm, can easily go unnoticed. Yet the loss of 1 mm of soil over 1 ha of cropland amounts to 15 tons. Re-forming or replacing this amount of soil under natural circumstances requires an average of 20 years. Erosion inﬂicts multiple serious damage in managed ecosystems like crops, pastures or forests – as well as in natural ecosystems. In particular, erosion reduces the water holding capacity of the soil because of rapid water runoff and reduced soil organic matter. As a result, nutrients and valuable soil organisms are reduced. Separately or together these factors diminish the productivity of all vegetation and animals in ecosystems. At the same time, biodiversity is signiﬁcantly reduced. Worldwide, soil erosion continues unabated, while the human population and its requirements for food, ﬁber, and other resources increase geometrically. Indeed, achieving future food security depends on conserving existing soil, water, and biological resources.

David PIMENTEL and Celia HARVEY

Conservation must receive high priority for effective protection of managed and natural ecosystems. It is vital for human survival.

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David PIMENTEL and Celia HARVEY Black, P., Valdes-Cogliano, S., Vinzant, B., Hudes, E. and Littman, S., 1980. Environmental quality and natural biota. BioScience, 30: 750–755. Pimentel, D., Stachow, U., Takacs, D.A., Brubaker, H.W., Dumas, A.R., Meaney, J.J., O’Neill, J., Onsi, D.E. and Corzilius, D.B., 1992. Conserving biological diversity in agricultural/forestry systems. BioScience, 42: 354–362. Pimentel, D., Harvey, C., Resosudarmo, P., Sinclair, K., Kurz, D., McNair, D., Crist, S., Spritz, L., Fitton, L., Saffouri, L. and Blair, R., 1995a. Environmental and economic costs of soil erosion and conservation beneﬁts. Science, 267: 1117–1123. Pimentel, D., Harvey, C., Resosudarmo, P., Sinclair, K., Kurz, D., McNair, D., Crist, S., Spritz, L., Fitton, L., Saffouri, L. and Blair, R., 1995b. Soil erosion estimates and costs. Science, 269: 463–465. Pimentel, D., Houser, J., Preiss, E., White, O., Fand, H., Mesnick, L., Barsky, T., Tariche, S., Schreck, J. and Alpert, S., 1997. Water resources: agriculture, the environment, and society. BioScience, 47: 97–106. Postel, S., 1992. Last Oasis: Facing Water Scarcity. W.W. Norton, New York, 239 pp. Raw, F., 1967. Arthropods (except Acari and Collembola). In: A. Burges and F. Raw (Editors), Soil Biology. Academic Press, London, pp. 323–362. Reid, W.S., 1985. Regional effects of soil erosion on crop productivity – northeast. In: R.F. Follett and B.A. Stewart (Editors), Soil Erosion and Crop Productivity. American Society of Agronomy, Madison, Wisconsin, pp. 235–250. Richards, B.N., 1974. Introduction to the Soil Ecosystem. Longman, New York, 266 pp. Ricou, G.A.E., 1979. Consumers in meadows and pastures. In: R.T. Coupland (Editor), Grassland Ecosystems of the World: Analysis of Grasslands and Their Uses. Cambridge University Press, Cambridge, pp. 147–153. Risser, J., 1981. A renewed threat of soil erosion: it’s worse than the Dust Bowl. Smithsonian, 11: 120–122, 124, 126–130. Roose, E., 1988. Soil and water conservation lessons from steep-slope farming in French speaking countries of Africa. In: Conservation Farming on Steep Lands. Soil and Water Conservation Society, Ankeny, Iowa, pp. 130–131. Schertz, D.L., Moldenhauer, W.C., Livingston, S.J., Weesies, G.A. and Hinz, E.A., 1989. Effect of past soil erosion on crop productivity in Indiana. J. Soil Water Conservation, 44: 604–608. Schuman, G.C., Spomer, R.G. and Piest, R.G., 1973. Phosphorus losses from four agricultural watersheds on Missouri valley loess. Soil Sci. Soc. Am. Proc., 37: 424–427. Shachak, M., Jones, C.G., Brand, S., Blume, H.P. and Berkowicz, S.M., 1995. The role of animals in an arid ecosystem: snails and isopods as controllers of soil formation, erosion and desalinization. Advances in Geoecology Vol. 28, Arid Ecosystems, pp. 37–50. Simons, M., 1992. Winds toss Africa’s soil, feeding lands far away. New York Times, October 29, 1992, pp. A1, A16. Singh, T.V. and Kaur, J., 1989. Studies in Himalayan Ecology and Development Strategies. Himalayan Books, New Delhi, 286 pp. Soil Conservation Service, 1993. Wind Erosion Report (Nov. 1992– May 1993). Soil Conservation Service, USDA, Washington, D.C. Southgate, D. and Whitaker, M., 1992. Promoting resource degradation in Latin America: tropical deforestation, soil erosion, and coastal ecosystem disturbance in Ecuador. Econ. Dev. Cultural Change, 40: 787–807.

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135 Walker, L.R., Zarin, D.J., Fetcher, N., Myster, R.W. and Johnson, A.H., 1996. Ecosystem development and plant succession on landslides in the Caribbean. Biotropica, 28: 566–576. Walter, H., 1985. Vegetation of the Earth and Ecological Systems of the Geo-Biosphere. Springer-Verlag, New York. 318 pp. Ward, L.K. and Lakhani, K.H., 1977. The conservation of juniper: the fauna of food-plant island sites in southern England. J. Appl. Ecol., 14: 121–135. Wen, D., 1993. Soil erosion and conservation in China. In: D. Pimentel (Editor), Soil Erosion and Conservation. Cambridge University Press, New York, pp. 63–86. Wendt, R.C. and Burwell, R.E., 1985. Runoff and soil losses from conventional, reduced, and no-till corn. J. Soil Water Conservation, 40: 450–454. Wendt, R.C., Alberts, E.E. and Helmers, A.T., 1986. Variability of runoff and soil loss from fallow experimental plots. J. Soil Sci. Soc. Am., 50: 730–736. Wood, M., 1989. Soil Biology. Blackie, Chapman and Hall, New York, 154 pp. World Resources Institute, 1992. World Resources 1992–93. World Resources Institute, Washington, D.C., 389 pp. World Resources Institute, 1994. World Resources 1994–95. World Resources Institute, Washington, D.C., 400 pp. Wright, D.H., 1983. Species-energy theory: an extension of speciesarea theory. Oikos, 41: 496–506. Wright, D.H., 1990. Human impacts on energy ﬂow through natural ecosystems, and replications for species endangerment. Ambio, 19: 189–194. Young, A., 1989. Agroforestry for Soil Conservation. C.A.B. International, Wallingford, UK, 276 pp. Young, A., 1990. Agroforestry, environment and sustainability. Outlook Agriculture (Oxon, UK: C.A.B. International), 19: 155–160. Zacharias, T.P. and Grube, A.H., 1984. An economic evaluation of weed control methods used in combination with crop rotations: a stochastic dominance approach. N. Cent. J. Agric. Econ., 6: 113–120.

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Chapter 5

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY Roger DEL MORAL and Sergei Yu. GRISHIN

INTRODUCTION

Volcanic eruptions are major natural disturbances with varied and complex consequences. Recent studies of how plants colonize newly created volcanic surfaces already have begun to clarify the understanding of succession. Because volcanoes create disturbance mosaics, it is imperative to understand the biotic responses more fully. The 1883 eruption of Krakatau captured the world’s imagination and initiated widespread and intensive studies of succession following volcanic events (Docters van Leeuwen, 1936). In this chapter, we review the state of knowledge of succession after volcanic events and attempt to develop a general understanding of the mechanisms that affect the development of volcanic landscapes. Volcanic eruptions punctuate the geological record and human history. The eruption at Santorini (Thera) in Greece (~B.C. 1470) probably destroyed Minoan civilization, and the collapse of the Maya is attributed to climatic impacts caused by eruptions of El Chich´on in AD 900 and AD 1250 (Anonymous, 1993). Indirect volcanic effects also destroy. When Iceland’s Laki ﬁssure spawned the largest volcanic episode in recorded history (AD 1783), Icelanders starved, summer failed, and European markets crashed (Gratten and Braysay, 1995). Despite monitoring warnings, ash and mud ﬂows from the eruption of Mount Pinatubo, in the Philippines in 1991, killed 700 people, displaced 500 000 villagers, closed a United States air base, and eliminated 40 000 jobs. Ash veiled the planet, and sulfuric acid reduced the thickness of the ozone layer. Altered weather patterns led to extreme ﬂoods of the Mississippi River Valley in the central United States (Tizon, 1996). Volcanoes occur along tectonic plate margins and within plates (Francis, 1993). There are currently about

550 active land volcanoes. Volcanoes associated with ocean plates occur in the Maldives and R´eunion in the Indian Ocean, the Azores in the Atlantic Ocean, and Hawaii in the Paciﬁc Ocean. Continental plates once produced ﬂoods of lava from ﬁssures (e.g., the Indian Deccan plateau), but today most active volcanoes are along plate margins. Diverging plates spawn openocean volcanoes such as Surtsey, Iceland, or rifts, as in East Africa. Converging plates spawn huge earthquakes and volcanism, such as routinely occur along the rim of the Paciﬁc Basin (the “Ring of Fire”). This chapter describes typical volcanic disturbance regimes and their ecological consequences. How does the biota respond to these impacts? We will emphasize disturbance regimes, recovery mechanisms, and recovery rates.

VOLCANIC IMPACTS

New volcanic surfaces are created by lava, pyroclastic ﬂows and air-borne deposits (tephra), debris ﬂows, lahars, and avalanches (see Table 5.1 for deﬁnitions). Primary succession is associated with lavas, pyroclastic ﬂows, and most lahars, but may occur on any material. Secondary succession occurs most frequently on thin tephra and some lahars, because some biota or soil may survive (Halpern et al., 1990; Grishin et al., 1996). Volcanic gases create strong, but local, impacts. Once the gases subside, primary succession usually results, modiﬁed by the nature of the substrate in question (Fig. 5.1). Climate, substrate type, and landscape factors all affect the recovery rate. Volcanism is associated with earthquake zones, being produced by the same fundamental forces. Tsunamis are a consequence of both earthquakes and volcanic activity. These sea waves are usually caused by

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Table 5.1 Deﬁnitions of major types of volcanic materials mentioned in this chapter; materials which have little direct impact on succession are not listed

Lava

Term

Deﬁnition 1

Lava

molten rock (magma) formed of basalt, andesite, dacite, or rhyolite, rocks (in order of increasing silicate content)

a’a

basalt forming a jumble of irregular cindery blocks; low in silicates

pahoehoe

basalt forming smooth, glossy surfaces; low in silicates

Lavas cover much of the earth, but large ﬂows are rare. Lavas are viscous and ﬂow slowly – thus they are the least dangerous of volcanic events. Molten lavas may ooze from ﬁssures, vents, hot spots (e.g., R´eunion and Hawaii), and rifts (e.g., East Africa) to form new land. For ecological purposes, lavas can be divided into those forming a mass of small, loose, irregularly shaped blocks (known by the Hawaiian term a’a), smooth, ropy lavas (known by the Hawaiian term pahoehoe), large angular blocks, and domes found within craters.

Block lava

formed of large, smooth-sided blocks; may form domes; usually andesitic

Pyroclastic rocks

Domes

viscous intrusions found in craters, commonly of dacite or rhyolite

Pyroclastic rocks

material ejected in solid fragments

Pyroclastic ﬂows

an eruption cloud consisting of gas and very hot solids driven by gravity and hugging the ground; including surges and nu´ees ardentes (incandescent mix of ash and large materials); forming deposits called ignimbrites

Pyroclastic falls

any solid material returning to earth after eruption into the air, called tephra

Ash

tephra less than 4 mm in size

Lapilli (little stones)

tephra between 4 and 32 mm in size

Blocks (bombs)

tephra greater than 32 mm in size

Scoria

fragmented, cindery-textured pyroclastic material, whether from falls or ﬂows of any sort

Pumice

frothy, low-density pyroclastic rock deposited by ﬂows or falls

Lahar (mud-ﬂow)

all water-transported debris ﬂows, regardless of origin, with >50% water

Debris ﬂow

a large, wet mass of material falling under the force of gravity, with <50% water

Avalanche

a mass of nearly dry material falling under the force of gravity

1

Deﬁnitions follow Francis (1993).

coastal earthquakes, but a devastating tsunami resulted from the collapse of the Rakata cone of Krakatau (1883) and killed more humans than the eruption itself (Francis, 1993). The effects of earthquakes and tsunamis on an ecosystem may be ephemeral or persistent. Space does not permit a review of these effects here, but the consequences share much with the principles developed in this chapter.

Any solid material forcibly ejected from a volcano is formed amid great heat and pressure. These pyroclastic (Greek = ﬁre-broken) events are variously classiﬁed (cf. Chester, 1993; Francis, 1993). Forms leading to distinct ecological results are noted below. Pyroclastic ﬂows Pyroclastic ﬂows include any incandescent mixture of gas and solids produced by explosive eruptions (Francis, 1993). They move rapidly along the ground (>100 km hr−1 ). Pumice ﬂows, the most common, leave behind deposits called ignimbrites. Nu´ees ardentes are less common and include a wide variety of solid materials, dominated by pulverized rocks. Scoria (Greek = refuse) may also be found in deposits from such ﬂows. Air-borne pyroclastics (tephra) Air-borne pyroclastic materials are the most widespread volcanic disturbances. Upon falling to earth, they are called tephra (Greek = ashes). Tephra deposits are size-sorted and, at any distance from a volcanic source, are homogeneous. Size categories vary, but ﬁne material is called ash, intermediate stones are termed lapilli, and large blocks are called breccia, blocks or bombs. Larger tephra fragments are composed of scoria or pumice. Scoria is coarse, fragmented, and dense. It forms slopes near the angle of repose. Pumice is a low density, frothy rock that ﬂoats in water. Lahars and debris ﬂows Lahars (Indonesian = mud-ﬂow) are cool, unsorted, slurries of rocks, mud, and vegetation with >50% water when emplaced. Debris ﬂows are less ﬂuid (<50% water) and usually more heterogeneous. Both can be

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Fig. 5.1. Steam and sulfur vent in crater of Mount Takumbe, Japan. Adjacent rocks lack vegetation.

produced by volcanoes (cf. Gecy and Wilson, 1990). Normally, they result from rapidly melting ice and snow during an eruption. Lahars and debris ﬂows can include heterogeneous material from many elevations and ages, so it is difﬁcult to generalize about their composition. Avalanches, which are primarily solid, are even more heterogeneous.

DISTURBANCE REGIMES

Disturbance regimes profoundly affect vegetation and have been discussed thoroughly by Pickett et al. (1987), Walker and Chapin (1987) and Chapin et al. (1994). Volcanic impacts normally are either so intense that ecosystem development starts on an abiotic substrate, or so infrequent that the term “regime” is inappropriate. Volcanoes are not associated with any vegetation type (as is ﬁre), climate (as is drought), biotic interaction (herbivory), or chronic event (hurricanes). Active volcanoes occur from the arctic to Antarctica. Together, geographic diversity, variations in magnitude, and the variety of responses blur global patterns. Individual eruptions, however, vary in scale, frequency, magnitude, and severity, and volcanic landscapes display community-level patterns related to these variations.

Below we review how some aspects of disturbance regimes affect ecosystem recovery and structure. Scale Volcanic events vary widely in scale. Lava ﬂows are usually quite localized, usually ﬂowing less than 10 km (Wadge, 1983). Pyroclastic ﬂows cover hundreds of square kilometers, tephra covers thousands of square kilometers, and lahars may descend for many kilometers from the volcano. Scale affects the type of colonists and the rate of recovery indirectly. Recovering vegetation is dramatically different in isolated sites from sites near intact vegetation. Isolation led to the dominance of wind-dispersed species after the eruption of Mount St. Helens (Washington, U.S.A.) in 1980 (del Moral and Bliss, 1993). Scale effects are attenuated in secondary succession. For example, tephra impacts are widespread, but shallow deposits do not eliminate all vegetation. Destroyed habitats recover more quickly if the impact is small, whereas areas with large-scale damage can recover quickly if there are survivors. For example, on Mount Rainier (Washington, U.S.A.), the recovery rate of the Kautz Creek lahar, a long, narrow deposit surrounded by intact vegetation, was much more rapid than that of the broad Muddy River lahar (Frenzen et al., 1988).

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Frequency Volcanoes normally have long intervals between major eruptions, though some, like Sakurajima in Japan (Tagawa, 1964) and Tolbachik in Kamchatka (Russia) (Grishin, 1992), erupt frequently. Return intervals of 200–300 years reduced species diversity and endemism on Mount St. Helens, which lacks endemics (del Moral and Wood, 1988a,b). Mount Vesuvius in Italy has 18 endemics in a ﬂora of over 300 species (Mazzoleni and Ricciardi, 1993), but has suffered only localized impacts since AD 79. Magnitude and severity The magnitude of a volcanic event can be described by volume or intensity, while severity describes its biotic impact. Lavas are of such magnitude and pyroclastic ﬂows of such severity that survival is rare. Lahars and debris ﬂows rarely permit survival, but along their margins and in isolated pockets, by chance, some organisms survive. Distance attenuates the magnitude of pyroclastic ﬂow events and tephra deposits become thinner, so that there is differential survival of the biota. Lahar deposits are heterogeneous. Experience on Mount St. Helens (del Moral, 1983; Franklin et al., 1985) suggests that rhizomatous species are more likely to survive. However, a variety of species survive on root wads and soil blocks, so survival on lahars is capricious. Recovery rate should be proportional to the ratio of topsoil to total volume, which serves as an indicator of the potential for residual species and debris to end up near the deposit surface. Whereas blast effects favor plants such as hemicryptophytes and geophytes growing at or below ground level, tephra impacts can be particularly harmful to mat-formers, mosses, lichens, low herbs, and species lacking rhizomatous growth or incapable of growing through thick deposits (Antos and Zobel, 1985a; Zobel and Antos, 1991, 1992; Grishin et al., 1996). Lateral eruptions in which directed pyroclastic blasts are followed by a deep tephra rain devastate the biota. Plant responses to a given impact depend on their size and growth form. At a sufﬁcient distance from the cone, directed blasts, and pyroclastic events will most severely impact taller vegetation. Ground layer and juveniles may be protected by topography, while buried structures are protected by soil. Dormant plants will be more likely to survive, and a snow pack will

Roger DEL MORAL and Sergei Yu. GRISHIN

enhance survival. Species with seed banks are more likely to dominate recovering vegetation. Stress Each volcanic habitat imparts different degrees of stress. Lavas break down slowly and present huge challenges to colonization, so that microsites that collect moisture, trap debris and offer refuge are crucial (Lohse et al., 1995). In the tropics, ferns often form a major group of colonists (Tagawa et al., 1985), whereas ﬂowering plants initially colonize dry tropical habitats. Lichens colonize exposed surfaces over a wide range of temperatures. Early succession in cold climates is dominated by cryptogams, especially nitrogen-ﬁxing Cyanobacteria (Griggs, 1933). Mosses dominate wetter climates, lichens drier ones. Vascular plants may pioneer in cracks, or invade following physical amelioration. Pumice and scoria are drought-prone and nutrientpoor. They weather to provide better growth conditions, but they dry quickly and may be unstable. Steep slopes remain bare where drought and chronic erosion prevail (Day and Wright, 1989). Figure 5.2 shows scoria habitats at least 2000 years since their formation in Idaho. Very cold, windy winters and summer drought combine to prevent rapid succession on these substrates (Day and Wright, 1989). Tephra is the most widespread volcanic deposit, and under ideal conditions can be colonized quickly. Deep tephra is colonized at rates related to the local climate. Wet habitats on Katmai, Alaska, required three years for colonization (Griggs, 1919), while moist uplands were colonized more slowly by liverworts, then mosses and ﬁnally willows, grasses, and shrubs (Griggs, 1933). In contrast, dry tephra on Ksudach in Kamchatka, ﬂoristically comparable to Katmai, was colonized by lichens, which were still dominant after 90 years (Grishin et al., 1996). On any substrate, the overall climate may ameliorate or accentuate stress. In temperate and tropical regions, available moisture is most likely to control the rate of succession. Along an elevation gradient of the same substrate, for example lava or lahars, the succession rate normally increases with increasing moisture, until decreasing temperatures curtail the growing season (Chevennement, 1990; Tagawa et al., 1994). Aplet and Vitousek (1994) demonstrated that the rate of biomass accumulation on Hawaiian lavas was correlated with precipitation. In cold regions and at high elevations, the

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Fig. 5.2. 2000-year-old scoria, Craters of the Moon, Idaho: Chrysothamnus nauseosus in gully and scattered Agropyron spicatum.

rate of succession is more likely to be correlated with temperature. Patchiness Habitats may be heterogeneous on either a local or a landscape scale. New deposits rarely result in a patchy landscape, though they may skip over or partially damage vegetation. Differential protection by snow, topographic irregularities, rock walls, and deep canyons may all shelter the biota from tephra deposits, lava ﬂows, lateral blasts, and pyroclastic ﬂows. Even if surviving species cannot colonize surrounding barren landscapes, birds that introduce suitable plants may be attracted to such residual sites. Tephra can extinguish habitat heterogeneity by ﬁlling in cracks to form smooth surfaces. Subsequent erosion can create new opportunities for colonists (Collins et al., 1983; Kadomura et al., 1983; Suwa and Okuda, 1983; Tsuyuzaki, 1994). Local heterogeneity is enhanced by differential physical amelioration, and by the establishment of early pioneers that facilitate subsequent establishment by other species (cf. Blundon et al., 1993). Inhibition of colonists by the initial invaders also can increase the variability of habitat conditions (Clarkson and Clarkson, 1983; Wood and

Morris, 1990). As a few dominant species create homogeneous canopy cover, site patchiness is signiﬁcantly reduced, leading to lower diversity and to successional changes (del Moral, pers. observ. on Mount St. Helens, Fuji-san, and Ksudach volcano).

RATES OF SUCCESSION

The rate of succession varies with the type of the impact, climate, and geographic factors. First we will discuss the effects of substrates, then the effects of climate on a given substrate. Substrate effects Table 5.2 lists representative papers that address primary succession on volcanoes. Lavas Succession on lavas is usually slow, but the type of colonizers vary greatly. Typically there are two juxtaposed successions. On ﬂat surfaces lichens (Eggler, 1941; Cooper and Rudolph, 1953) establish and slowly form soil. In cracks, dust and organic matter accumulate and moisture collects. In warm, moist climates, vascular plants can establish more quickly. Tagawa

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Table 5.2 Summary of major papers dealing with common substrates on different volcanoes Volcano or area

Impact type

References

Gal´apagos Islands

lava

Hendrix (1981)

Hawaii

lava

Atkinson (1970); Eggler (1971); Drake and Mueller-Dombois (1993); Aplet and Vitousek (1994); Karpa and Vitousek (1994); Kitayama et al. (1995)

Kekla, Iceland

lava

Bj¨arnason (1991)

Krakatau, Indonesia

lava

Tagawa et al. (1985); Whittaker et al. (1989); Tagawa (1992); Partomihardjo et al. (1992)

La R´eunion, Mascarene Islands

lava

Chevennement (1990)

Papua New Guinea

lava

Taylor (1957)

Sakurajima, Japan

lava

Tagawa (1964, 1965)

Mount Tarawera, New Zealand

lava, scoria

Clarkson and Clarkson (1983); Timmins (1983)

Mount Tolbachik, Russian Far East

lava, tephra

Grishin (1992, 1994)

Mount Kula, Turkey

lava

Oner and Oﬂas (1977)

Vesuvius, Italy

lava

Mazzoleni and Ricciardi (1993)

Mount St. Helens, U.S.A.

lateral blast

Franklin et al. (1985); Halpern et al. (1990)

Mount Lassen, U.S.A.

lahar

Heath (1967)

Mount Ontake, Japan

lahar

Nakashizuka et al. (1993)

Mount Rainier, U.S.A.

lahar

Frenzen et al. (1988)

Mount St. Helens, U.S.A.

lahar

Halpern and Harmon (1983); Dale (1989, 1991)

Mount Taranaki, New Zealand

lahar

Clarkson (1990)

Fuji-san, Japan

pumice, scoria

Ohsawa (1984); Masuzawa (1985); Nakamura (1985)

Mount St. Helens, U.S.A.

pumice, pyroclastic ﬂows, lahars

Wood and del Moral (1987); del Moral (1993); del Moral and Bliss (1993); del Moral et al. (1995)

El Chich´on, Mexico

pyroclastic ﬂows, tephra

Burnham (1994)

Montagne Pel´ee, Martinique

pyroclastic ﬂows, lahar

Beard (1976); Sastre and Fiard (1986)

Motmot, Papua New Guinea

scoria

Ball and Glucksman (1975)

Crater Lake, U.S.A.

scoria, lapilli

Jackson and Faller (1973); Horn (1968)

El Par´ıcutin, Mexico

scoria, lapilli, ash

Eggler (1948, 1963); Rejm´anek et al. (1982)

Surtsey, Iceland

scoria

Fridriksson and Magnusson (1992)

Mount St. Helens, U.S.A.

ash

Antos and Zobel (1985a,c, 1986); del Moral (1983, 1993); del Moral and Bliss (1993); Zobel and Antos (1986)

Kyushu, Japan

ash

Tagawa et al. (1994)

(1964, 1965) described juxtaposed successions on lavas on Sakurajima, Kyushu, Japan, dating from 1476, 1779, 1914 and 1946. Seed plants rapidly colonized crevices and dominated the 1914 ﬂow, and cryptogams dominated the smooth surfaces. Sakurajima continues to eject tephra that covers the lava, the climate is warm and rainfall well-distributed, so soil formation is relatively rapid and succession relatively quick. Figure 5.3 shows young Alnus ﬁrma, Ficus erecta,

and Pinus thunbergii invading the 1914 ﬂow of Sakurajima. Local biological effects modify colonization rates. The density of safe sites (microsites favorable for seedling establishment) is crucial for establishment; but safe sites are subject to competition. For example, lava domes on Mt. Tarawera in New Zealand (which erupted in 1886) were ﬁrst colonized in a limited number of cracks by mosses and the mat-forming

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY

143

Fig. 5.3. The 1914 lava ﬂow of Sakurajima, Japan: a wide variety of woody species, e.g., Alnus ﬁrma, Ficus erecta, Pinus thunbergii, Rhus ambigua, and Rosa polyantha, with Miscanthus sinensis and Polygonum japonicum.

Muehlenbeckia axillaris, thus limiting invasion by other species (Clarkson and Clarkson, 1983). In contrast, nitrogen ﬁxers can facilitate colonization and accelerate biomass production, thus promoting succession (Hirose and Tateno, 1984; Clarkson, 1990; Halvorson et al., 1991). Microsite availability and dispersal distance are two major factors that affect the rate of succession on lava. Pyroclastic ﬂows There are few studies of succession on pure pyroclastic deposits. On Mount St. Helens, erosion of surges and nu´ees ardentes created heterogeneity and resulted in recovery rates similar to those on pumice. Vascular plants dispersed by wind dominated the early colonists. In contrast, Beard (1976) found that early colonists of nu´ees ardentes on Montagne Pel´ee in Martinique, a warmer, wetter site, were dominated by mosses and lichens, but that vascular plants had returned within 60 years. Pitcairnia sulphurea, a rock-outcrop pioneer species, dominated these communities, while shrubs occurred only in protected locations. Seed plants with succulent fruits dispersed by birds and bats had also returned within 60 years (Sastre and Fiard, 1986). Pyroclastic falls Air-fall deposits are usually size-sorted. Lapilli fall

near the source. Ash can travel hundreds of kilometers. Ash and sulfuric acid may be injected high into the atmosphere and thus affect climate and ecosystems hundreds of kilometers away (Inbar et al., 1995). Coarse tephra is inhospitable because it has little water-holding capacity, is extremely poor in nutrients, and is unstable. Coarse basaltic tephra formed Surtsey, Iceland, (erupted in 1963: Fridriksson and Magnusson, 1992). In addition to the hostile substrate, Surtsey is so isolated that few disseminules arrive. The limited ﬂora is dominated by bird- and water-dispersed species. Isolation accentuates stress and affects succession on most volcanoes, whether or not they occur on islands. Climatic stress further affects succession on airfall deposits. At timberlines with scoria, succession proceeds slowly. On Fuji-san, Japan, the normal timberline is at 2600 m. Figure 5.4 shows an advancing timberline at 2500 m, below the Hoei crater, which erupted in 1707. The forest slowly continues its advance, limited by low nitrogen levels, lack of organic matter, instability, lack of germination sites, deep snow, and difﬁculties in dispersal (Masuzawa, 1985; Maruta, 1994). Where disturbance has been frequent, timberline successions may lack regionally common species that have limited dispersal (Ohsawa, 1984; Kruckeberg, 1987; del Moral and Wood, 1988a).

144

Roger DEL MORAL and Sergei Yu. GRISHIN

Fig. 5.4. Advancing timberline on scoria, 2500 m a.s.l. on Fuji-san, Japan. Trees in distance are Larix sp., common herbs include Calamagrostis sp. and Polygonum weyrichii.

Climatic stress also affects the succession rate on pumice deserts dominated by lapilli-sized rocks, which are nutrient-stressed. One such desert near Crater Lake is over 6500 years old (Horn, 1968). It has half the species richness of the Abraham Plain pumice desert on Mount St. Helens, which is 17 years old, but has comparable cover (del Moral and Wood, 1993a). Less intense drought, thinner pumice, and less isolation contribute to the rapid recovery on Abraham Plain. Succession on scoria cones is slower than on ﬁnetextured pyroclastic materials, but can be accelerated by erosion. Eggler (1948) noted that erosion of the cone of El Par´ıcutin in Mexico (which erupted in 1943) facilitated seedling establishment by creating safe sites and by removing tephra (cf. Riviere, 1982). However, for the buried seed bank to emerge, nearly complete erosion was required. Eggler (1959, 1963) later found that removal of tephra by water and wind had permitted recovery to be dominated by surviving seeds and plants. By 1977, there were 39 vascular plant species and 20% vegetation cover, mostly due to residual species (Rejm´anek et al., 1982). Among classic studies of succession on ash is that of Griggs (1933), who followed recovery from the eruption of Katmai (Alaska) in 1912. Previous papers by Griggs described hot lahars (1918a), recovery of

buried vegetation (1919), and recovery on thinner deposits (1918b). He demonstrated that deep deposits precluded emergence of survivors, so that only longdistance dispersal could initiate succession. The cool, wet climate permitted pioneering leafy liverworts to form turf, and these turfs facilitated the colonization of mosses, willows, grasses, and forbs. The season of deposition contributes to determining the degree of damage and survival (Antos and Zobel, 1982). Where snow protected vegetation during the eruption of Mount St. Helens in May, 1980, damage to the understory from ash-fall was signiﬁcantly less and recovery signiﬁcantly more rapid than where vegetation was exposed (Halpern et al., 1990). Tephra buries and can kill components of existing vegetation and this initiates succession. Deep burial precludes survival, and primary succession ensues. However, tephra can increase succession rates on lava by providing a rooting material (Tagawa et al., 1994). Thin deposits also may have ephemeral mulching effects that enhance biomass production for some species (Harris et al., 1987; Chapin and Bliss, 1988). Between these extremes, ecosystem recovery patterns are complex. Recovery is affected by differential survival, erosion and concentration of tephra deposits,

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY

145

Fig. 5.5. Toutle River lahar, Mount St. Helens, U.S.A., with fringes of rapidly developing Alnus rubra and Salix spp., shown after 16 years.

proximity to potential colonists, and many other factors. Lahars and debris ﬂows Lahars can wreak havoc over huge areas (Mizuno and Kimura, 1996). Stabilized, narrow lahars and debris ﬂows can be colonized rapidly from adjacent vegetation (Dale, 1989). Figure 5.5 shows the lower Toutle River lahar, Mount St. Helens, after 16 years. The stabilized margins are densely vegetated by Salix spp. and Alnus rubra, while erosion continues to scour the banks. The mixture of soil types and rocks provides heterogeneity and potential safe sites. However, large, high-elevation lahars, such as that near Mt. Taranaki, New Zealand (Clarkson, 1990) and on Mount St. Helens (del Moral, 1993), develop slowly due to isolation and climatic stress. Trees often colonize lahars, provided seed sources are adjacent (Heath, 1967; Frenzen et al., 1988). Figure 5.6 shows invasion 28 years after the Kautz Creek lahar occurred. However, on the Muddy River lahar, Mount St. Helens, Halpern and Harmon (1983) found that species richness clearly decreased away from the forest edge. Local diversity was enhanced by stumps, soil clumps, and root wads providing seeds, organic matter, and safe sites. By 1995, dense Alnus rubra stands had developed at an altitude of 800 m a.s.l., where conditions were mild and dispersal

did not limit invasion. At 1000 m a.s.l., the vegetation remained open and diverse. Unexpected species from higher elevations dominated (del Moral, pers. observ.). Nutrient limitations inhibited conifer growth, but Pseudotsuga menziesii had become more common. Mosses such as Racomitrium canescens did not occur until several years after signiﬁcant colonization by vascular plants. They are pioneer species only in that they establish on barren microsites. The importance of residual soil or plant fragments has been observed repeatedly on lahars. Nakashizuka et al. (1993) found that the debris avalanche from the Ontake volcano in Honshu, Japan, was colonized by wind-dispersed species, but that residual soil and root fragments accelerated succession. The early dominance of Lupinus spp. on Mount St. Helens lahars resulted in part from surviving seeds and root fragments (del Moral, pers. observ.). Tropical lahars and debris ﬂows can recover quickly, but often are colonized by only a few species (Eggler, 1959; Whittaker et al., 1989). Burnham (1994) described hot debris ﬂows from El Chich´on in Mexico (1982). Vascular plant diversity was low because vines inhibited colonization by other vascular plants. Biomass on lahars in temperate climates develops slowly, but diversity recovers quickly (Frenzen et al., 1988; Dale, 1991). This anomaly may occur because on temperate or higher-elevation lahars, where physiolog-

146

Roger DEL MORAL and Sergei Yu. GRISHIN

Fig. 5.6. Kautz Creek, Mount Rainier, U.S.A., lahar of 1947, photo taken in 1975, showing dead conifer trunks that accelerated succession.

ical stresses often limit productivity, many species still can establish in the relative absence of competition. Comparisons of rates of succession Succession on lava is always slow, but its rate can be affected by surface heterogeneity, subsequent tephra deposits, distance from the volcanic cone, and climatic stress. Coarse air-fall deposits permit more rapid development, but the rate of succession may be slowed by site instability and low nutrient status. Succession on ash and ignimbrites may develop rapidly if erosion ameliorates initial conditions. Lahars are composed of older, more mature substrates, so that dispersal limitations strongly inﬂuence development. The degree of environmental stress can modify the rate of succession on each substrate. Environmental stress Factors that facilitate biomass accumulation will accelerate succession. This rule is apparent on volcanoes since they occur under most combinations of environmental stress. Recovery on two identical volcanic substrates varies with temperature and moisture conditions. Recovery is rapid in the tropics and slow at high latitudes and high elevations. Succession is more rapid

in moist climates than in semi-arid ones at the same latitude. Substrate microtopography alters local succession rates by affecting environmental stress. Grishin (1992, 1994) described succession on lavas of two Kamchatka volcanoes. Primary succession and soil development are slow on a’a lava in this cool, temperate forest region. Initially, herbs such as Chamaenerion angustifolium, Lerchenfeldia ﬂexuosa, Leymus interior, and Poa platyantha colonized fractures, while lichens (Stereocaulon spp.) and mosses (Polytrichum juniperinum) dominated exposed surfaces. Woody species (e.g., Betula ermanii, Populus suaveolens, and Salix caprea) were rare during the ﬁrst 50 years, and remained stunted on 500-year-old surfaces. On 1000year-old lavas, soil is shallow, supporting dense forests of stunted Pinus pumila. Gradually, forest communities of Alnus kamtschatica, Betula, and Larix cajanderi develop, and after 2500 years the vegetation begins to resemble mature surrounding communities. This succession takes at least four times as long as succession described in warm-temperate southern Japan (Tagawa, 1964, 1965). In any given region, on any substrate, moisture availability is crucial. Tagawa et al. (1994) showed that succession on lavas on subtropical islands was more advanced in the cloud layer at 800 m a.s.l. than in

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY

the drier zone at 500 m a.s.l. Fern´andez-Palacios and de Nicol´as (1995) found that zonation on Tenerife in the Canary Islands was related to both temperature and moisture along an altitudinal gradient. On Mount St. Helens (del Moral et al., 1995), stable streams and small oases found on pumice developed much more rapidly than in the surrounding uplands. Aplet and Vitousek (1994) found that on Mauna Loa, Hawaii, biomass decreased with decreasing precipitation and temperature. In contrast, Vel´azquez (1994) determined that temperature was more important than moisture gradients in controlling succession in central Mexican volcanoes; however, his study covered only a small moisture range, but a large elevational range.

TYPICAL VEGETATION RESPONSES TO VOLCANIC DISTURBANCE

Habitats resulting from volcanic eruptions are as varied as any on earth. The nature of colonists, rates of succession, degree of predictability, and similarity to adjacent vegetation all vary widely. Below we discuss some vegetation patterns common to many volcanoes. Landscape effects Timberline depression Brown (1994) predicted that any disturbance lowers timberlines from levels predicted from the geomorphic and climatic factors. Timberline depression on volcanoes is widespread [Beamon (1962), near Mexico City; Fosberg (1959), on Mauna Loa, Hawaii]. Clarkson (1990) determined that upper limits of Weinmannia racemosa forests on Mt. Taranaki, New Zealand, were depressed and that species diversity was reduced by chronic tephra deposits. Veblen et al. (1977) reported that catastrophic volcanism depressed the Nothofagus betuloides timberline in south-central Chile by 100 to 300 m. Ohsawa (1984) and Masuzawa (1985) both noted that the timberline on Fuji-san, Japan, was depressed; both indicated that the upward colonization recapitulates succession. These studies also showed that alpine climax species can act as pioneers on lower sites. Reasons for the slow upward forest migration generally include nutrient limitations, poor soil development, substrate instability (Jackson and Faller, 1973), poor upward seed dispersal, wind, snow effects, and inability of tree seedlings to establish in most years.

147

Reduced diversity and turnover Initial turnover rates on volcanoes appear to be lower than might be expected, and accumulation of species is usually gradual. Their ﬂoras are rarely at equilibrium. The best evidence that species richness is reduced on volcanoes comes from studies showing continual increases in species richness through time. The accumulation of species and, where data are available, the species turnover are discussed below. Table 5.3 summarizes representative studies of richness during early succession. Tagawa (1964, 1965) listed vascular plants found on lavas of four distinct ages. The youngest was dominated by lichens and mosses, and ﬁve vascular species. The 47-year-old ﬂow had 54 vascular species, dominated by herbs, with high cover of non-vascular plants. The 475-year-old ﬂow supported a late seral forest with 76 species. Whittaker et al. (1992) summarized changes in the ﬂora of Rakata, Indonesia, during 100 years following the eruption in 1883. By 1989, there were 397 species, with richness continuing to increase. The warm tropical habitat, habitat complexity, proximity to sources of propagules and Rakata’s escape from subsequent eruptions has permitted the ﬂora to develop rapidly and richness continues to increase. Anak Krakatau emerged from the sea in 1930. Bush et al. (1992) recorded 138 species in 1989–91, though a combined total of 157 species have been noted. This relatively low turnover rate (19 species once found, not present in latest survey) occurs despite subsequent disturbances that set succession back. Each restart has led to different colonization patterns. Turnover here is directly related to volcanic events, not to extinction. In 1987, Fridriksson summarized his work on Surtsey, formed in 1963. Initially, there were two species. By 1986, 24 vascular plant species had been recorded (Table 5.3), but most of them were rarely encountered there. By 1990, 28 species had been recorded, of which 24 occurred in that year (Fridriksson and Magnusson, 1992). Some species had several colonization and extinction events. Surtsey combines elements that ensure high rates of turnover: it is small, climatically harsh, isolated by sea water, unstable, and composed of substrates very low in nutrients and organic matter. Therefore, colonization rates and population densities are low, and stochastic variation is high. Mount Tarawera, in New Zealand, erupted massively in 1886 (Clarkson and Clarkson, 1983; Clarkson,

148

Roger DEL MORAL and Sergei Yu. GRISHIN

Table 5.3 Number of vascular plant species during early development, middle years, and the last measurement in a particular study Volcano

Eruption Type of date surface

Age (yr)

References

Number of species Young surface

Intermediate surface

Oldest surface

Sakurajima, Japan

1476

lava

475

5

54

76

Rakata, Krakatau, Indonesia

1883

mixed

106

24

253

397

Whittaker et al. (1992)

Anak Krakatau, Indonesia

1930

scoria

60

17–20

45

138

Bush et al. (1992)

Mount Tarawera, New Zealand

1886

lava

93

nd a

63

74

Clarkson (1990)

El Par´ıcutin, Mexico

1943

pumice

34

2

17

39

Rejm´anek et al. (1982)

Mount Rainier, U.S.A.

1947

lahar

33

17

21

32

Frenzen et al. (1988)

Surtsey, Iceland

1963

lava/tephra

27

2

15

24

Fridriksson and Magnusson (1992)

pumice

16

nd a

21

44

del Moral and Wood (1993a,b)

blasted

16

3

16

27

del Moral, unpubl.

lahar

16

5

31

46

del Moral and Wood (1988b)

Mount St. Helens, U.S.A.

a

1980

Tagawa (1964)

No data.

1990). Between 1964 and 1979, species richness increased from 63 to 74, representing 28 colonization events and 17 local extinctions. The increase in richness was modest, but there was a substantial increase in cover, primarily by shrubs that expanded as resources accrued. This succession is in a phase of slow species accumulation with modest turnover and gradual expansion of plant cover. The Kautz Creek lahar described by Frenzen et al. (1988) on Mt. Rainier (Washington, U.S.A.) demonstrated a gradual accumulation of species as well as spread of vegetative cover. Species that had previously been encountered, but were absent in 1980, were weeds that had been excluded by developing forest cover. Here, turnover is due neither to disturbance nor to biogeographic effects, but to competition among plants, probably for light as well as soil resources. The record of vegetation recovery for Mount St. Helens includes time-series studies in several habitats. The lahar sample is near intact vegetation, so that rapid recovery was expected. Colonists were ﬁrst encountered in 1982, and by 1994 richness was comparable to that of the adjoining meadow, but cover was much lower (del Moral and Bliss, 1993). Studebaker Ridge was sampled by ten permanent plots starting in 1984. It is within 1 km of potential colonists. Total richness and cover were very low in 1995. There had been no turnover, only accumulation of species. Sampling on the isolated Abraham Plain started in 1988

on a barren grid (del Moral and Wood, 1993a). There were then 23 species, but total cover was less than 1%. By 1996, only one species had disappeared, while an additional 23 species had colonized. Cover remained sparse; the mean number of species in 400 plots continues to increase annually. These results suggested that, in stressful sites, turnover is concentrated on rare species. Extinctions may become more frequent after the canopy closes and trees or shrubs become dominant. Disharmony Floristic disharmony occurs when isolation leads to a species composition distinct from the regional ﬂora. An immature or isolated ﬂora is drawn from a subset of the surrounding ﬂora, producing vegetation that is markedly distinct from the donor vegetation. Disharmony can be reﬂected in life-form spectra, dispersal types, or dominance hierarchies. The degree and nature of the disharmony depends on the type and degree of isolation from sources, and the age of the developing vegetation. Surtsey provides ample evidence of disharmony. Of the few species that have arrived, only 9% are wind-dispersed and 27% sea-borne. The remaining 64% are bird-dispersed, in sharp contrast to Iceland (Fridriksson and Magnusson, 1992), where most species are wind-dispersed and few are sea-borne. Chronic minor disturbances also differentially reduce the ﬂora (Antos and Zobel, 1984, 1985a,b; Zobel and Antos, 1986; Grishin et al., 1996),

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY

and accentuate ﬂoristic growth-form disharmony. For example, mosses, lichens, and non-rhizomatous perennials are more susceptible to tephra impact than trees or shrubs. Whittaker et al. (1992) described several interesting aspects of disharmony. On Krakatau’s remnant, Rakata, sea birds and bats distributed 30.0% of the species while wind deposited 48.4%. Sea-borne dispersal has been signiﬁcant (17.1%), while humans have introduced 4.5%. Compared to nearby islands, animaldispersed taxa are under-represented in this ﬂora and wind-dispersed taxa are over-represented. On Anak Krakatau, 42.9% of the ﬂora is sea-dispersed, 28.1% is wind-dispersed, 19.3% is ﬂight dispersed, and humans introduced 11.9% of the species. Ferns account for 25.4% of the ﬂora, and epiphytic orchids and epiphytic ferns often occur only on the ground (Partomihardjo et al., 1992). Structural disharmony was studied on Mount St. Helens lahars (del Moral and Bliss, 1993). A lahar adjacent to intact forest has been invaded by conifers, and is more similar to the forest (Percent Similarity = 59%) than the meadows (PS = 41%). A more isolated lahar is equally similar to the woodland (PS = 47%) and meadow (PS = 48%). The comparison ﬂora was already reduced as a result of frequent large-scale volcanic activity and the youth of Mount St. Helens (del Moral and Wood, 1988b), so that differences were less than they might have been. The cone of Mount St. Helens had only 72 native species in 1987, compared to at least 95 before the eruption. Three nearby volcanoes had from 185 to 276 subalpine and alpine species. Kruckeberg (1987) listed 70 expected species missing from Mount St. Helens after the eruption of 1980. Most of the missing species disperse poorly. Chance and contingency Isolation and differential dispersal permit singular events or historical accidents (contingent factors) to affect community development, even as abiotic processes lead predictably to habitat amelioration. Novel assemblages may develop in which the relationship between vegetation and environmental factors is weak (McCune and Allen, 1985). Bush et al. (1992) stated that successional pathways may be extrapolated somewhat, but that too many stochastic biotic forces act for long-term predictions on Krakatau to be valid. Each island of the group was developing differently as a result of unique colonization events and subsequent disturbance. Whittaker et al.

149

(1992) noted that the lack of larger mammals conﬁned many plant species to the coast. Such hierarchical links, highly subject to stochastic effects between plants, animals, and fungi are important determinants of succession on these islands. Tagawa (1992) emphasized that stochastic dispersal events determining the initial colonizers force succession for centuries and create novel communities (Tagawa et al., 1994). In southwestern Japan, Ardisia forms a stable pioneer scrub community, which resists invasion indeﬁnitely. One group of stochastic elements that affect succession is unusual climatic events. Lohse et al. (1995) suggested that drought could alter primary succession on pahoehoe lavas on Hawaii. Stochastic effects weaken environmental relationships. Del Moral et al. (1995) applied canonical correspondence analysis (ter Braak, 1987) to communities on Mount St. Helens. There was little correlation between species distributions on primary successional sites. Only 14% of the variance was explained by the analysis, and only geographical factors were signiﬁcant. It appears that, in the early stages of succession, chance plays the dominant role in determining which species occur where, and at what abundance. Differential effects An interesting aspect of tephra is its differential impact on vegetation. The gradual nature of tephra deposition over space blurs the distinction between primary and secondary succession. Grishin et al. (1996) described recovery from tephra from the Ksudach volcano, Kamchatka (erupted in 1907). They mapped three vegetation zones differing in the depth of tephra. In deposits deeper than 100 cm, all vegetation was killed, and trees were crushed; after 90 years, primary succession was only in a second stage, dominated by lichen mats, with only scattered herbs. Deposits between 30 and 100 cm deep deﬁned a complex zone. Deposits over 70 cm destroyed all vegetation, but left standing dead trunks, and the rate of primary succession that dominated recovery was variable. Lichen stages were followed by a dwarf-shrub–herb mosaic and a secondary birch-forest stage. Isolated trees survived in deposits less than 70 cm and primary and secondary stages formed a mosaic. Deposits less than 30 cm permitted trees to survive, while deposits of 10 to 20 cm eliminated mosses and lichens but only damaged dwarf shrubs and herbs. Deposits under 10 cm damaged herb, moss, and lichen layers but did

150

Roger DEL MORAL and Sergei Yu. GRISHIN

Fig. 5.7. Remnant of old Betula ermanii stump on volcano Ksudach, Kamchatka Peninsula, Russia, invaded by Epilobium (Chamaenerion) angustifolium and Rubus sachalinensis.

not eliminate any species, so recovery, not succession, has been the predominant process. The succession rate was related to survival. Without survivors, the primary succession rate related to deposit depth, distance to survivors, and presence of standing dead trees. The latter provided perches for birds which imported ﬂeshy fruits (McDonnell and Stiles, 1983). The dead trees subsequently dropped leaves that enriched the pumice beneath (Fig. 5.7). These oases have been colonized by a haphazard assortment of bird-dispersed species intermingled among common wind-dispersed species. Effects on animals While volcanoes occasionally kill many humans (e.g., 20 000 in the lahar spawned by Nevado del Ruiz, Colombia), animal populations are usually devastated. Tephra fallout decimates insect populations, since ash abrasion causes swift desiccation (Edwards, 1986). Ant colonies survived on Mount St. Helens (del Moral, 1981), but survivors lacked resources and most colonies failed. Wildlife losses during the Mount St. Helens eruption included 5000 deer (Odocoileus hemionus), 1500 elk (Cervus canadensis), 200 black bear (Ursus americanus), and 15 mountain goats (Oreamnos americanus), in addition to uncounted

rodents, birds, and ﬁsh. Fifty-seven humans were killed. Small animals have recovered from pools of survivors (e.g., gophers – Thomomys sp.) and by invasion (e.g., mice). Wildlife can recover quickly after volcanic events. Many species thrive in post-eruption ecosystems when forage recovers quickly. Most animal species have at least moderate dispersal powers, so populations of insects, birds, and mammals follow vegetation recovery. Fish recovery may be much slower in new or sterilized systems. For example, Spirit Lake adjacent to Mount St. Helens was boiled and displaced. Only in 1994 was the ﬁrst ﬁsh found, and populations remain very low.

RECOVERY MECHANISMS

Recovery from devastation on volcanic landscapes has several common features. The importance of these features varies, and is conditioned by several factors. An understanding of recovery mechanisms is of both practical and theoretical importance. On the practical side, one can learn how to design more efﬁcacious restoration of devastated landscapes, and how to alleviate potential volcanic destruction

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Fig. 5.8. Rill edge on Mount St. Helens, U.S.A.: invading Anaphalis margaritacea, Hieracium albiﬂorum, and other species, conﬁned to edges where seeds are trapped, protection is offered against the wind, and water erosion is minimal.

better. Aspects of succession theory will be improved if abiotic amelioration, dispersal, survivors, biotic interactions, and factors affecting predictability are better understood.

Abiotic amelioration Invasion of new volcanic landscapes cannot occur without physical amelioration (del Moral and Bliss, 1993). Chemical weathering is one crucial ameliorating factor (Ugolini et al., 1992), but physical weathering (del Moral, 1983, 1993), organic fallout (Edwards and Sugg, 1993), and erosion (Franklin et al., 1985; Tsuyuzaki and Titus, 1996) all play important roles. These abiotic processes create safe sites suitable for seedling establishment (Fridriksson, 1987; Tsuyuzaki and del Moral, 1994). Figure 5.8 shows a rill formed by minor erosion, with invaders concentrated on the rill edge. Physical processes form microsites that are extremely important for nucleation in early succession (Yarrington and Morrison, 1974). All studies of lavas demonstrate that vascular plant establishment early in succession occurs in cracks. Surface manipulations on

Mount St. Helens demonstrated the crucial role of physical amelioration (del Moral and Wood, 1988b, 1993b; Titus, 1995). Safe sites have often been found to be the sites of ﬁrst establishment (del Moral and Bliss, 1993), but this pattern becomes more diffuse as mature plants survive and expand on gradually ameliorating surfaces. As a rule, seedling establishment and survival on barren volcanic substrates is enhanced by any factor that promotes seed capture, reduces drought stress, concentrates organic matter, or increases nutrient uptake ability. The physical destruction of a site through erosion can promote succession. Tephra buries vegetation, but if erosion soon occurs, most plants survive (del Moral, 1981; Riviere, 1982; Kadomura et al., 1983). On Mt. Usu, Japan (Tsuyuzaki, 1987, 1989, 1991, 1995; Tsuyuzaki and del Moral, 1995), the rate of recovery was greater in gullies from which tephra had been eroded than in adjacent sites. The seed bank and surviving buried organs contributed greatly to initial recovery. Tephra-covered sites recovered more slowly, and colonizers were dominated by wind-dispersed taxa such as Betula maximowcziana, Salix hultenii, and Asteraceae such as Anaphalis margaritacea, Petasites japonicus, and Picris hieracioides.

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Dispersal Dispersal barriers clearly affect the recovery of primary volcanic surfaces. Ocean barriers lead to high proportions of water-dispersed species, and wind-dispersed species are usually well represented. Oceanic volcanoes may also have more species dispersed by birds than by ants and mammals. In contrast, continental volcanoes are likely to be colonized primarily by wind-dispersed species (Tsuyuzaki and del Moral, 1995). Where birddispersed plant species are common, frugivorous birds may ﬂy considerable distances over barren terrain to isolated dead trunks. Colonists are drawn from species that can reach a site, not from a suite of pre-adapted “pioneer” species. Wood and del Moral (1987) demonstrated that latesuccessional species survived better than pioneers on barren surfaces. They concluded that the ﬁrst colonists were “pioneer” species only where their superior dispersal abilities overcame their inferior establishment ability. Late-successional species that somehow reach isolated sites accelerate succession. The mysterious case of isolated patches formed on Mount St. Helens by the late-successional species Lupinus lepidus (Morris and Wood, 1989) demonstrated that dispersal barriers can be more important than stress in determining the identity of the ﬁrst colonists. Dispersal barriers and poor dispersal ability of stresstolerant species affect high-elevation vegetation. On Mount St. Helens, there are steep richness gradients upward along ridges. In one case, richness declined from 18 species to nine species per plot over an elevation increase of 250 m (del Moral, pers. observ.). Species missing at higher elevations have poor to moderate dispersal abilities. Invasion near the margin of a lahar or lava ﬂow is typically described as a phalanx, where vegetation advances along a gradually advancing front. Phalanx invasions are common where dispersal distances are short and colonization probability high. Smathers and Mueller-Dombois (1974) described such an invasion on a lava ﬂow from Kilauea, Hawaii, near undisturbed Metrosideros polymorpha forests. Beyond 100 m or less from the forest, the seed rain is typically extremely diffuse and initial populations are conﬁned to safe sites. Here populations expand from a few isolated colonists, a process termed nucleation (Yarrington and Morrison, 1974). Seemingly trivial events, such as where a spider spins a web (Dale, 1991), can affect establishment. A sparse vegetation with many

Roger DEL MORAL and Sergei Yu. GRISHIN

unoccupied safe sites results. Once a plant germinates, establishes, and matures, it becomes the predominant source of the local seed rain. The capricious nature of dispersal in stressful, isolated sites strongly affects subsequent succession. Survivors Residual species have survived the impact of a volcanic disturbance, while relict sites escaped the devastation. The best-known relicts are kipukas, the Hawaiian term describing vegetation surrounded by, yet spared from, lava. Such relicts retain mature vegetation, but result in few colonists of young lavas. Many colonists appear drawn from among a pool of ferns, orchids, Asteraceae, etc., dispersing over a long distance and capable of tolerating initially stressful conditions (Smathers and Mueller-Dombois, 1974), though Metrosideros does colonize from kipukas. In such cases, species adapted to mature soils and forests cannot establish. Del Moral et al. (1995) described residual species patches on the Pumice Plain of Mount St. Helens. These patches retained understory plants, but the canopy disappeared. The hypothesis that these patches would accelerate primary succession was not conﬁrmed. After 15 years, only seedlings of residual species (e.g., Rubus spectabilis) occurred outside a patch. These patches were themselves being invaded by species dominant on barrens and did not accelerate forest development. Rather, these tattered remnants provide sites where pioneers grow and reproduce vigorously to enhance the seed rain over the surrounding barrens. The cases of Hawaiian relicts and Mount St. Helens residuals imply that there is a limit to the ability of plants from late succession or climax conditions to colonize barren substrates. Franklin et al. (1985) described residual vegetation within the blast zone of Mount St. Helens. Rhizomes found in lahars, seeds lodging near the surface, and plants protected by snow were common. Succession was accelerated by these survivors, though species composition had been altered substantially. Alteration results from differential survival and from invasion by small or light-seeded taxa such as Alnus rubra, Epilobium angustifolium, and Salix spp. Biotic interactions Facilitation and inhibition Facilitation involves any biotic effect that promotes

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colonization, while inhibition refers to any process that precludes colonization (Connell and Slatyer, 1977). Inhibition typically includes both resource competition and allelopathy, the chemical inhibition of adjacent plants. Soil amelioration has been noted for taxa as different as Alnus rubra (Dale, 1989), Lupinus lepidus (Wood and del Moral, 1988), Myrica faya (Vitousek et al., 1987), Polygonum cuspidatum (Hirose and Tateno, 1984), and liverworts (Griggs, 1933). Trapping windblown soil and seeds (Chambers et al., 1991), nurseplant effects (Hirose and Tateno, 1984) and delayed facilitation (Eggler, 1963; del Moral and Wood, 1993b) are recognized as important factors in succession. Very early primary succession on volcanoes seldom yields densities sufﬁcient to inhibit invasion, but inhibition will occur where sufﬁcient cover develops. Clarkson (1990) described how Coriaria arborea, a nitrogen-ﬁxing shrub on Mt. Tarawera, formed a dense thicket that reduced species richness from 50 to about 10 species. The balance between facilitation and inhibition is delicate. Lupinus lepidus plays a dramatic role on Mount St. Helens (Halvorson et al., 1991, 1992; Halvorson and Smith, 1995). Dense Lupinus prevented invasion of wind-blown plants, whereas scattered Lupinus enhanced survival of planted seedlings (Morris and Wood, 1989). Del Moral (1993) showed that while Lupinus improved local soil and microclimatic conditions, competition excluded establishment by seedlings of other species. Once an adult clump died, seedlings of several species become common on the edge but not in the clump.

Animals Animals obviously disperse seeds and subtly contribute to soil development. Edwards et al. (1986) trapped arthropods on Mount St. Helens from 1980 to 1985 and found that while most soon perished, the organic fallout accelerated recovery of volcanic barrens (Edwards, 1988; Edwards and Sugg, 1993). The organic matter contributed by the victims, as well as by seeds and pollen, is crucial to plant establishment, improving water holding capacity and soil nutrients that promote seedling survival (Edwards, 1986). Spiders, a major group of colonists in barren sites, spin ground webs that trap seeds in and near safe sites (Dale, 1991; del Moral, pers. observ.). Burrowing species often survive to emerge into a landscape devoid of predators with an abundance of below-ground food. Gophers can break up tephra layers and accelerated the recovery of buried vegetation (Andersen and MacMahon, 1985a,b). Soil fertility was improved, as evidenced by yields signiﬁcantly higher in gopher-worked tephra on Mount St. Helens, compared to undisturbed sites (del Moral and Clampitt, 1985). Large animals may create safe sites. Griggs (1919) provided a wonderful photograph of seedlings growing in bear tracks on Kodiak, Alaska, after the Katmai eruption. Though most large mammals avoid devastated landscape until vegetative recovery has provided food, elk herds on Mount St. Helens are an exception. They have routinely traversed barren sites, importing seeds and creating safe sites where they travel through mud.

Mycorrhizae Mycorrhizae play a major role in community structure and function, but their role in primary succession on volcanoes is poorly understood. Allen (1987) and Allen and MacMahon (1988) suggested that the lack of mycorrhizae limits colonization on Mount St. Helens. However, Titus (1995) demonstrated in ﬁeld and greenhouse experiments that mycorrhizae had little effect on growth or reproduction of colonizing plants, though they were common in wellvegetated sites. Koske and Gemma (1997) studied sand dune succession involving Ammophila arenaria in Massachusetts (U.S.A.). Mycorrhizae were not found in barren sites, but only in association with planted grasses. These results suggest that mycorrhizae usually follow initial colonizers and that they then can facilitate later colonists.

Predictability In addition to intrinsic site qualities, the accumulation of species on volcanoes is determined by chains of lowprobability events, conditioned by landscape effects that are often historically unique. Differentially isolated identical habitats can develop different vegetation in a seemingly chaotic way (Kazmierczak et al., 1995). Distance alone affects the rate and direction of early succession on many volcanoes. Vegetation is initially sparse, and many empty safe sites occur. Any lucky success contributes disproportionately to recruitment and accentuates the initial colonization effect. Biotic interactions that might structure plant communities remain weak. Such results pose difﬁculties for those who argue that primary succession is deterministic and vegetation develops predictably. Therefore, reductionist

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deterministic approaches to predicting successional pathways may fail (Talling, 1951). Succession may be explained by traditional mechanisms and be arbitrarily divided into stages (e.g., pioneers, seral species, and climax species), but a particular result is not predetermined – it is merely the current lottery draw. Novel species combinations not predictable from the local ﬂora are often observed early in succession. Examples include studies on Krakatau (Tagawa et al., 1985; Whittaker et al., 1989, 1992; Partomihardjo et al., 1992; Tagawa, 1992), Sakurajima (Tagawa, 1964), Surtsey (Fridriksson and Magnusson, 1992), and Mount St. Helens (del Moral, 1993), and in New Zealand (Clarkson and Clarkson, 1983). Sale (1977) proposed a lottery model which stated that, if any of several species might occupy a site and colonization rates are low, then identical habitats will support different species (cf. Lavorel and Lebreton, 1992). Establishment patterns on Mount St. Helens agree with this model (del Moral, 1993). The suite of common colonizing species overlap broadly in their microsite distributions, yet they rarely occur in close proximity. Long-term successional patterns may be described by the “carousel” model (Hanski, 1982; Collins et al., 1993; van der Maarel and Sykes, 1993), which states that communities are dynamic at small scales, and that cyclic replacement by any of a guild of species can occur. Colonists win lotteries, and winners are “rewarded” with disproportionate contributions to nearby safe sites. Then, a further series of stochastic events determine where the next generation establishes. Biotic interactions only gradually become discernible, as competition, local dispersal, herbivory, and facilitation dampen stochastic variation and species patterns become more predictable. Despite such predictable abiotic processes as erosion and soil weathering, much stochastic variation remains and may contribute to the large unexplained variance found in most vegetation studies. As long-lived perennials begin to dominate, the carousel slows.

MANAGEMENT

Alien species Weeds are increasingly common on volcanoes. Most 1

See footnote to Table 14.5, p. 373.

Roger DEL MORAL and Sergei Yu. GRISHIN

20th Century eruptions have occurred in the new biological context provided by the “Columbian Exchange” (Crosby, 1972) – that massive redistribution of the biota that followed the European encounter with the Americas. Alien species may have ephemeral effects, but some successional paths have been altered completely by exotic species. Northern Hemisphere conifers were introduced to New Zealand, altering normal succession. Exotics such as Cirsium vulgare, Holcus lanatus, Hypochaeris radicata, Lupinus arboreus, Pinus spp., Senecio jacobaea, and Trifolium spp. now dominate many successional sites (Clarkson and Clarkson, 1983). The 49 introduced species encountered on Mt. Tarawera were more abundant than the 116 native vascular species. Smathers and Mueller-Dombois (1974) indicated that exotics on Hawaii were rare as pioneers on lava, but frequently invaded tephra sites. However, Myrica faya, a nitrogen-ﬁxing Canary Islands native, has displaced the native Hawaiian Metrosideros polymorpha on more recent tephra (Vitousek et al., 1987; Walker and Vitousek, 1991). Myrica can grow rapidly under stressful conditions and is at a competitive advantage over any Hawaiian woody species. The invasion of Myrica makes it plain how a single invasion can have a major effect on successional development, and may transform a landscape. Herbaceous seed plants, poorly represented in Hawaii, also are becoming dominant. Smathers and Mueller-Dombois (1974) asserted that “herbaceous exotics ﬁll a practically vacant niche”, and will not interfere with successional development. We are not so sanguine. Hawaiian volcanic grasslands are also affected by exotics such as Anthoxanthum odoratum (Karpa and Vitousek, 1994). MacDonald et al. (1991) showed that exotics have severely altered the remaining native vegetation of lavas in La R´eunion, and it is likely that a detailed survey of insular volcanoes would echo this pattern. On Mt. Usu (Hokkaido, Japan), exotics such as Festuca rubra, Poa compressa, and Trifolium repens were seeded for erosion control (Riviere, 1982). Dale (1991) reported that misguided attempts to stabilize the slopes of Mount St. Helens through hydroseeding1 introduced many European weeds, reduced tree density, and inhibited the development of native species on the debris avalanche. The presence of weeds in new volcanic landscapes is inevitable, and again reminds one that plant com-

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Fig. 5.9. Mount Usu, Japan: installing mixed forest species for erosion control.

munities are assembled from the available components. While plant communities result from many factors, the vegetation of any habitat is dependent on its historical and landscape context. Management of volcanic hazards Active management of volcanic landscapes is uncommon (see p. 154). Erosion control through hydroseeding and tree planting has been attempted in Japan (Fig. 5.9) and the United States. On Mount Usu and on Mount St. Helens, native tree seedlings were planted widely in sites with surviving soil. These actions appear to have had few unexpected effects, and are analogous to replanting after clear-cut logging. Forest composition will not return to some pre-eruption condition, but will be conditioned by reforestation efforts. Engineering approaches to erosion control are common in Japan. On Sakurajima (Japan), where ash-fall is chronic (several kg m−2 yr−1) , local inhabitants build small concrete shelters for protection against the ash-falls (see Fig. 5.10). Engineering approaches were also applied on Mount St. Helens. The U.S. Army Corps of Engineers built a check dam on the Toutle River, ﬂowing from Mount St. Helens, that effectively trapped sediment for a few years. The outﬂow of the new Spirit Lake was dammed

by the debris avalanche. A tunnel was drilled to lower the lake level by several meters, ensuring that towns on the lower Toutle would not suffer a catastrophic ﬂood. Many volcanoes with a history of recent eruptions are monitored to provide early warnings. Mount Rainier may be the most dangerous volcano in the United States – lahars resulting from earthquakes or an eruption easily could kill over 100 000 people and cause uncountable damage (S. Malone, pers. commun., 1996). Few have left the drainages, however, depending instead on a civil alert system that would provide up to 30 minutes warning. Lahars may revegetate quickly, if dispersal and fertility limitations are overcome. The 1991 Mount Pinatubo lahars destroyed over 30 000 ha (Mizuno and Kimura, 1996). Experimental studies showed that, with nitrogen fertilization, rice grew moderately well, and agriculture may return to this lahar in record time. Minor volcanism may be manageable, but in most cases, escape is the only rational response to imminent volcanic hazard. Chester (1993) reviewed hazard mapping from various data. He found that recent work has assessed hazard potential well, but the magnitude of impacts may be underestimated, as at Mount St. Helens. Dobran et al. (1994) modeled the AD 79 Vesuvius event (see Sigurdsson et al., 1985), and determined that pyroclastic ﬂows could destroy

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Fig. 5.10. Ash-protection huts on Sakurajima, Japan.

all life within 7 km. Hazard maps are the basis for evacuation and development planning. Many volcanoes are monitored for such tell-tales as earthquake swarms, tilts, and bulging, and increasingly this monitoring employs geographical positioning systems and satellite systems. Armed with such information, planners are better prepared to meet challenges posed by active and currently dormant volcanoes.

CONCLUSIONS

Volcanoes impact all ecosystems and represent the most intense of nature’s forces. Francis (1993) notes historic examples of eruptions with demonstrable global climatic effects (e.g., Thera, ~1500 BC; Mount Katmai, 1912; and Mount Pinatubo, 1991). The largescale global climatic effects of such eruptions are only now beginning to be understood. Ecosystem responses to volcanism vary with the type, scale, frequency, and severity of the event, the nature of impacted vegetation, and contingent factors. Lavas destroy all biota and recovery rates on lavas are slow, but their texture helps to determine the rate of colonization and the nature of colonists. Lava succession rates appear to be more responsive to moisture than to temperature. Summer-dry lavas

2000 years old on Mount St. Helens remain dominated by mosses and lichens, with vascular plants in the cracks. Kamchatka lavas are well forested within 1500 to 2000 years. The effects of pyroclastic events and air-borne tephra depend on their intensity and scale, and the impacted biota. Forests are more resilient after tephra events than are shrub or grassland vegetation, since trees often survive impacts that kill other growth forms. Forests are also resilient to pyroclastic events, since growth-form diversity enhances the possibility that some individuals survive. Biogeographic factors strongly condition community development on volcanoes. The vegetation on many volcanoes is isolated from pools of suitable colonists. Extensive volcanic impacts require longdistance dispersal before plant succession can commence. Therefore, dispersal limitations often result in stochastic succession, disharmony, and novel species combinations. One implication is that restoration can be accelerated by overcoming dispersal barriers and planting species from more mature vegetation. A second, major and general implication is that one should not be concerned if the resulting species composition represents a mixture of native species different from that present prior to the eruption. Natural processes

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY

could also produce different successional patterns at different times and places within the same habitat. Understanding how volcanic landscapes recover provides information about fundamental ecological processes. Biogeographic effects largely determine species richness and composition, but abiotic factors such as weathering, erosion and nutrient inputs control the rates of change. Because the landscape context affects many factors in a stochastic way, whereas amelioration effects are more deterministic, the rate of succession may be more predictable than the course of succession after major volcanic eruptions. ACKNOWLEDGMENTS

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Chapter 6

BOREAL FOREST DISTURBANCES Ola ENGELMARK

INTRODUCTION

The boreal region is a northern, circumpolar forest belt covering 11% of the globe’s terrestrial surface (Bonan and Shugart, 1989). It includes a spectrum of forest vegetation types as well as disturbance regimes (Engelmark et al., 1993; Goldammer and Furyaev, 1996; Esseen et al., 1997). Boreal-forest dynamics are governed by a combination of various types of disturbance, both endogenic and exogenic, among which ﬁre is often claimed to be paramount (Heinselman, 1973; Tolonen, 1983; Johnson, 1992; Bradshaw et al., 1997). The role of ﬁre may, however, be over-emphasized in some boreal systems, and alternative types of disturbance may be more common than is generally realized (Bonan and Shugart, 1989; van der Maarel, 1993; Hofgaard, 1997). Other natural disturbances, such as herbivory, wind, ﬂooding, landslides, avalanches, and climate variability, may be more subtle but still have signiﬁcant effects on ecosystem properties (Oliver and Larson, 1990; Bergeron and Frisque, 1996). In addition, disturbances directly or indirectly caused by human land use may be of great importance for forest dynamics (Tolonen, 1983; Pyne, 1982, 1996). In this chapter, I use the term ‘disturbance’ as any event, discrete or chronic, that changes the environment and makes new growing space available (Pickett and White, 1985; Oliver and Larson, 1990; Engelmark et al., 1993). Effects of interactive disturbances may be very important in forest dynamics, as a low frequency of one type of disturbance may enhance both the occurrence and the ecological effects of another. Such variations between sites make predictions of successional pathways complicated (van der Maarel, 1993; Engelmark et al., 1993; Bergeron and Dubuc, 1989). In addition, it is urgent to consider the variations

Fig. 6.1A. Circumpolar distribution of boreal forest (from Wein and MacLean, 1983).

in disturbance regimes through time, which cause signiﬁcant differences in composition of the vegetation (Sprugel, 1991). Several authors have reported displacements of tree limits or changes in tree species distribution due to Holocene variability in climate and/or disturbance regimes (Kullman, 1990, 1996; Payette, 1992; Desponts and Payette, 1993). Such variability and disturbance interactions are also likely to occur in the future, and therefore, no disturbance

161

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Ola ENGELMARK

Fig. 6.1B. Distribution of dominant Pinus communities of the circumpolar north (from Wein and MacLean, 1983).

Fig. 6.1C. Distribution of dominant Picea communities of the circumpolar north (from Wein and MacLean, 1983).

or vegetation equilibrium is likely to exist over long temporal or large spatial scales (Cooper, 1913; Sprugel, 1991; Hofgaard, 1997). Thus, one must pay attention to temporal and spatial diversity when interpreting disturbance dynamics. Although similarities between boreal regions certainly exist, one must be careful not to generalize too much in trying to ﬁnd common ecological patterns. Instead, I suggest that spatially and temporally precise interpretations of vegetation dynamics following disturbance may contribute to a detailed understanding of mechanisms involved in controlling forest dynamics in the global boreal landscape. Further, regarding society’s urgent need for an ecologically sound and long-term sustainable use of natural forest resources, it turns out to be increasingly more important to base management and conservation strategies on a profound understanding of natural forest dynamics (Burton et al., 1992; Hunter, 1993; Gordon, 1996; Angelstam, 1997; Bergeron and Harvey, 1997). This chapter is not intended to be a detailed review of the copious literature available on disturbance dynamics in the large and spatially diverse borealforest belt. My aim is to delineate patterns for the

boreal zone, but also to elucidate ecological effects resulting from interactions between disturbances. I have chosen to review the predominant disturbance patterns in the boreal forest from the large scale (ﬁre and herbivory) to the small scale (internal stand-disturbance dynamics). I also mention the role of human land use, yet all disturbances are best interpreted within the broad context provided by climate. Further, ﬁre seems to have been the ‘agent of preference’ among scientists studying boreal forests, and is consequently best represented in the literature. I have therefore made the hopefully pedagogical and logical choice of presenting a gradient from the most ﬁre-prone boreal ecosystems (Pinus banksiana and P. contorta in North America), through Eurasian and North American ﬁrefavoured Pinus and Larix forests, through even less ﬁre-prone Abies and Picea forests, to ecosystems where ﬁre has no importance at all. The latter include northern European high-altitude Picea abies forests controlled by climate variability, and Betula pubescens forests in Fennoscandia altered by outbreaks of phytophagous insects.

BOREAL FOREST DISTURBANCES

Fig. 6.1D. Distribution of dominant Abies communities of the circumpolar north (from Wein and MacLean, 1983).

BOREAL BIOME CHARACTERISTICS

The boreal-forest belt, also called the taiga, extends over a vast area in the circumpolar north (North America: 3 730 000 km2 , Eurasia: 8 410 000 km2 : Wein and de Groot, 1996), and is thereby one of the two largest forest belts on the globe (Fig. 6.1A–E). The North American part is located primarily between 45º and 65ºN latitude, and the Eurasian part between 45º and 70ºN. Characteristic for the zone is a macromosaic formed by coniferous forests of Abies, Larix, Picea, and Pinus. The mosaic structure is due to variations in climate, topography, soil texture, and disturbance regimes. Pines are characteristically found on drier sites. In North America, lodgepole pine (Pinus contorta) grows in the west and jack pine (Pinus banksiana), red pine (Pinus resinosa) and white pine (Pinus strobus) in the east. Scots pine (Pinus sylvestris) is the most widespread pine in Eurasia, and is found from Scotland in the west to the east coast of Siberia. Stone pine (Pinus sibirica)

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Fig. 6.1E. Distribution of dominant Larix communities of the circumpolar north (from Wein and MacLean, 1983).

and dwarf pine (Pinus pumila) are also found in Siberia. Species of Picea, as well as Abies, are found on moister sites. Black spruce (Picea mariana) and white spruce (Picea glauca) grow over the entire boreal North America, and in the east balsam ﬁr (Abies balsamea) is also found. In Fennoscandia, Norway spruce (Picea abies) is characteristic, but it is replaced to the east by Altai spruce (Picea obovata). In boreal Siberia, Siberian larch (Larix sibirica) is gradually replaced by Siberian ﬁr (Abies sibirica), Larix sukaczewii and Dahurian larch (Larix dahurica) as one goes from west to east (Wein and MacLean, 1983; Walter and Breckle, 1989; Arnborg, 1990; Lenihan, 1993; Esseen et al., 1997). What distinguishes the boreal-forest belt from other forest biomes is primarily the climatic setting. Abiotic and biotic characteristics of the boreal forest, as compared with other types, include high seasonality, shorter growing season, lower temperatures, long summer day lengths, lower biological productivity, occurrence of permafrost, and a lower species diversity (Wein and MacLean, 1983; Walter and Breckle, 1989)

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than other forested biomes. This set of ecological parameters creates an environment prone to a spectrum of disturbances that vary greatly both in temporal and spatial aspects. The different disturbances occur on a variety of scales from a small gap in the canopy created by a single tree fall, to large areas over which the trees are killed and the canopy opened. Following the terminology given by Oliver (1981), and Oliver and Larson (1990), disturbances can be divided into two main groups, ‘major’ and ‘minor’. A major disturbance is deﬁned as one that kills practically all the trees above the forest ﬂoor (scale of stand), whereas a minor disturbance leaves groups of trees alive. Severe ﬁres or windstorms, glaciers, and extensive clear-cuts are examples of major disturbances, which promote the establishment of new, even-aged cohorts. Wind-throws, ﬁres, insect outbreaks, browsing, and partial cuttings are types of minor disturbance. Recruitment subsequent to the disturbance occurs in the canopy openings, and results in multiple-aged stand structures. Different species have different post-disturbance regeneration strategies (e.g., sprouting, seeds preserved in the soil, seeds stored in closed cones, shade intolerance) that confer various degrees of competitive ﬁtness. For example in relation to ﬁre, Rowe (1983) grouped plant species as endurers, resisters, evaders, invaders and avoiders.

FIRE

Fires generally create large openings in the forested landscape, although there is a great variety in boreal ﬁre regimes and related ecological responses (Sannikov, 1983; Swetnam, 1996; Bergeron et al., 1997; Flannigan et al., 1998). Higher ﬁre frequencies and more intense ﬁres have been reported in North America than in northern Eurasia, and species show different preferences to sites with different ﬁre regimes. In North America, the recurrent ﬁres have favoured trees with such ﬁre adaptations as serotiny (Pinus banksiana and P. contorta) and shade intolerance (P. resinosa and P. strobus). Species such as Picea glauca and Thuja occidentalis are more competitive at sites where ﬁre is of low importance (Heinselman, 1973; Bergeron, 1991; Quinby, 1991; Desponts and Payette, 1993). Spatial differences in ﬁre regimes in northern Europe also determine species distributions. Pinus sylvestris is found on regularly burnt sites with a dry, continental climate as found in northern Sweden (Fig. 6.2) (Uggla,

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1958; Engelmark et al., 1994). Picea abies dominates on sites where ﬁres occur rarely or not at all (e.g., high-altitude or maritime areas as in the Scandinavian mountain chain and in western Norway) (Arnborg, 1943, 1990; Engelmark, 1993a; Hofgaard, 1993a). Appendices A (Canada) and B (Sweden) present detailed examples of ﬁre frequencies and resultant forest structure and species composition. Different types of pine forests present different ﬁre regimes (i.e. ﬁre intensity, ﬁre frequency, timing (season) of the ﬁre and ﬁre size; Malanson, 1987). The North American species jack pine (Pinus banksiana) is considered to be the most ﬁre-prone and ﬁre-dependent tree species in the boreal area (Rowe and Scotter, 1973; Cayford and McRae, 1983; Weber and Stocks, 1998). It is the most widespread pine species in Canada, and extends from the Mackenzie Valley in the Northwest Territories southeastward in a broad swath to the upper Great Lakes and thence eastward to the Atlantic Ocean (Rudolph and Laidly, 1990). Although it is commonly found on sandy, gravelly, and nutrient-poor soils on terrace and outwash sandplains or on rock outcrops, it is also found on soils of medium to high fertility (Burns and Honkala, 1990; B´eland and Bergeron, 1993). Jack pine is a shade intolerant species, and stands usually begin to decay after 80 years. The foliage of jack pine is very combustible, dense forests often burn as intense crown ﬁres and the trees get killed (Van Wagner, 1967). Jack pines are adapted to frequent ﬁres by early seed production and rapid seedling growth in response to full light conditions (Rowe and Scotter, 1973; Cayford and McRae, 1983). Of particular interest is the serotiny of most jack pine cones. Temperatures of 50ºC or higher such as occur during forest ﬁres, are required to melt the resin and allow the cones to open (Mutch, 1970; Johnson and Gutsell, 1993). These serotinous cones allow jack pine to regenerate even if all the jack pine trees are killed by ﬁre. Fires remove organic material, sometimes expose the mineral soil, and create a generally favourable seed bed. If temperature and moisture conditions are favourable during the ﬁrst growing season, a new jack pine stand will probably establish (Cayford and McRae, 1983). The approximate ﬁre frequency in jack pine stands is 20–60 years (Wein and MacLean, 1983). Jack pine population dynamics depend on the prevailing ﬁre regime (Bergeron, 1991; Gauthier et al., 1993). Even-aged populations are more common in ﬁre regimes where lethal ﬁres occur with a frequency

BOREAL FOREST DISTURBANCES

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Fig. 6.2. Pinus sylvestris forest with ﬁre-scarred trees in Muddus National Park, northern Sweden (altitude 450 m). The oldest individuals germinated in the 1200s. Calluna vulgaris and Empetrum hermaphroditum are ﬁeld-layer dominants, with Cladonia spp. in the bottom layer (photograph by O. Engelmark).

of less than every 100 years. Further, lower ﬁre frequencies and/or non-lethal ﬁres allow for unevenaged structures to develop, as such ﬁre regimes seem to decrease the degree of serotiny (Gauthier et al., 1996). Stands in mesic habitats are more even-aged (Cogbill, 1985) than stands in xeric habitats, partly because natural openings in xeric sites allow for regeneration without ﬁre (Desponts and Payette, 1992). Lodgepole pine (Pinus contorta) is considered as the second most ﬁre-dependent species in the boreal area. It has a wide ecological amplitude, from dry to moist, from well-drained to poorly-drained soils (Lotan and Critchﬁeld, 1990). It grows throughout the Rocky Mountain and Paciﬁc coast regions, from Yukon Territory in the north and south to Baja California. Fire is the common agent for regeneration of the species. Fire prepares the seed bed by removing the organic layer as mineral soil is required for seedling establishment. Fire also removes competition and opens up the canopy. As with jack pine, lodgepole pine cones are often serotinous. Thus, jack and lodgepole pines have adapted and are dependent upon a ﬁre regime

of recurrent, high-intensity ﬁres. These recurrent ﬁres are usually lethal crown ﬁres which are common in the boreal forest. Fire frequencies in lodgepole pine stands are similar to ﬁre frequencies in jack pine stands (Johnson and Gutsell, 1993), and often lead to evenaged cohorts, although lodgepole pine stands are often multi-aged in the absence of ﬁre (Despain, 1983). Red pine (P. resinosa) and white pine (P. strobus) have ranges that extend from southeastern Manitoba to the Atlantic Ocean, and as far south as Pennsylvania for red pine and northern Georgia for white pine. Although both pines are commonly found on sandy and gravelly, poor soils on terrace and outwash sand-plains or on rock outcrops, they are also found on soils of medium to high fertility (Burns and Honkala, 1990). Red and white pine stands persist for 200 years or more. Red pines are shade-intolerant while white pine is somewhat shade-tolerant. Red and white pine are also adapted to forest ﬁre. Fire is the common agent for regeneration, although white pine regeneration by other means is possible because of some shade tolerance. Red and white pine

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do not have serotinous cones and the trees must survive ﬁre for regeneration. Mature red and white pine both have a thick bark which enables the trees to survive all but the most intense surface ﬁres and crown ﬁres. An ideal disturbance regime for red and white pine would be for a moderate-intensity surface ﬁre to occur every 50 to 150 years. This ﬁre environment is more common in the Canadian mixed-wood forests of Betula, Picea, Pinus, and Populus, and in ﬁre-sheltered areas of the boreal forest (Heinselman, 1981; Bergeron and Brisson, 1990). Fire-sheltered locations include islands, protected lake-shores, and areas of rugged terrain (van Wagner, 1970). Major ﬁres are not necessary for white pine recruitment, as white pine can re-establish following small ﬁres, wind-thrown trees, or death of individual trees (Quinby, 1991). The result is either even-aged or multi-aged stands. At the northern limit of red and white pine, where the ﬁre regime is one of frequent and intense crown ﬁres, these pines are restricted for the most part to ﬁre-sheltered locations, and the populations are therefore small and scattered. Scots pine in Fennoscandia and Siberia grows preferably on coarse, well-drained, and sometimes nutrientpoor soils. Stands of Scots pine most often burn as lowto medium-intensity surface ﬁres – that is, the organic material, the ground-layer (lichens) and the ﬁeldlayer vegetation (dwarf shrubs) constitute the principal fuel. During drought conditions, dense populations of Cladina, Cladonia, Empetrum, and Vaccinium species are important as a continuous fuel source for surface ﬁres. Some tall pines survive and often become ﬁrescarred (Fig. 6.2). Pines and other shade-intolerant trees such as Betula pubescens and Populus tremula regenerate on the burnt area. Trees of Scots pine that suffer mechanical damage (e.g., bark-stripping, ﬁrescarring) tend to produce much resin, which gives them better resistance against pathogens (Gref and Ericsson, 1985) and favours their longevity (Engelmark, 1987). The oldest known pines in Fennoscandia are 700– 800 years old and were found in ﬁre-prone areas (Engelmark and Hofgaard, 1985; Engelmark et al., 1994). Naturally, such old pines with ﬁre scars are very useful when studying ﬁre history. According to Engelmark (1984, 1987) the mean ﬁre frequency in pine forests at 67ºN close to the Arctic Circle in Sweden, was once in 110 years (range 40–277 years). At 65ºN, in pine forests along a river valley in northern Sweden, Zackrisson (1977) obtained a mean frequency of once in 55 years. An overview of papers on ﬁre history (Flannigan

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et al., 1998) covering the last millenium shows a range of frequencies from once in 30 to once in 300 years in northern Fennoscandian pine forests, the most common intervals between recurrent ﬁres being about 70–80 years. Extending the time span, Tolonen (1983) showed in Finnish spruce stands that natural (pre-settlement) ﬁre frequency in Holocene times (between 5000 and 2000 BP) was once in 158 years for local ﬁres, and 238 years for ﬁres on a regional scale. Historically, ﬁres have been more frequent in the interior parts of Fennoscandia. The intervals between ﬁres are generally longer at high altitudes and in tundra areas. This is partly coupled to the climatic situation, where lightning frequency, amount of precipitation, and air pressure are important factors affecting ﬁre occurrence. More than 60 species of vascular plants, fungi, lichens, and invertebrates in Fennoscandia are considered ﬁre-dependent, due to certain regeneration strategies (see Esseen et al., 1992; Wikars, 1997). Among vascular plants, Esseen et al. (1992) suggested that Anemone patens, A. vernalis, Chimaphila umbellata, Geranium bohemicum, and G. lanuginosum were ﬁre-dependent. There are additional species favoured by ﬁre though not dependent on it, such as the vascular plant Epilobium angustifolium, and mosses such as Ceratodon purpureus, Funaria hygrometrica, Polytrichum juniperinum, and P. piliferum (Esseen et al., 1997). In Fennoscandia, few studies on the dynamics of post-ﬁre vegetation are available (but see Uggla, 1958). However, Schimmel and Granstr¨om (1996) have suggested that depth of burn (a measure of ﬁre severity) greatly affects the range of plants which establish. This interacts, of course, with the composition of the pre-ﬁre vegetation and the survival of different species in relation to their respective regeneration strategies (seeds to be dispersed after ﬁre, as for Epilobium angustifolium; or having a soil seed bank, as for Luzula pilosa; or through rhizomatous regeneration, as for Vaccinium myrtillus and V. vitisidaea). Scots pine regeneration is reported to occur in a multi-aged pattern, partly related to post-ﬁre situations (Engelmark et al., 1994). But in the northern boreal area this pattern might also be related to the climate in the sense that summer warmth is decisive for the success of seed ripening and seedling establishment (Zackrisson et al., 1995). Deciduous forests mainly of birch (Betula spp.) and aspen (Populus tremula), establishing early after forest ﬁres on mesic soils, may often host a large variety of species. Examples from Fennoscandia are Pyrola

BOREAL FOREST DISTURBANCES

media; and lichens such as Collema curtisporum, C. fragrans, and Ramalina sinensis; and fungi such as Daedaleopsis septentrionales, Gloiodon strigosus and Haploporus odorus (see Esseen et al., 1997, and literature cited therein). The areal extent of deciduous forests has diminished in Fennoscandia, both as a result of fewer ﬁres and also because of a forestry landuse directed towards coniferous silviculture (Sj¨oberg and Lennartsson, 1995; Esseen et al., 1997). As a consequence, the lack of deciduous habitats obstructs and even threatens species favoured by such habitats. In Canada, on the other hand, large deciduous-forest landscapes can still be found, but have to be carefully managed to preserve diversity (Burton et al., 1992; Stelfox, 1995). The Siberian taiga is divided into two main parts, the dark and the light taiga. The dark taiga has a tree stratum of Altai spruce (Picea obovata), Siberian ﬁr (Abies sibirica), Siberian larch (Larix sibirica), and stone pine (Pinus sibirica), whereas the light taiga is dominated by Dahurian larch (Larix dahurica) in a spectrum of different associations with other trees (cf. Walter and Breckle, 1989). Larch can broadly be regarded as being as ﬁre-adapted as pine (cf. Sannikov, 1981). In the eastern light taiga the average interval between ﬁres varies between 5 and 20 years, and the regeneration of Dahurian larch is principally a postﬁre phenomenon. If ﬁre does not occur more or less regularly, the proportion of the less ﬁre-resistant species Siberian pine (xeric habitats), or Siberian ﬁr (mesic habitats) increases (Babintseva and Titova, 1996). The ﬁre ecology of larch forests is not well studied, but the following example of how ﬁre may change the successional patterns has been presented by Walter and Breckle (1989). Dahurian larch covers more than 200×106 ha in the eastern part of Siberia, north of 49ºN, where permafrost soil is 250–400 m thick. The existence of permafrost creates particular conditions for tree growth and ﬁre. In the continental parts, extremely cold winters are followed by very warm summers. The thawing of permafrost in summer, which creates the water supply, is a prerequisite for the existence of larch. Furthermore, the freezing/thawing conditions cause thermokarst phenomena of considerable proportions. Large-scale natural forest ﬁres occur on average at intervals of 180–240 years. The heat of the ﬁre itself does not appear to inﬂuence the permafrost. However, the insolation on the ground increases after a forest ﬁre, which leads to deeper soil thawing (i.e., increased active layer), soil collapses, and undrained depressions

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that ﬁll with the rising ground water. Eventually, new ice forms and ice-ﬁlled pingos rise. Under these conditions the soil is well drained, and a Scots pine forest can establish. The pingos continue to grow for up to 1000 years, and then collapse to form new depressions. Another development is described from Siberia outside the permafrost zone by Uemura et al. (1990). They found that larch dominated most of the sites except the most xeric ones where Scots pine dominated, and the moistest sites which favoured spruce. Sites of intermediate soil moisture were dominated by larch. Scattered larches were also found in the spruce stands, indicating that larch had preceded spruce. The dominance of shade-intolerant larch is presumably related to the high ﬁre frequency (Uemura et al., 1990). Fire is an obstacle to spruce, especially in dry sites, but favours larch and pine. Stands of black spruce (Picea mariana) in northern Canada and Alaska usually burn frequently. According to Van Wagner (1983) and Viereck (1983), black spruce forests in North America burn as high-intensity, large, combined surface and crown ﬁres, with a frequency of about once in 90 years (Rowe et al., 1974). The revegetation of black spruce forests after ﬁre is complex. Depending on ﬁre intensity, aspect, altitude, presence of permafrost, and other factors, the successional patterns are highly variable. Black spruce seems to be ﬁre-adapted as it has semiserotinous cones, and a burnt black spruce stand is often succeeded by a new even-aged cohort of black spruce. In some cases, one generation of deciduous species such as paper birch (Betula papyrifera) and aspen (Populus tremuloides) precede the black spruce regeneration (Fig. 6.3). Several authors have also pointed out the interactive role of ﬁre and outbreaks of spruce budworm in the Canadian black spruce belt, which has contributed to multi-aged forests of black spruce and balsam ﬁr (Abies balsamea) (Viereck, 1983; Morin, 1994), and see pp. 168–172 below. Forests of white spruce in Canada and Alaska most often grow where ﬁre frequencies are lower, for instance in association with mires, rivers, or more nutritious sites, or in wet depressions where ﬁres generally do not spread (Quirk and Sykes, 1971). Norway spruce (Picea abies) in Fennoscandia prefers silty and nutrient-rich soils. It is favoured by low ﬁre frequencies or the absence of ﬁre. Spruce stands commonly burn as crown ﬁres and few trees survive, except as unburnt patches that play an important

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Fig. 6.3. Theoretical successional relationship of black spruce forests in relation to ﬁre, site type, and other forest types in Alaska and Canada (from Viereck, 1983).

role in the forest recovery process (Tir´en, 1937; Uggla, 1958; Engelmark, 1987). Survivors may die due to fungal attack. On the burnt area, broadleaved deciduous species such as birch and aspen regenerate abundantly. Spruce starts to regenerate in these young broadleaved stands. After slightly more than a century, when birch and aspen senesce, a new spruce stand establishes. After approximately another 300 years the spruce forest has accumulated sufﬁcient fuel (e.g., dead standing stems) which, during dry conditions, makes the forest susceptible to renewed burning. This pattern is not applicable everywhere, and there are contradictory views on the post-ﬁre dynamics of Norway spruce forest. One study from northern Sweden (Engelmark, 1993b) following 26 years of post-ﬁre tree regeneration, showed that although the closed pre-ﬁre stand was dominated by Norway spruce, the post-ﬁre stand consisted of roughly equal amounts of Norway spruce and Scots pine (Pinus sylvestris). This was ascribed to increased browsing by reindeer (Rangifer tarandus) of the post-ﬁre birch cohort facilitating pine establishment. In other words, the interactions of ﬁre and browsing led to a shift in dominance of tree species. Similar results were obtained in another northern Swedish stand of Norway spruce by Steijlen

and Zackrisson (1987). Although ﬁre had not occurred for about 500 years, Scots pine had recruited in gaps of the spruce canopy. This was perhaps related to the climatic amelioration from the early 1900s, which gave pine a competitive advantage over spruce. Boreal ﬁr forests have a relatively low ﬁre frequency, mainly because many of these forests grow in areas with high rainfall. High air humidity and high fuel moisture make ignition less probable. The ﬁre frequency of ﬁr forests is poorly quantiﬁed, but a very approximate ﬁgure is once every 300– 400 years (Furyaev et al., 1983; see also Bergeron and Leduc, 1998). High-intensity crown ﬁres may occur after severe drought, and are even more likely following insect outbreaks causing tree defoliation or tree mortality and increased ﬂammability. This theory is called “the insect–wildﬁre-hypothesis” (Heinselman, 1981; Furyaev et al., 1983). Firs are shade-tolerant and not ﬁre-resistant, and all these forests change type after a ﬁre. Balsam ﬁr (Abies balsamea), in eastern North America tends to be replaced by species of Populus or Betula after ﬁre (Bergeron and Dubuc, 1989). Smirnov (1970) suggested that Alnus fruticosa or Spiraea media may establish after ﬁre in stands of Siberian ﬁr (Abies sibirica), and also persist for a long time. However,

BOREAL FOREST DISTURBANCES

the post-ﬁre succession in ﬁr stands is closely related to outbreaks of the North American spruce budworm (Choristoneura fumiferana) and the Siberian silkworm (Dendrolimus sibiricus). Although restricted by a successively harsher climate, ﬁre also plays a role far north and at high altitudes in controlling tree distribution limits. In Canada, it has been suggested that the northern limits of pine species are more or less ﬁre-controlled rather than being controlled solely by climate – for instance, jack pine (Desponts and Payette, 1992), red pine (Flannigan, 1993; Flannigan and Woodward, 1994) and white pine (Holla and Knowles, 1988; Abrams et al., 1995). As mentioned earlier, black spruce is also a ﬁre-adapted, semi-serotinous species, and can therefore regenerate successfully after ﬁre, but can also recruit without ﬁre at its northern distribution limit provided that ripe seeds are available (Payette et al., 1989a,b). Even though information on ﬁre regimes at high altitudes in northern Fennoscandia are scanty, some kind of ﬁre/climate relationship seems reasonable in the sense that ﬁre-free sites tend to favour spruce; however, this is not straightforward. Kullman and Engelmark (1997) suggested that the present latitudinal tree limit for Norway spruce is not determined by variations in summer temperature, but instead mediated by early and late winter conditions, primarily snow depth, surface water mobility, and consequently also soil freezing and related occurrence of permafrost on ﬁre-free sites. The more northerly limit for Scots pine is instead mainly controlled by summer temperature (Kullman, 1996). In contrast to this latitudinal pine/spruce distribution, Norway spruce generally reaches higher altitudes than Scots pine in its western distribution along the Scandinavian mountain chain (Nilsson, 1897; Kullman and Engelmark, 1991). These altitudinal differences are generally explained by the lower temperature demands for success of spruce regeneration (Mikola, 1971; Kielland-Lund, 1981), although other factors may also be important.

HERBIVORY AND TREE DEFOLIATORS

Browsing, grazing and tree defoliation are well-known disturbances structuring the boreal forests. Common herbivores include ungulates (Bryant and Chapin, 1986) such as moose (Alces alces) (Danell et al., 1991; Edenius, 1991; Thompson and Curran, 1993) and reindeer (Rangifer tarandus) (Oksanen et al., 1995;

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V¨are et al., 1996), beavers (Castor canadensis and C. ﬁber) (Johnston et al., 1993), other rodents (Ericson and Oksanen, 1987), and insect defoliators (Volney, 1988; Holling, 1992; Morin et al., 1992; Bylund and Tenow, 1994). Grazing or browsing can be regarded as a disturbance acting more or less continually in forests, and may therefore be termed ‘chronic’ (Engelmark et al., 1993), in contrast to ‘pulsed’ disturbances such as ﬁre. For example, the effects on vegetation of grazing might not be obvious until the grazing changes in magnitude, or alternatively ceases and a vegetative recovery process begins (Hofgaard, 1997) [see also elsewhere in this volume (Schowalter and Lowman, Chapter 9; Willig and McGinley, Chapter 27) for more discussion on herbivory and disturbance]. The effects of herbivores on plant communities seem to be dichotomous in the sense that these disturbances may retard or alternatively hasten plant succession (Peet, 1988; Veblen et al., 1989; Davidson, 1993) (Table 6.1). For the boreal forest, however, a majority of studies show that succession is accelerated by herbivory. Usually the succession starts with deciduous trees and shrubs and proceeds with evergreens (mostly conifers) (Bergeron and Dansereau, 1993; Esseen et al., 1997). When early-successional deciduous species (generally palatable) diminish due to herbivory, this enables an earlier start of a later successional sere of conifers (Walker et al., 1986). Further, conifers are generally both less palatable and less digestible than deciduous trees (Bryant et al., 1983; Bryant and Chapin, 1986; Davidson, 1993). Beavers act as keystone species in remoulding structure and dynamics of boreal landscapes (Naiman et al., 1986; Johnston et al., 1993). By feeding, tree-felling, and dam-building they impact both vegetation dynamics and ecosystem properties. Beavers also have the capacity to affect the tree canopy. Tree-felling causes canopy openings which change the light regime, alter competition and species composition, and increase the amount of coarse woody debris. The dam-constructing activities change the water ﬂow, and consequently sediment transportation and biogeochemical pathways, but also cause ﬂooding. Forests in areas inﬂuenced by beavers have a patchy structure, characterized by a series of canopy openings all of different ages, dying trees and stands, or stands which are rejuvenating following beaver-caused disturbance. Certain areas may become deforested because of sediment coverage or a more or less permanent ﬂooding. This structural diversity contrasts with areas where beavers do not

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Table 6.1 Examples of how herbivory on species of earlier seral stages can accelerate succession (sensu Davidson, 1993) Herbivore

Earlier seral stage

Later seral stage

Place

Odocoileus spp.

pioneer trees

Betula papyrifera

Minnesota

Cervus elaphus

Sambucus shrubs

Pseudotsuga douglasii

Washington

Ungulates

grasses

Acacia trees

Uganda

Lepus americanus

deciduous shrubs, trees

Picea

Alaska

Castor canadensis

Populus tremula

Picea mariana

Minnesota

Alces alces

Abies balsamea

Picea glauca

Wisconsin

deciduous trees

Abies balsamea, Picea

Michigan

Caterpillars

Populus tremula, shrubs

Abies balsamea

Minnesota

Choristoneura fumiferana

Pinus contorta, overmature Picea

Picea, Abies

Colorado

occur. Although the size of beaver populations has varied through time, beavers are presently common both in northern Eurasia and in North America, and their ecological impact is visible everywhere in the landscape (Naiman et al., 1986; Nilsson, 1992). Johnston et al. (1993) showed how beaver impoundments in Minnesota increased the aquatic habitat by 438%, and thereby converted a large peninsula measuring 298 km2 to a spatially diverse mosaic of aquatic and terrestrial habitat. Moose are habitat-selective animals showing seasonal preferences for different stand and vegetation types (Bergstr¨om and Hjeljord, 1987). They frequent a range of habitats from northern mountain birch forests and slowly growing coniferous stands to more productive forests, such as riparian or southern deciduous forests. During winter their main fodder is Scots pine (Pinus sylvestris). If there is a shallow snow cover, moose also feed on willow (Salix spp.) and lingonberry (Vaccinium vitis-idaea) (Cederlund et al., 1980; Bergstr¨om and Hjeljord, 1987). The main summer fodder besides willow is birch leaves and twigs, and ﬁreweed (Epilobium angustifolium). The preference for feeding on shade-intolerant species such as ﬁreweed, pine, and birch, may favour the establishment of late-successional species such as the shade-tolerant spruce (cf. Walker et al., 1986), particularly in mesic soils. Palatability of the pines may reﬂect soil fertility (Bryant et al., 1987; Edenius, 1991). Reindeer have been raised as a semi-domesticated animal for at least the past 1000 years in northern Fennoscandia (Baudou, 1988; Aronsson, 1991). As a

result of increasing reindeer herds, grazing of lichens is also increasing. Some areas are today so overgrazed that winter feeding with hay or imported lichens is necessary. According to V¨are et al. (1995, 1996), high grazing pressure has resulted in a signiﬁcant decrease of the late-successional species Cladina stellaris (grazed: 3% coverage, ungrazed 60%), while early colonizers such as C. arbuscula and C. rangiferina beneﬁt, as do the mosses Dicranum spp. and Pleurozium schreberi. Also, the biomass of dwarf-shrubs such as Calluna vulgaris, Empetrum nigrum, and Vaccinium vitis-idaea and pine root biomass decreased signiﬁcantly. All exchangeable nutrients decreased by 30–60% in the organic layer, and this may affect longterm site productivity. On the other hand, the reindeer grazing/browsing has been reported as positive for maintaining a continuous seedbed for shade-intolerant Scots pine recruitment (Engelmark et al., 1998). Reindeer sometimes also trample pines and scrape the trees with their antlers. Eriksson and Osterman (1996) found that 0.2% of the pines were subjected to scraping of antlers, and thereby damaged (as against 10% of the pines in the same area that were damaged by moose). Although ﬁre is the dominant type of disturbance in the Canadian black spruce belt, if ﬁre does not occur in these generally ﬁre-prone systems (e.g., in eastern Canada), a succession towards dominance by balsam ﬁr may occur (Damman, 1964; Morin and Laprise, 1990). Balsam ﬁr forests undisturbed by ﬁre are modiﬁed instead by regular outbreaks of spruce budworm (Choristoneura fumiferana). The more intense outbreaks seem to favour spruce as ﬁr is more vulnerable. In less intense outbreaks, ﬁr seems to be more competitive

BOREAL FOREST DISTURBANCES

171

Fig. 6.4. Epirrita autumnata outbreaks reported in Fennoscandia north of 64ºN latitude. (a) The extent of defoliation is expressed as incidence in 0.2º latitudinal belts; note that zero on the Y-axis corresponds to latitude 64º. (b) Serial correlation analysis shows cycle periodicity to be 9 years (from Haukioja et al., 1988).

than spruce (Holling, 1992). The spruce budworm denudes and causes extensive mortality in the overstory trees, and afterwards the larvae may also feed on the seedlings and saplings. However, the seedling and sapling mortality is considered low, allowing for a new ﬁr/spruce forest to establish. Therefore no major changes in forest composition usually occur (Mattson et al., 1988). As suggested by Baskerville (1975) and Morin (1994), this regime of recurrent outbreaks of spruce budworm in balsam ﬁr forests is forming an independent, self-regulating system in the black spruce belt. The outbreaks occur approximately every 40 years (Holling, 1992; Morin, 1994). Stand and forest patch composition seem important to explain the variation in outbreak frequency and intensity, and a higher proportion of deciduous species decreases the severity (Bergeron et al., 1995). This cyclic system is in contrast to outbreaks of spruce beetle (Dendroctonus ruﬁpennis), which caused large canopy gaps in forests in the southern Rocky Mountains, releasing subcanopy trees and thereby hastening the succession towards dominance of shade-tolerant ﬁr (Veblen et al.,1989, 1991).

Another example of tree defoliation is from highaltitude Fennoscandian Betula forests. Above the coniferous forests in the Scandinavian mountain chain there is a belt of mountain birch forests. Mountain birch (Betula pubescens ssp. tortuosa) is often a polycormic tree that is only used for small scale cutting of ﬁrewood. The dynamics of mountain birch forest are governed primarily by the climate, and ﬁre is an insigniﬁcant disturbance. However, outbreaks of phytophagous insects occur. Outbreaks of the autumnal moth Epirrita autumnata, are best known, and have large impacts on the ecosystem. During some outbreaks, defoliation by Epirrita causes extensive birch mortality at the treeline (e.g., in Utsjoki, northern Finland, where 10% of an area of 5000 km2 became treeless tundra) (Haukioja et al., 1988). More commonly, defoliated trees recover, while the ground ﬂora changes, as grasses and herbs increase in cover (Haukioja et al., 1988). The outbreaks occur generally at regular intervals of 9 or 10 years (Haukioja et al., 1988) (Fig. 6.4), but there are also factors that interact to inﬂuence the disturbance regime. For example, in landscape depressions, cold winter air temperatures destroy the Epirrita eggs.

172

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Fig. 6.5. Schematic cross-section of a regeneration wave of Abies balsamea, described by Sprugel (1976), from Whiteface Mountain in the state of New York, U.S.A.

Epirrita outbreaks may promote the polycormic state of birch, because monocormic birches are more common in these ‘cold air lakes’ (Tenow and Nilssen, 1990). Tree age affects growth of Epirrita (mature forests are less resistant). Larval densities may be about twice as high in birch stands growing on less fertile soils – that is, the foliage of vigorous birches has a higher resistance. The leaves formed after defoliation are smaller, have a lower quality as food, and are covered with a denser layer of hairs. These patterns suggest a phenotypic plasticity characteristic for stress-tolerant plants. In the years following defoliation, production of eggs by Epirrita decreases drastically. Two defoliations in one year would be fatal, as the plant’s reserves are depleted. There is no information on whether these outbreaks also occur at lower altitudes, in the mountain conifer forests, or how they may interact with ﬁre occurrence.

INTERNAL STAND DISTURBANCES

In areas where ﬁres, extensive insect outbreaks, or other major canopy disturbances are unimportant, minor openings in the canopy play a central role for ecosystem rejuvenation (Hytteborn et al., 1987; Hofgaard, 1993a,b; Kuuluvainen, 1994). These openings are often made by a variety of causes, principally the ageing and death of single trees. Another cause is wind disturbance (e.g., Mitchell, 1995; Peltola, 1996), which uproots trees or snaps tree stems. Depending on wind speed (Peltola, 1996), possible snow load in trees (Solantie, 1994; Nyk¨anen et al., 1997), and stand characteristics, wind disturbance can create canopy gaps of many sizes. The canopy gaps cause a mosaic pattern of stand stages, which, on a landscape scale, present a multi-aged stand structure (Oliver and Larson, 1990; Hofgaard,

1993a,b; see also Webb, Chapter 7, and Whigham et al., Chapter 8, this volume). A peculiar type of wind-triggered disturbance pattern (wave regeneration) was described from the northeastern United States by Sprugel (1976) (Fig. 6.5). Balsam ﬁr died off at the edge of a ‘wave’ and was replaced by a new cohort of ﬁr. This implies that when old and/or weakened trees were exposed in a forest opening, and blew down, this would expose the trees behind them, which died in turn. This cyclic regeneration pattern has a periodicity of about 60 years. A similar wave regeneration pattern of about 100 years has been described in ﬁr forests in Japan (Iwaki and Totsuka, 1959). The forest under this type of disturbance regime becomes multi-aged and might, on a landscape level, be regarded as maintaining a stable, mosaic structure of regeneration and degeneration stages (Sprugel, 1976). Although ﬁr waves are characteristic of temperate forests, such structural stability on a landscape level has also been discussed for boreal forests (e.g., Zackrisson, 1977; Bergeron and Dubuc, 1989). Alternatively, a steady state might never be reached (in time or space) in a continually changing climate. Hofgaard (1993a,b, 1997) suggested that oldgrowth stands of Norway spruce forests in northern Sweden can be maintained by small-scale disturbances such as tree-falls, which in turn create a continually changing mosaic of differently-aged patches. Where spruce is the dominant conifer in Fennoscandian forests, it may reach ages of 400–450 years (Hofgaard, 1993a; Kullman, 1995; Rolstad et al., 1996). Openings in such forests (as a result, for instance, of tree-falls and uprootings) cause a spectrum of microenvironments which alter light regimes, and competitive interactions, and may thereby increase the diversity of species. Such forests may thus be of critical importance to many species of vertebrates,

BOREAL FOREST DISTURBANCES

173

Fig. 6.6. Relation between degree of decomposition (DD) of spruce (Picea abies) logs and age of pooled regeneration of spruce and birch (Betula pubescens) on logs. DD ranges on a scale from 1 (solid wood) to 8 (pulverized wood). Number of regenerated individuals, mean age and range are shown. Colonizing bottom- and ﬁeld-layer species on logs: Cla. spp. (Cladonia), Des. ﬂe. (Deschampsia ﬂexuosa), Dic. spp. (Dicranum), Hyl. spl. (Hylocomium splendens), Par. spp. (Parmeliopsis), Ple. sch. (Pleurozium schreberi), Vac. myr. (Vaccinium myrtillus) (from Hofgaard, 1993a).

invertebrates, and cryptogams (Berg et al., 1993). Some species of fungi, bryophytes, lichens, and insects are conﬁned to old-growth stands, and are therefore often used as indicators of biodiversity and forest continuity (Jonsson and Esseen, 1990; Berg et al., 1993; Renvall, 1995; Esseen et al., 1997), although their value as indicating old-growth stages can be questioned (Ohlson et al., 1997). Further, decomposing wood is an important factor for the regeneration of spruce (Harmon et al., 1986). Hofgaard (1993a) reported that 40% of spruce regeneration occurred on decomposing logs and stumps, although they covered only 6% of the forest ﬂoor (Fig. 6.6).

HUMAN IMPACT

Human land-use may cause both major disturbances such as clear-cuts, as well as minor disturbances such as may result from gathering wood to build a campﬁre. All these land-uses potentially disturb the environment by creating new growing space. Many species are generalists, and can therefore colonize areas disturbed by people. However, in the case of forestry, new, often exotic species are planted after cutting. Such

major disturbances thus cause highly unnatural, often unpredictable, secondary successions. Dissemination of toxic compounds, lake acidiﬁcation, mining, river exploitation, road construction, and city extension are other examples of human impact on boreal forest ecosystems that are beyond the scope of this chapter. Man has utilized boreal forests almost continuously since the latest deglaciation. If one looks at Sweden as an example, the oldest known Mesolithic sites are 8000 years old (Baudou, 1988). Hunting and gathering gradually were replaced by agrarian activities in Sweden, beginning around 5000 BP. Commercial tree cutting did not start until the 1600s in connection with the development of the mining industry. Large-scale tree harvesting for saw timber began in the 1800s in the boreal parts of Sweden. The pulp industry started to develop in the late 1800s. The efﬁciency of commercial tree harvesting increased rapidly, and some far-sighted foresters realized that forest resources were not endless. Therefore, as a consequence, the ﬁrst Forestry Act in 1903 required that tree regeneration measures had to be undertaken following forest removal. Hunting, animal husbandry, and recreation have also affected Swedish forests, and there has recently been much public debate regarding environmental concerns on forested land. For

174

further reading on past but also present uses of land in northern forests, the reader is directed to Hyt¨onen (1995) and Peterken (1996). Human land-use, especially large-scale commercial forestry, involves several types of disturbance, but has also changed the stand and landscape structures drastically and thereby also inﬂuenced natural disturbance regimes. For example, large clear-cuts are not ecologically equivalent to burnt forests with respect to post-disturbance stand structure or vegetation dynamics. This changed landscape structure gives different conditions for ﬁres to start and spread. A decreased ﬁre frequency observed in many boreal areas during the past century might partly be related to this structural change due to forestry, but also partly due to climate variations (see below). Another example is that 565 000 ha of forested land in Sweden are planted with the North American species lodgepole pine (Pinus contorta). Despite the large area of these alien stands, very few ecological assessments have been done – to ﬁnd out, for example, what can be expected regarding self-dispersal, effects on native plant and animal communities, and ﬁre occurrence (cf. Sj¨oberg and Lennartsson, 1995). At present however, there is a global consensus that natural resources should be used with a sustainable vision. It has therefore been suggested many times that natural disturbance dynamics may be used as a template for forest management, with the overall aim to maintain both productivity and biodiversity in the ecosystems. These thoughts have been developed by Hunter (1993), Attiwill (1994) and Angelstam (1997), and by Binkley in Chapter 18 of this volume, all with the basic concept that planning should be on the scale of a landscape.

CLIMATE CHANGE AND DISTURBANCE

Ecosystem change is the norm in the boreal biome. During the Quaternary period (the last 2.3×106 years) there has been a cycle of warm and cold stages where recurrent glaciations have been the most common state, whereas the interglacials represent less than 10%. The present interglacial period, the Holocene, started about 10 000 years ago (Donner, 1995). Whatever future climate situation the boreal zone will face during the remaining part of the Holocene, it will be different from that today and from those earlier in the Holocene. One can expect new disturbance patterns and new ecological responses. As Sprugel (1991) put it; ‘Every

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point in time is special’, implying that an ecosystem cannot always, if ever, return to its predisturbance state. Instead, one has to recognize the transitory character of concurrent disturbances that have unpredictable but also long-lasting effects on the boreal ecoystem. It is important to strive towards increased understanding of what triggers different disturbances and attendant dynamic effects, so that a more complete picture of the dynamics of boreal ecosystems can be envisaged. Climate change may affect the distribution of the boreal zone, but disturbance patterns and processes within the boreal area also may be affected in the case of global warming or cooling. Fire and herbivory are key disturbances, and, as suggested by Bergeron and Flannigan (1995) and Fleming (1996), effects of climate change on disturbance dynamics might be greater than the climate change per se. Bergeron (1991) and Bergeron and Archambault (1993) demonstrated that, at the onset of an earlier warming (after the Little Ice Age, about 1870), the forest-ﬁre frequency in eastern Canada decreased signiﬁcantly. The same pattern was also reported from northern Sweden (Engelmark et al., 1994). This might be explained by the fact that a warmer climate also involves higher humidity and thereby a decreased ﬁre hazard. This is further supported by Swetnam (1996) who, in the Yenisey region in Siberia, found that ﬁres occurred more often during dry years prior to 1880 than after that date. He interpreted this as implying that the causes of ﬁre occurrence after 1880 were related rather to the increased inﬂuence of people than to the climate. On the other hand, in some areas a global change might involve a drier climate, which naturally would increase the risk of ﬁres (Clark, 1990). Johnson and Larsen (1991) suggested that the effect of regional climate on fuel moisture and ﬁre risk was the most useful factor in understanding changes in ﬁre frequency. Changes in disturbance regimes have ecological consequences for forests. For example, Landh¨ausser and Wein (1993) suggested that, for the western Canadian Arctic, warming involving more frequent ﬁres would result in a stepwise northward migration of the treeline, with each ﬁre providing a favourable seedbed (Shankman, 1984; Scott et al. 1987). In more general terms this is supported by Payette and Gagnon (1985), who concluded that failure or success of post-ﬁre tree regeneration close to the tundra is delicately balanced by ﬁre and climate variability. A drier situation would also induce water stress for trees, in turn increasing the sucrose concentrations

BOREAL FOREST DISTURBANCES

(Mattson and Haack, 1987) which in turn would improve the feeding possibilities for insect larvae such as spruce budworm (Albert et al., 1982; Mattson et al., 1988). Such climate effects could eventually lead to more extensive insect outbreaks (Fleming, 1996), affecting the composition of the boreal forest considerably. In addition, as emphasized by Flannigan et al. (1998) any future climate change must be viewed in a spatially dependent context. By combining records of ﬁre history and simulation results for future climate and ﬁre hazard, they could show that, even if a global warming occurs, ecological consequences as well as effects on ﬁre regimes would probably be very different both between North America and northern Europe, and within each area. Prentice et al. (1991) pointed out that, consistent with past climate changes and related vegetation responses, a future climate change would reasonably imply that different tree species will respond individualistically in range expansion or contraction through their species-speciﬁc climatic demands. Such individualistic responses may in turn lead to new species composition in forests (see also Davis, 1989; Payette et al., 1989a,b). To conclude, any synthesis regarding past, but also possible future changes in boreal vegetation or plant succession has to be done with great caution. In a biome such as the boreal forest, where many disturbances interact continually and where their respective magnitude and occurrence vary through time, where the climate changes, and where species may have unique responses to sequential disturbances, it is even difﬁcult to use the term ‘succession’ (Rowe, 1983). A more reasonable suggestion is that species respond functionally with certain survivalstrategies simply to cope with the disturbances. This continually interactive adjustment of the ecosystem components also involves an evolutionary impact on the ever-changing, complex, boreal biological systems (Rowe, 1983; van der Maarel 1993). Future dynamics are thus neither a straightforward process, nor easily predicted, although retrospective analyses of past ecosystem dynamics can provide indications of what future changes may be like (Willis et al., 1997; Flannigan et al., 1998).

ACKNOWLEDGEMENTS

I am grateful to the editor, Dr Lawrence Walker, for inviting me to contribute and for inspiring advice, and to the Editor-in-Chief Dr David Goodall for ﬁnal

175

reﬁnements. I thank Dr Yves Bergeron, Dr Mike Flannigan, Dr Annika Hofgaard and Dr Christer Nilsson for discussions, and two anonymous referees for pertinent comments. Elisabet Carlborg and Anders Hedefalk helped with literature search. Financial support was provided by the Swedish Council for Forestry and Agricultural Research.

Appendix A. SOUTHERN BOREAL FOREST CASE STUDY. LAKE DUPARQUET RESEARCH AND TEACHING FOREST, QUEBEC, CANADA

The Lake Duparquet Research and Teaching Forest (9300 ha; 48º30 N, 79º25 E) is situated in the Abitibi region in western Quebec, about 300 km south of James Bay. Dominant trees include Abies balsamea, Betula papyrifera, Picea glauca, P. mariana, and Populus tremuloides. Xeric sites (bedrock) are dominated by A. balsamea, P. mariana, Pinus banksiana, and Thuja occidentalis; Pinus resinosa and P. strobus can also be found. On xeric to mesic sites (moraine) A. balsamea, B. papyrifera, P. glauca, and P. mariana predominate, whilst mesic sites (clay) are dominated by A. balsamea, B. papyrifera and P. glauca and T. occidentalis. On bogs and hydric sites Larix laricina, can also be found (Bergeron and Dubuc, 1989). Fires have been recorded since 1593, and the ﬁre occurrence is depicted in Fig. 6.7. The ﬁre occurrence decreased from about 1870, when the ﬁre cycle was signiﬁcantly prolonged from 63 to 99 years. This seems to be related to a lower frequency of drought periods, in turn creating a generally more humid situation and thereby a lower ﬁre hazard (Bergeron, 1991; Bergeron and Archambault, 1993). Post-ﬁre tree regeneration follows different patterns for different sites, but is also related to disturbance regime. Post-ﬁre tree regeneration occurred on burnt sites (Figs. 6.8 and 6.9) from ﬁres in 1760, 1797, 1823, 1870, 1907, 1919, and 1923. Regeneration also occurred on nine other types of sites, including ﬂooded land and shallow organic soils over bedrock. Betula papyrifera and Pinus banksiana were the most common indicators in the post-ﬁre cohorts (Dansereau and Bergeron, 1993). Regarding post-ﬁre dynamics of understorey species, Table 6.2 (De Grandpr´e et al., 1993) presents changes among 33 taxa in a chronosequence from 26 to 230 years of post-ﬁre stand age, in the Duparquet area.

176

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Fig. 6.7. Fire occurrence in the Lake Duparquet Research and Teaching Forest, Quebec. Major ﬁres that have burnt >1 km of the lakeshore are indicated by an asterisk (from Bergeron, 1991).

Fig. 6.8. Lake Duparquet map showing major ﬁres from 1760 (white area) and 1923 (shaded area). Dots indicate plots where post-ﬁre recruitment was identiﬁed, while triangles show sites where ﬁres have been dated by ﬁre scars. When 1760 is in parentheses it indicates that the area burnt again in 1923 (from Dansereau and Bergeron, 1993). See Fig. 6.9 for an enlargement.

BOREAL FOREST DISTURBANCES

177

Fig. 6.9. An enlargement of Fig. 6.8 showing smaller ﬁres that were recorded (from Dansereau and Bergeron, 1993).

Appendix B. NORTHERN BOREAL FOREST CASE STUDY. MUDDUS NATIONAL PARK, NORTHERN SWEDEN

Muddus National Park (57 000 ha; 66º55 N, 20º15 E) is situated above the Arctic Circle. The altitudinal range is from 170 to 660 m, and the forest limit is slightly above 600 m. About 53% of the park consists of forest, 45% of mires and 2% of lakes. The principal type of solid rock is granite. A special type of till, ‘Muddus till’ (sandy gravel) predominates throughout the park, although it is sometimes interspersed with glacioﬂuvial material. The tree species distribution in the park is as follows: mixed deciduous/coniferous forest, 45%;

pine (Pinus sylvestris) forest, 21 %; spruce (Picea abies) forest, 18%; mixed coniferous forest, 13%; and pure deciduous forest, 3%. The vegetation varies along two gradients, one from dry to wet and the other from nutrient-poor to nutrientrich. These two gradients are also, however, closely related. The vegetation has here been grouped in forest types named after the mature stand in this boreal part, excluding the subalpine region (see also Uggla, 1958; Arnborg, 1990; Engelmark, 1987, 1993a). A. The pine/lichen type is a xeric forest type with stands of scattered pines. The ground is covered by mats of Cladina, Cladonia, and Stereocaulon species. The type exists on dry, well-drained, often sandy soils and is characterized by a high ﬁre frequency.

178

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Table 6.2 Frequency (F) 1 and cover (C) 2 of the most frequent understory species in the Lake Duparquet area for each post-ﬁre age 3 Species

Time since ﬁre 26 yr

46 yr F

C

74 yr

F

C

F

C

Acer spicatum

++

30.9

++

47.8

++

Aralia nudicaulis

+−

8.9

++

19.4

++

Aster macrophyllus

++

25.4

26.1

++

56.8

+−

7.0

23.2

+−

2.7

−−

2.1

++

9.5

++

31.6

++

35.2

7.0

−−

0.2

Brachythecium spp.

+−

0.2

++

1.8

++

1.2

Carex spp.

++

2.5

−−

0.1

−−

0.1

Cinna latifolia

++

2.6

−−

0.1

Circaea alpina ++

11.0

Cornus canadensis Corylus cornuta

++

31.0

Dicranum polysetum

143 yr

C

Athyrium ﬁlix-femina +−

Clintonia borealis

120 yr F

+−

1.8

−−

0.6

++

1.3

−−

0.1

−−

0.2

F

++

C

1.1

167 yr F

174 yr

C

F

++

28.9

++

13.0

230 yr

C

F

C

+−

7.0

+−

12.6

++

17.6

++

4.7

++

17.4

−−

0.1

+−

2.8

−−

0.1

++

0.7

++

0.4

++

1.5

+−

0.3

−−

0.1

+−

0.7

++

9.6

+−

2.7

+−

3.0

−−

0.1

++

5.9

+−

3.1

+−

4.7

+−

0.4

+−

0.4

−−

0.1

−−

0.2

+−

2.0

+−

0.9

−−

0.1

+−

4.0

+−

2.4

+−

3.1

−−

0.1

+−

1.1

+−

2.7

−−

0.1

−−

0.1

−−

0.1

+−

0.1

+−

0.1

−−

0.1

+−

0.1

+−

0.3

+−

0.1

+−

0.2

Dicranum spp.

−−

0.1

+−

0.3

+−

0.1

+−

0.1

++

0.4

+−

0.2

+−

0.1

Diervilla lonicera

+−

3.1

+−

4.7

+−

0.5

−−

0.1

−−

1.6

+−

2.1

−−

0.2

+−

0.2

+−

0.2

+−

0.2

+−

0.1

+−

0.1

Drepanocladus spp. Dryopteris disjuncta

−−

0.1

−−

0.1

−−

0.1

−−

3.2

−−

0.1

+−

1.3

Dryopteris spinulosa

+−

3.5

−−

0.1

−−

0.1

−−

0.9

−−

0.1

+−

0.2

−−

0.1

Galium triﬂorum

++

1.9

+−

0.4

+−

0.3

+−

0.4

+−

0.1

+−

0.5

−−

0.1

+−

0.2

−−

0.1

−−

0.1

+−

0.2

+−

0.2

−−

0.2

−−

0.2

Goodyera spp. Linnaea borealis

+−

0.3

++

2.1

−−

0.2

−−

0.7

++

2.4

++

1.9

+−

1.3

Lonicera canadensis

+−

0.3

+−

2.7

+−

1.7

+−

0.9

−−

0.2

+−

0.9

+−

1.9

+−

0.9

Maianthemum canadense

+−

0.8

++

4.3

+−

0.6

+−

0.4

−−

0.1

++

1.7

++

1.7

+−

0.1

Mitella nuda

++

3.8

Pleurozium schreberi

+−

2.0

++

4.0

++

2.3

−−

0.4

++

2.6

+−

0.4

++

1.7

−−

0.1

+−

0.2

+−

0.1

++

0.9

+−

0.4

+−

0.1

+−

0.4

Pteridium aquilinum

−−

2.3

−+

6.1

+−

4.7

Ribes glandulosum

++

4.2

+−

0.6

+−

0.1

−−

0.7

−−

0.1

+−

0.3

−−

0.1

−−

0.1

Ribes lacustre

−−

0.1

−−

0.2

+−

2.2

−−

1.0

−−

0.3

+−

0.1

−−

0.1

−−

0.1

Rosa acicularis

−−

0.4

+−

0.3

+−

0.7

−−

0.1

−−

0.1

−−

0.2

+−

2.1

Rubus idaeus

++

11.5

Rubus pubescens

++

5.8

+−

1.0

++

3.0

−−

0.7

−−

0.2

++

3.0

+−

0.8

Solidago rugosa

+−

2.6 −−

1.0

−−

2.6

++

45.2

+−

9.6

−−

1.7

−−

Trientalis borealis

−−

0.4

+−

1.0

+−

0.8

+−

0.8

−−

0.2

+−

0.7

+−

Viola spp.

++

2.8

++

1.3

++

2.0

+−

1.9

−−

0.2

++

2.9

+−

0.7

Taxus canadensis

1

−−

0.1

−−

0.1

0.1

++

57.9

0.2

+−

0.3

++

1.5

Frequency in 100 m2 quadrats and 1 m2 quadrats; ++, species frequent on the whole site (100 m2 quadrat scale) and also frequent at the 1 m2 quadrat scale; +−, species frequent on the whole site but infrequent at the 1 m2 quadrat scale; −+, species not frequent on the whole site but frequent at the 1 m2 quadrat scale; −−, species infrequent to rare at both scales. 2 Cover (%) in 100 m2 quadrats. 3 From De Grandpr´ e et al. (1993).

BOREAL FOREST DISTURBANCES

Fig. 6.10. Forest ﬁres in Muddus National Park, northern Sweden (from Engelmark, 1984).

179

180

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Fig. 6.11. Sample from a live Pinus sylvestris tree in Muddus National Park, which germinated in 1490 and thereafter has registered ﬁve ﬁre years as consecutive ﬁre scars (from Engelmark, 1984).

B. The pine/lichen/dwarf shrub type is not as xeric as the pine/lichen type. Arctostaphylos uva-ursi, Calluna vulgaris, Empetrum hermaphroditum, and Vaccinium vitis-idaea are common ﬁeld-layer species, and also Betula pendula occurs in the tree layer. This pine-forest type is more or less dependent on repeated ﬁres. If ﬁres do not occur, the shade-tolerant spruce can regenerate in the closed pine stands, and in the long run the forest can change towards mesic conditions. C. The spruce/Vaccinium type is a mesic forest type (spruce, Betula pubescens and occasionally solitary pines). Juniperus communis occurs as a shrub. The ﬁeld layer is dominated by Vaccinium myrtillus, some V. vitis-idaea, but also Deschampsia ﬂexuosa. Characteristic herbs are Cornus suecica, Linnea borealis, Luzula pilosa, Lycopodium annotinum, Melampyrum pratense, M. sylvaticum, Solidago virgaurea, and Trientalis europea. Common mosses are Dicranum spp., Hylocomium splendens, and Pleurozium schreberi. Natural ﬁres occur with low frequency. On slightly drier soils, the ﬁeld layer may be dominated by Empetrum hermaphroditum. D. The spruce/Vaccinium type is found on soils with good water supply and slightly more nutrient-rich. The same ﬁeld and bottom layer species are found as in the spruce/Vaccinium type, but also Gymnocarpium dryopteris, Listera cordata, Maianthemum bifolium,

Oxalis acetosella, Ptilium crista-castrensis, and Rubus saxatilis can be found. This type rarely burns. E. The spruce/herb type is found on even more nutrient-rich soils. Deciduous trees such as Alnus incana, Betula pubescens, Salix caprea, and Sorbus aucuparia are common. In addition to ﬁeld-layer species found in the other forest types, the following species are found: Aconitum septentrionale, Calamagrostis purpurea, Cirsium helenoides, Deschampsia cespitosa, Dryopteris expansa, Filipendula ulmaria, Geranium sylvaticum, Gnaphalium norvegicum, Lactuca alpina, Matteuccia struthiopteris, Milium effusum, Thelypteris phegopteris, and Valeriana sambucifolia, with Rhytidiadelphus triquetrus in the bottom layer. Forests of this type burn exceptionally seldom. F. Mire forest may occur in ﬂat terrain. Stagnant ground water results in slow turnover of nutrients and in peat accumulation. Dense cover by Polytrichum and Sphagnum mosses dominates the bottom layer. The tree layer may be composed of pine and/or spruce and/or Betula pubescens. Common ﬁeld-layer species are Equisetum sylvaticum, Ledum palustre, Rubus chamaemorus and Vaccinium uliginosum. Such forests may occasionally burn during dry conditions. Traces of past ﬁres in Muddus National Park were found in all types (A, B, C, D and F) except the spruce/herb type (E). About 75% of the stands studied in the park have burned at one time or another. By

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181

Fig. 6.12. Age–structure diagrams for Picea abies and Pinus sylvestris, sampled from an area of 3500 km2 including Muddus National Park and surrounding nature reserves. See original reference for discussion on changes in ﬁre regime and regeneration patterns (from Engelmark et al., 1994).

182

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Chapter 7

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS Sara L. WEBB

INTRODUCTION

Forests everywhere are subject to the force of the wind. This chapter probes the community-level consequences in temperate-zone forests for winds strong enough to uproot or snap trees. For windstorms, the major ﬁnding that emerges from this literature review is a surprising degree of variation between forests in their structural and compositional responses to windstorm disturbances. Thus generalizations are elusive, and predictions about succession and diversity are difﬁcult to draw. For tropical forests, Whigham et al. (Chapter 8, this volume) describe an equally broad range of windstorm-induced dynamics. A full understanding of windstorms as ecological disturbance events is hampered by the global loss of intact temperate forests in the face of agriculture and timber harvesting; by salvage logging of windthrown timber; and by the uneven geographic scope of windstorm studies. Most research on temperate-zone windstorms has focused on eastern North America (Table 7.1). Studies in other regions provide key insights and challenge the existing body of theory about disturbance. Some regions such as western Europe have virtually no unmanaged forests remaining; here, heavy wind damage to plantations provides insight into return intervals, but not into responses and consequences for native forests. In eastern North America, oldgrowth forest fragments are scarce (Davis, 1996) and might not be representative of the extensive forests now gone, because atypical features such as wet soil or steep slopes often explain their escape from logging. Meanwhile nearly every forest preserve in eastern North America has been struck by catastrophic windstorms within recent decades, demonstrating the important role of wind and the need for extensive

nature reserves so that the full range of disturbance responses are preserved (White, 1987). This chapter ﬁrst provides theoretical background on how disturbance by windstorms might inﬂuence community attributes and processes, and identiﬁes alternative pathways that enrich, reduce, or maintain diversity. The second section describes windstorm disturbance regimes within temperate-zone forests, discussing major storm types and their geography over time. The third section examines mortality and damage imposed by windstorm events, particularly storms with differential impacts on different species. The ﬁnal section considers regeneration opportunities provided by light gaps, microsites (such as windthrow mounds), and below-ground resource pulses. This section also highlights the importance of prestorm understory structure for windstorm consequences and brieﬂy treats questions about seed ecology and the responses of nonwoody plants. For more background and related literature reviews, the reader is referred to other chapters in this volume (especially Whigham et al., Chapter 8, and Pickett et al., Chapter 31), and to several other key information sources: the seminal book, “Ecology of Natural Disturbance and Patch Dynamics” by Pickett and White (1985), including several chapters on temperate-zone forests (Canham and Marks, 1985; Collins et al., 1985; Runkle, 1985; Veblen, 1985b); a review of forest uprooting by Schaetzl et al. (1989); ecological chapters by Everham (1995), Foster and Boose (1995), and Wooldridge et al. (1995) in the Coutts and Grace (1995) volume; and recent books on vegetation dynamics that incorporate new awareness of disturbance ecology (Burrows, 1990; Oliver and Larson, 1990; Glenn-Lewin et al., 1992). Another larger body of literature explores light

187

188

Sara L. WEBB

Table 7.1 List of studies of windstorms and gaps, by geographic locality and project type Location

Forest type

Study type

References

Asia Taiwan

plantations

hurricane

Chen and Horng (1993)

China

Abies–Betula

gaps

Taylor and Zisheng (1988)

Japan

cold-temperate Fagus

gaps

Hara (1983, 1985, 1987); Nakashizuka (1984); Watanabe et al. (1985); Yamamoto (1989);

Japan

warm-temperate hardwoods

gaps

Naka (1982); Nakashizuka (1991); Nakashizuka et al. (1992); Yamamoto (1992); Abe et al. (1995)

Japan

subalpine conifer

gaps

Yamamoto (1993, 1995)

India

subtropical broadleaf

gaps

Barik et al. (1992)

podocarp

gaps and microsites

Gibson and Brown (1991)

South Island/Canterbury

Nothofagus

blowdowns

Jane (1986)

South Island

Nothofagus

gaps and microsites

Stewart and Rose (1990); Stewart et al. (1991)

South Island/Fiordland

Nothofagus and other hardwoods

gaps and disturbance

Stewart (1986)

New Zealand

podocarp

microsites

Adams and Norton (1991)

Argentina

Nothofagus

gaps

Rebertus and Veblen (1993b); Veblen (1989b)

Chile

rainforest

gaps

Veblen (1985a,b)

Chile

xeric forests

gaps

Veblen (1989a)

Chile

Nothofagus, old growth

gaps

Veblen (1989b); Armesto et al. (1992)

Picea, old growth

gaps

Qinghong and Hytteborn (1991); Leemans (1991)

Australia Tasmania New Zealand

South America

Europe Sweden Sweden

Picea (boreal)

gaps and microsites

Hytteborn et al. (1991); Hofgaard (1993)

Sweden

Picea (boreal)

microsites, disturbance regime

Dynesius and Jonsson (1991); Jonsson and Dynesius (1993)

Finland

Picea and Betula (boreal)

wind damage

Kauhanen (1991); Mansikkaniemi and Laitenen (1991)

Poland

Picea

microsites (uprooting)

Falinski (1978)

Germany

Picea

gaps

Fischer (1992)

Germany

Picea and Fagus

ants in windfalls

Theobald-Ley and Horstmann (1990)

Germany

several

microsites (uprooting)

Heinrich (1991); Semmel (1993)

Germany

plantations

wind damage

Hutte (1968)

Ireland

plantations

wind damage

O’Cinn´eide (1975); Gallagher (1975)

Scotland

plantations

wind damage

Milne (1992) continued on next page

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

189

Table 7.1, continued Location

Forest type

Study type

References

Great Britain

plantations

wind damage

Gloyne (1968)

Ukraine

Picea

wind damage

Kalinin (1991)

Ural Mountains

coniferous

disturbance regime

Shiyatov (1990)

France

plantations (conifers)

wind damage

De Seze (1987); Despres (1987); Becquey and Riou-Nivert (1987)

Slovakia

coniferous/deciduous

wind damage

Konopka (1985)

Canada Quebec

northern hardwoods

gaps

Payette et al. (1990)

British Columbia

subalpine

gaps

Lertzman and Krebs (1991); Lertzman (1992)

Pinus–Quercus

hurricane damage

Arriaga (1988)

conifers: Tsuga/Picea

wind damage

Harris (1989) Hunter and Parker (1993)

Mexico Baja, Mexico United States Alaska (southeast) California

Conifers: Abies/Sequoia

gaps and disturbance regime

Colorado

Conifers: Picea–Abies

gaps

Veblen (1986); Aplet et al. (1988)

Florida

southern hardwoods

gaps

Canham et al. (1990)

Illinois

hardwoods

herbaceous plants and gaps

Thompson (1980)

Indiana

hardwoods

gaps

Williamson (1975)

Kentucky

mixed hardwoods

tornado regeneration

Held and Bryant (1989)

Maine

Pinus–hardwoods

gaps

Kimball et al. (1995)

Maine

northern conifers

disturbance history

Lorimer (1977)

Massachusetts

northern hardwoods

experimental gaps

Sipe and Bazzaz (1994, 1995)

Massachusetts

northern hardwoods

hurricane and other disturbances

Spurr (1956); Oliver and Stephens (1977); Hibbs (1982, 1983); Kelty (1986); Foster (1988b, 1995); Foster et al. (1992); Foster and Boose (1995)

Michigan

northern hardwoods

disturbance regime, gaps

Frelich and Martin (1988); Frelich and Lorimer (1991)

Michigan (central)

Pinus banksiana, Thuja occidentalis

gaps

Abrams and Scott (1989)

Michigan (central)

Tsuga–hardwoods

belowground gaps

Mladenoff (1987)

Michigan (central)

Pinus spp.

disturbance history

Whitney (1986)

Michigan (southwest)

hardwoods (Warren Woods)

gaps

Brewer and Merritt (1978); Poulson and Platt (1989, 1996)

Minnesota

hardwoods

gaps

Bray (1956)

Minnesota

Pinus–Abies and Pinus–hardwoods

thunderstorm damage, microsites

Webb (1988, 1989)

New Hampshire

northern hardwoods

hurricane and other disturbances

Henry and Swan (1974); Foster (1988a); Merrens and Peart (1992)

New Hampshire

Picea–Abies

gaps

Worrall and Harrington (1988)

New Hampshire

northern hardwoods

gaps

McClure and Lee (1993) continued on next page

190

Sara L. WEBB

Table 7.1, continued Location

Forest type

Study type

References Matlack et al. (1993)

New Jersey

Pinus–Quercus

thunderstorm damage

New Jersey

hardwoods

gaps

Ehrenfeld (1980)

New Jersey

Pinus rigida barrens

below-ground gaps

Ehrenfeld et al. (1995)

New York (central)

northern hardwoods

microsites and seed pool

Beatty (1984, 1991); Beatty and Sholes (1988)

New York (north)

northern hardwoods

gaps

Canham (1985, 1990)

New York (west)

northern hardwoods; conifers

disturbance history

Seischab and Orwig (1991)

Ohio

hardwoods

gaps

Runkle (1990b); Canham et al. (1990); Cho and Boerner (1991); Kupfer and Runkle (1996)

Oregon and Washington

old growth conifers

microsites

Minore (1972); Christy and Mack (1984); Harmon and Franklin (1989)

Oregon and Washington

old growth and successional conifers

gaps

Spies and Franklin (1989); Spies et al. (1990a); Canham et al. (1990)

Pennsylvania

northern hardwoods

gaps, herbaceous plants

Collins and Pickett (1987, 1988a,b)

Pennsylvania

northern hardwoods

gaps

Orwig and Abrams (1994)

Pennsylvania

northern hardwoods

tornado damage, microsites

Peterson and Pickett (1990, 1991, 1995)

S. Appalachians

hardwoods and Picea–Abies

disturbance regime (esp. ﬁre)

Harmon et al. (1983)

S. Appalachians (NC and TN)

hardwoods/cove forests

gaps

Lorimer (1980); Barden (1980, 1981); Runkle and Yetter (1987); Clebsch and Busing (1989); Clinton et al. (1994)

S. Appalachians (NC)

Picea–Abies (subalpine)

gaps

White et al. (1985a,b); Foster and Reiners (1986); Canham et al. (1990); Pauley and Clebsch (1990)

S. Appalachians (TN)

Tsuga–hardwoods

gaps

Barden (1979)

South Carolina

upland Pinus–hardwoods

hurricane damage

Gresham et al. (1991)

South Carolina

coastal swamps, ﬂoodplains

hurricane damage

Gresham et al. (1991) Sharitz et al. (1992, 1994); Duever and McCollom (1993); Allen et al. (1997); Battaglia et al. (1999)

Texas

Pinus–Quercus and mixed hardwoods

tornado regeneration

Glitzenstein and Harcombe (1988)

Vermont

montane and subalpine conifers

gaps

Perkins et al. (1992)

Virginia

mixed hardwoods

gaps

Orwig and Abrams (1994)

Wisconsin

northern hardwoods

blowdown damage

Stearns (1949); Dunn et al. (1983); Canham and Loucks (1984)

Wyoming

Pinus contorta forests

below-ground gaps

Parsons et al. (1994a,b)

gaps, structural openings within a forested matrix that are sometimes caused by wind-throw of canopy trees. This literature must be interpreted with caution as one seeks an understanding of windstorm-induced dynamics; light gaps are not all caused by windstorms, and forest turnover does not always involve discrete gaps (Lieberman et al., 1989). Another concern is that few gap studies include non-gap control areas.

A special feature in the journal Ecology (Platt and Strong, 1989) included 13 articles on tree-fall gaps; another set of articles on gaps appeared in the Canadian Journal of Forest Research (20: 617–667; Denslow and Spies, 1990). Caldwell and Pearcy (1994) have reviewed ecophysiological responses by plants to environmental heterogeneity, with several chapters focussing on responses to gaps. Table 7.1 catalogs

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

references by geographic locality and distinguishes gap-only studies from windstorm studies. THEORETICAL FRAMEWORK

Windstorms play an important role in most temperatezone forests at several spatial scales. However, this role is complex and not always in accord with predictions from theory. As ecologists began to acknowledge natural disturbances and the dynamic nature of forests, they moved away from equilibrial, deterministic successional models and from the notion of the stable climax forest. Meanwhile, paleoecology revealed how unstable forest assemblages have been over all time scales (Davis, 1981a, 1986; Foster et al., 1990), as a result of continual climatic change. Here I consider how recent views on disturbance and patch dynamics mesh with empirical observations of windstorm-driven forest dynamics. Disturbance Windstorms that damage trees ﬁt into ecological deﬁnitions of disturbance provided elsewhere in this book: “a relatively discrete event causing a change in the physical structure of the environment” (Whigham et al., Chapter 8, this volume) or, similarly “a force that abruptly disrupts the physical or biological structure of an ecological system” (Pickett et al., Chapter 31, this volume). Like other types of disturbance discussed in this book, a windstorm has consequences that depend on three interrelated aspects of the disturbance event: magnitude, intensity, and severity. In addition, to move beyond single windstorm events to an understanding of the disturbance regime over time requires knowledge of windstorm frequencies (return times) across the range of magnitude, intensity, and severity. Magnitude Magnitude is the areal extent of wind damage, ranging from individual tree-falls to blow-downs that cover many hectares. The size of the forest opening, if any, follows from characteristics of both the storm and the forest itself, with pre-storm structure and wood strength of the species present both important except in storms of the very highest-intensity. Windstorm magnitude will determine the scale and grain of the forest mosaic. Intensity Intensity is the physical force of the windstorm,

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measured primarily by wind-speed but also inﬂuenced by storm duration and gustiness. The strongest winds characterize tornadoes, whereas hurricanes exhibit lesser wind-speeds (but often spawn tornadoes); thunderstorm systems can generate winds as destructive as tornadoes, but more often produce winds of only moderate intensity. Severity Severity is the degree of change initiated by the windstorm in several components of the forest ecosystem. For windstorms, severity is not equivalent to intensity (Peterson and Pickett, 1991). The consequences of windstorms result not only from windstorm characteristics but also from characteristics of the forest itself: its location, and the life histories of its trees. Two equivalent storms can produce very different dynamics in two forests that differ in such key characteristics as understory development, canopy evenness, windﬁrmness of canopy and understory trees, wood-rotting fungal pathogens, or resprouting potential (Glitzenstein and Harcombe, 1988; Webb, 1989; Putz and Sharitz, 1991). Furthermore, two similar forests can diverge in windstorm consequences on the basis of latitude (which affects the angle of light into a gap; Canham et al., 1990) or topography (which can affect wind-speeds as well as root depth; Foster and Boose, 1992, 1995). Return time Return time, or frequency of disturbance events over time, is the key to understanding forest turnover times and dynamics at the landscape scale; this component of the disturbance regime is discussed later. Patch dynamics, neighborhood effects, and elaborations “Patch dynamics” is a model that incorporates ecological disturbance by viewing the forest as a mosaic of disturbance-generated patches. When one patch is disturbed, its composition and/or structure is modiﬁed as it enters the gap phase, while surrounding patches are in various stages of succession toward the mature phase (Watt, 1947; Bray, 1956). The patchwork can form at a landscape scale (with massive blow-downs for example: Pickett et al., Chapter 31, this volume) or on a ﬁner scale (with smaller tree-falls). In forests, the most conspicuous patchwork is a shifting mosaic that follows light-gap dynamics (Watt 1947; Bray, 1956), following disturbance or death of canopy trees; but ﬁner-grained mosaics can be overlain on this mosaic of light, as in

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the case of microsites such as wind-throw mounds and rotting logs. The structural diversity within this patchwork is generally expected to promote species diversity, but this prediction will not always be correct. Diversity enrichment takes place only if certain conditions are met. First, the disturbance must rearrange the conﬁguration of limiting resources (light, moisture, nutrients, germination sites) within the mosaic to produce a signiﬁcantly heterogeneous pattern of resource availability. Second, if the patch is large and discrete enough, then the disturbed patch will admit species that require higher levels of resources than the closed forest provides. For example, a large light gap might admit shade-intolerant taxa (such as Betula spp.) into a matrix of undisturbed shade-tolerant species where they are otherwise excluded. The ﬁnal requirement for diversity enrichment involves the pattern of damage and mortality. If the heaviest damage is to shadetolerant trees, then the result of the windstorm can be an enrichment of diversity as the shade-intolerant species are admitted into the disturbed patches. In this case, succession is set back within these patches and diversity is enriched at the scale of the multiple-patch mosaic. However, if instead the same species that regenerate in disturbed patches are most heavily damaged, the result of the windstorm might be compositional constancy – the case of positive neighborhood effects from the perspective of Frelich and Reich (1995). The neighborhood-effects model relates the fate of a disturbed forest patch to the nature and strength of the overstory/understory relationship. Self-replacement within a species (positive neighborhood effect) is common for some forest types, but rare in others. Where a positive neighborhood-type effect is strong, as in Tsuga/Fagus/Acer forests of North America, a modest windstorm will cause little compositional change because wind-thrown trees are often replaced by trees of the same species. A third possible outcome of disturbance, often overlooked but, as will be seen, quite common for windstorms, is a loss of diversity (Webb, 1989; Battaglia et al., 1999). This outcome results when shade-intolerant species sustain heavy mortality and/or are unable to colonize disturbed patches because of a pre-established understory of shade-tolerant species. Succession in its classic sense is thus accelerated by the disturbance in this case, in direct contrast with the usual prediction that succession will be set

Sara L. WEBB

back in disturbed patches. Thus, a forest operating through patch dynamics may or may not be enriched by disturbance. To accommodate the complexity of windstorm consequences, one needs to consider these three possible outcomes at the scale of the multiple-patch mosaic: (1) diversity enrichment, where the disturbance causes discrete and sizable pulses of limiting resources in disturbed patches, and where mortality is proportionally high for tolerant species; (2) compositional maintenance, where neighborhood effects are positive; (3) loss of diversity, where there is little or no pulse in limiting resources and/or where shade-intolerant taxa are the most heavily damaged. One must furthermore recognize the several interacting factors that inﬂuence windstorm consequences. These factors are explored in this chapter; they include features of the storm, of the site, and of the life histories of the species present in the system. The storm and the site together inﬂuence patch size and the magnitude of pulses of limiting resources such as light, whose elevation can admit shade-intolerant species into disturbed patches. Several key life-history features inﬂuence both regeneration in these patches and mortality patterns within and between species. For regeneration, these include tolerance of resource conﬁgurations in new patches, tolerance of competition for these resources, resistance to prevailing levels of herbivory in the understory, and seed ecology including biotic interactions with seed predators and seed dispersers. Together these factors – storm, site, and life history – structure the biotic response to the formation of a gap. For mortality, risks are related to tree size and to wind-ﬁrmness, itself a function of wood strength and fungal infections; but mortality risk also relates to site characteristics such as topography and soils. These mortality patterns must be superimposed on the regeneration continua, linking for example the wind-ﬁrmness and shade tolerance of major tree species, for a wide spectrum of possible consequences. Clearly, forests exhibit a wide range of responses to windstorms, and only some cases ﬁt the diversityenriching patch-dynamics model, which is perhaps too widely assumed to describe disturbance responses. DISTURBANCE REGIMES

Temperate-zone forests vary greatly in their disturbance regimes, which in addition to windstorms include

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

disturbance by ﬁre (Heinselman, 1973, 1981; Frissell, 1973; Wright and Heinselman, 1973; Harmon et al., 1983; Clark, 1988, 1989b), insect outbreaks (Collins, 1961; Ehrenfeld, 1980; Runkle, 1990a; see also Schowalter and Lowman, Chapter 9, this volume), pathogen epidemics (Davis, 1981a; Allison et al., 1986; Foster, 1988a) and anthropogenic forest clearance and management (Binkley, Chapter 18; Hartshorn, Chapter 19 and Barrow, Chapter 28, this volume). The overall disturbance regime for a given forest incorporates these multiple forces which interact in complex ways (Bormann and Likens, 1979; White, 1979; Foster, 1988a; Frelich and Lorimer, 1991; Clark, 1992; Hunter and Parker, 1993; Veblen and Alaback, 1995; Frelich and Reich, 1995). Even when one considers windstorms alone, the disturbance regime is highly variable in space and in time, for reasons tied to both global climatology and local physiography. Windstorm events also range across broad continua of both intensity and magnitude (White, 1979). Storm categories such as hurricanes, tornadoes, and thunderstorms overlap in their consequences, complicating efforts to reconstruct the windstorms and windstorm disturbance regimes of the past. Reconstructing the past It is difﬁcult to reconstruct the frequency of windstorm events over ecologically meaningful time spans longer than a few decades. Spatial patterns are also obscured in today’s fragmented landscape. Valuable insights come from work with historical land surveys which indicate blow-down locations and extent (as well as ﬁre-disturbed areas) before extensive Euro-American settlement of North America: 1834–1873 in Wisconsin (Stearns, 1949; Canham and Loucks, 1984); 1793– 1827 in northern Maine (Lorimer, 1977); 1788 onwards in western New York (Seischab and Orwig, 1991); 1836–1859 in Michigan (Whitney, 1986). Other historical records also help to interpret clues on the landscape scale and cover a longer time span. By dissecting forest stands in New England, Henry and Swan (1974) and Oliver and Stephens (1977) reconstructed past disturbances dating back several hundred years. In addition, tree-fall mounds have residual effects on soil proﬁles, and thus offer clues to windstorms of the past (Lutz and Griswold, 1939; Lutz, 1940; Goodlett, 1954; Stephens, 1956; Lyford and MacLean, 1966; Dynesius and Jonsson, 1991). Unlike stumps and logs, mounds generally form only when windstorms, or

193

occasionally heavy snow and ice loads, are the cause of tree damage. More clues to past disturbance events come from detailed age–structure reconstruction using tree-ring analysis. Disturbance dates can be marked by the establishment of shade-intolerant trees and by pulses of radial growth in shade-tolerant trees (Lorimer, 1985; White et al., 1985b; Glitzenstein et al., 1986; Lorimer et al., 1988, 1992; Canham, 1989; Lorimer and Frelich, 1989; Shiyatov, 1990; Payette et al., 1990; Stewart et al., 1991; Clark, 1991; Frelich and Lorimer, 1991; Foster et al., 1992). This approach cannot distinguish windstorm events from other causes of gap formation and canopy turnover unless used in conjunction with wind-throw mounds or historical records (Foster, 1988a). Aerial photographs and other remotely-sensed images can be valuable for dating and tracking large blow-downs from recent decades (Rebertus et al., 1997). Meteorological reconstructions and models are shedding light on hurricane history in New England (Boose et al., 1994). Taken together, these methods provide a fascinating picture of the recent past, extending several centuries in the most intensively studied areas (Oliver and Stephens, 1977; Foster, 1988a; Rebertus et al., 1997). Historical records show the extent of blow-downs across survey lines. Stearns (1949) and Canham and Loucks (1984) examined notes of U.S. General Land Ofﬁce surveyors and found that blow-downs were widespread at the time of Euro-American settlement (1834–1873). Stearns found records of an extensive (64 km+ long) blow-down dating from a storm in 1872. Canham and Loucks (1984) reviewed survey records for the entire extent of hemlock (Tsuga)–northern hardwoods forests of Wisconsin. Blow-downs covered a surprisingly large portion of the landscape, with an estimated return time of 1210 years, the average time for the entire study region to be disturbed by blowdowns. This abundance of blow-downs goes beyond what tornado frequencies could explain, and the blowdown shapes are also wider than clearings formed by tornadoes. These Wisconsin results differ from other reconstructions farther east in North America, with an apparent gradient of decreasing blow-down frequency from midwest to east, but with hurricane damage becoming more important nearer the east coast. Just east of Wisconsin, before Euro-American settlement, forests in northern lower Michigan had an estimated

194

recurrence interval of 1220 years for blow-downs in upland Tsuga–Pinus strobus–hardwood forests, but with more frequent blow-downs in swamps (Whitney, 1986). Farther east in western New York State, Seischab and Orwig (1991) found geographic variance in wind-throw distribution: virtually no wind-throw in a ﬂat till-plain region, but conspicuous wind-throw coverage (0.5%) in the more dissected, steep terrain of the Allegheny plateau. Return times were much longer (3190 years) in the western half than the eastern half (980 years) of this plateau (overall average, 1720 years), perhaps because of hurricanes toward the east. However, most of these New York blow-downs were caused by thunderstorms (Seischab and Orwig, 1991). For presettlement lands in northern and western Maine, Cogbill (1996) calculated 1.4% coverage by fallen timber and a return time of 1400 years between events that induce large blowdowns. In northeastern Maine, Lorimer (1977) found few blow-downs recorded for upland forests; most wind-falls were in lowland conifer forests with shallowrooted species, with a spatial distribution that suggested damage by a single storm (possibly 1795); an overall recurrence interval was estimated at 1150 years. These return times are estimated from a single point in time. In the absence of long-term records, all estimates of return intervals must be considered extremely tentative because of their inherent assumptions about an equilibrial storm regime, blow-down persistence, and constant vulnerability to wind-throw across the landscape. Records of timber loss to windstorms provide a limited window onto the disturbance regime in Europe. Central France was struck by a windstorm in 1982 that caused extensive damage to plantations (Despres, 1987). More widespread windstorms crossed several European nations in 1990, initiating long-term research in tree-fall areas of Bavaria (Germany) (Fischer, 1992). A compilation by Gallagher (1975) documents forestdamaging windstorms almost annually in Ireland. It is unclear what the magnitude and severity of these windstorms would be in a naturally forested landscape. Unfortunately, the distant past is unknown because windstorms do not provide fossilized signatures (Foster, 1995), as wildﬁres do in the form of charcoal deposits (e.g. Swain, 1973; Clark, 1988, 1989b). Further obscuring our understanding is the certainty that windstorm disturbance regimes are nonequilibrial. Windstorms do not occur at regular intervals and do not affect all locations equally often (Jonsson and Dynesius, 1993).

Sara L. WEBB

As with ﬁre effects, windstorm responses are inevitably linked to previous disturbance events on the site (Clark, 1989a). For example, Putz and Sharitz (1991) found that trees with heart-rotting fungi from previous wind-breakage were most susceptible to damage in a subsequent hurricane. Global windstorm patterns are likely to have shifted with a naturally dynamic climate, in the same way that ﬁre frequencies have changed (Wright and Heinselman, 1973; Clark, 1989b). Given our limited window into the past, long-term studies must be initiated to monitor the occurrence of disturbance events within ﬁxed geographic areas over time. Sizable natural areas need to be protected from post-storm salvage logging for a clear understanding of wind-mediated dynamics. Hurricanes Although hurricanes are tropical storms, occasional hurricanes reach temperate zones and achieve landfall, damaging forests. Hurricanes create a mosaic of forest damage, with zones of massive blowdown surrounded by wider zones with scattered treefalls. A tropical storm becomes a hurricane (or typhoon), in the northwestern Paciﬁc Ocean) when wind-speeds reach 33 m s−1 (119 km hr−1 ; Tufty, 1987). Sustained wind-speeds typically peak around 45– 67 m s−1 (160–240 km hr−1 ) but occasionally exceed 89 m s−1 (320 km h−1 ; Battan, 1961). The meteorology of hurricanes is summarized clearly by Foster and Boose (1995). Much hurricane damage to trees is associated with wind gusts which vary by 30–50% depending on the topographic roughness (Foster and Boose, 1995). Where wind-speeds reach peak velocities, some 30 km from the storm’s center, virtually all trees may be wind-thrown. With distance from the center, and also as the storm moves inland, wind velocities decline and damage is less severe. (Merrens and Peart, 1992). Some damaged trees resprout, thus surviving even in high-intensity zones although losing canopy stature (Sharitz et al., 1994). As with storms of more moderate intensity, hurricanes have variable consequences depending upon features of the plants present: their stature, strength, health, and regeneration requirements. Among temperate forests, the hurricane disturbance regime has been best studied in New England (northeastern U.S.A.). Meteorological modelling here and in the tropical forests of Puerto Rico show some promise

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

for relating hurricane damage to topography, windspeed, and storm duration (Boose et al., 1994). One site in southwestern New Hampshire shows evidence of 13 windstorms, including at least 4 hurricanes, with local to regional impact, between 1635 and 1938 (Foster, 1988a). This amounts to one detectable windstorm per 23 years, including one hurricane per 75 years on average. New England as a whole was inﬂuenced by 6 hurricanes that caused severe forest damage between 1635 and 1944 (Boose et al., 1994), approximately one every 50 years. Farther south along the same coast, South Carolina experiences more frequent hurricane landfalls: an average of one hurricane in 5.8 years, with severe hurricanes (comparable to Hurricane Hugo in 1989; see below) every 27.5 years (based upon weather records for 110 years: Purvis, 1973; Gresham et al., 1991). A stretch of coast in eastern Texas, adjacent to Lousiana, has an estimated 14% annual probability of hurricane occurrence, a 7-year return interval (Simpson and Lawrence, 1971). In the Paciﬁc Ocean, hurricanes struck temperate forests of Baja California Sur (Mexico) three times in a recent 11-year period (Arriaga, 1988). In southwestern Japan, Naka (1982) surveyed damage and gap formation from a 1979 typhoon, and also linked other gaps in subalpine coniferous forests to previous typhoons. A typhoon struck beech (Fagus) forests of central Japan (Nagano) in 1982 (Watanabe et al., 1985). Throughout Japan, a pattern of small gaps of unspeciﬁed origin characterizes Fagus crenata forests (Yamamoto, 1989). In western Japan (Kyushu district) typhoons are responsible for extensive small gaps within species-rich evergreen broadleaved forests (Yamamoto, 1992). Temperate broad-leaved forests in northeastern Japan also appear to be free of extensive blow-downs, but instead contain a relatively large areal coverage of small tree-fall gaps; the role of windstorms in their formation is unknown (Nakashizuka, 1984). Individual hurricanes and their consequences have been scrutinized in several temperate-zone localities; results are summarized later in this chapter. One major hurricane in 1938 (estimated landfall winds at 25 m s−1 ) caused extensive forest damage in the northeastern United States. Subsequent forest dynamics were examined in central Massachusetts (Spurr, 1956; Kelty, 1986; Foster, 1988a), in southern New Hampshire (Foster, 1988b), and farther north in New Hampshire where damage was less intense, with a maximum of 80% and more frequently only ~20% canopy mortality

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(Hubbard Brook Experimental Forest: Merrens and Peart, 1992; Peart et al., 1992). This same 1938 hurricane extended as far north as New Brunswick, Canada (Lyford and MacLean, 1966). More recently in 1989, Hurricane Hugo struck forests in South Carolina with sustained winds of 62 m s−1 (222 km hr−1 ) (Gresham et al., 1991), spurring studies of damage and forest responses in bottomlands and uplands (Putz and Sharitz, 1991; Gresham et al., 1991; Shefﬁeld and Thompson, 1992; Sharitz et al., 1992; Gardner et al., 1992; Shefﬁeld and Thompson, 1992; Duever and McCollom, 1993; Allen et al., 1997; see Chapter 8 for consequences of the same hurricane in Puerto Rico). Another well-studied hurricane was Hurricane Andrew, which crossed Florida and also struck Louisiana in 1992, with sustained winds of 242 km hr−1 (Pimm et al., 1994; Whigham et al., Chapter 8, this volume). Tornadoes Tornadoes carry the strongest winds in nature, reaching perhaps twice the wind-speeds of hurricanes. Tornado winds of 54 m s−1 (200 km hr−1 ) have been measured but typical peaks of 180–225 m s−1 (600–800 km hr−1 ) are likely, judging from the damage these storms wreak (Battan, 1961; Tufty, 1987). However, tornadoes also differ from hurricanes in the lower magnitude (areal extent) of the damage they cause. Thus, the higher wind-speeds of tornadoes inﬂuence smaller areas whereas the lower wind-speeds of hurricanes inﬂuence larger regions. This suggests that disturbances should not be classiﬁed by either magnitude or intensity alone. The highest tornado wind-speeds will entirely open the canopy, as in the case of a 1985 tornado that ﬂattened 400 ha of old-growth forest in Pennsylvania, cutting a swath 600–900 m wide (Peterson and Pickett, 1991, 1995). In other cases, forest damage is more diffuse. A Kentucky tornado created single-tree and multiple-tree gaps amidst trees left standing, although many survivors later died (Held and Bryant, 1989). A Texas tornado followed a storm track 42 km long, but touched down only sporadically, creating two discrete blow-downs measuring 31.1 ha and 10.4 ha (Glitzenstein and Harcombe, 1988). Even in these blow-downs, mortality was not total. Some 35% of canopy trees escaped serious damage; canopy coverage was reduced by 83% and 46% in the two blow-downs, respectively.

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Tornado frequency is well known to follow geographic patterns, both continentally and within regions. Trenchard (1977) calculated a national average annual frequency of tornadoes for the United States of 0.66 per 10 000 km −2 , and further estimated regional rate for eastern Texas of 5.8 tornadoes per 10 000 km−2 . Kessler and White (1983) have mapped the average annual tornado frequency by state, for 1953–1974; the highest ﬁgures were for Oklahoma (3.17 per 10 000 km−2 ), Indiana (2.39), Kansas (2.3), and Florida (2.12), followed by (in order, from 2.0 to 1.2) Massachusetts, Illinois, Iowa, Alabama, Missouri, Nebraska, Lousiana, Georgia, Mississippi, Connecticut, Arkansas, Ohio, and Wisconsin. At this broad scale, tornadoes are most frequent, particularly on the plains and on other level lands. Patterns exist on a more localized scale as well. In Arkansas, tornados are concentrated in level physiographic regions but are uncommon in mountainous areas (Gallimore and Lettau, 1970); roughness can decrease tornado wind-speeds and gusts (Dessens, 1972). However, the path of a tornado can be entirely independent of physiography (Peterson and Pickett, 1991; Foster et al., 1999). Several recent studies have scrutinized individual tornado events, often with invaluable permanent plots for monitoring postdisturbance dynamics (Dunn et al., 1983; Glitzenstein and Harcombe, 1988; Held and Bryant, 1989; Peterson et al., 1990; Peterson and Pickett, 1991, 1995; Toyooka et al., 1992; Peterson and Carson, 1996; Peterson and Rebertus, 1997). Other catastrophic windstorms Temperate-zone forests are struck not only by hurricanes and tornadoes, but also by a variety of other windstorms, from moderate to gale-force in intensity. So-called catastrophic blow-downs result from storms that clear a dozen to hundreds of canopy trees in a contiguous area. Catastrophic winds differ in both wind-speeds and ecological consequences from more moderate thunderstorm winds, which instead cause scattered tree-falls within an intact canopy (see next section). However, thunderstorms themselves often spawn bursts of catastrophic winds including tornadoes. Several studies examine blow-downs of uncertain or unspeciﬁed meteorological origin (Jane, 1986; Veblen, 1986; Rebertus et al., 1997). The down-burst is one important type of catastrophic wind that emerges from thunderstorm cells (Fujita, 1978). Such storms blow down large patches of

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forest, with wind-speeds around 35–50 m s−1 (120– 180 km hr−1 ) (Canham and Loucks, 1984). Downbursts are perhaps most similar to tornadoes in their localized occurrence and powerful winds, but downbursts typically cover much larger areas. A 1977 thunderstorm in Wisconsin that spawned 25 down-bursts along a 266-km path (Dunn et al., 1983; Canham and Loucks, 1984) provides a framework for interpreting past blow-downs as recorded in historical records. The down-bursts occurred within a storm track 27 km wide and 266 km long, where 344 000 ha of forest were leveled (Dunn et al., 1983). Within the Flambeau tract (Fig. 7.1), a virgin hardwood forest struck by this storm, 89% of stems and 94% of basal area were lost, indicating near-total canopy clearance. Stearns (1949) observed a similar severe blow-down in 1936, also in Wisconsin. Another category of catastrophic wind is the derecho event, like that which caused extensive blow-downs in 1995 in the Adirondack Park of New York State (Jenkins 1996). Moderate windstorms/thunderstorms Not all windstorms create extensive blow-downs. More moderate winds often damage scattered trees within a matrix of undamaged forest. The ecological consequences are more complex and selective from the standpoint of mortality, but less dramatic from the standpoint of resource rearrangements and regeneration. Such less catastrophic windstorms are sometimes the source of the canopy light-gaps so intensively studied in recent years, although other causes of tree mortality produce similar and often indistinguishable openings. Scattered tree-falls are found along the fringes of blow-down zones from hurricanes, tornadoes, and down-bursts, but are perhaps more commonly caused by lower-intensity windstorms. For example, the gusty winds that precede thunderstorms, called “plow winds,” can exceed 35 m s−1 (200 km hr−1 ; Battan, 1961) and create sporadic tree-falls (Webb, 1989). Thunderstorm frequencies in the United States are highest in the southeast, central plains, and northern Arizona (Kessler and White, 1983). In one Minnesota forest, summer thunderstorms and winter windstorms together represented a regular disturbance regime damaging 1.6% of all trees over a four-year period (1983–1987), two thirds (1.1%) during a single storm (Webb, 1989). Only a few studies have examined individual moderate windstorm events. Such studies are valuable for

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

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Fig. 7.1. Blow-down area seven years after the 1977 down-burst event in the old-growth Flambeau Tract in northern Wisconsin, showing heavy development of Rubus thicket. This storm and consequences have been discussed by Dunn et al. (1983) and Canham and Loucks (1984).

mortality information (Webb, 1989; Fischer, 1992; Matlack et al., 1993) and are yet more valuable if marked plots are resurveyed periodically for storm damage and also for post-windstorm regeneration. In general, little is known about the global frequency and distribution of moderate windstorms that cause treefalls. Forest turnover from gap surveys How long does it take for the natural regime of windstorm events to cause forest turnover? This question can be answered only with knowledge of disturbance return intervals and of mortality from the average storm, within a framework appropriate to the scale of interest. Although mortality patterns and percentages have been documented for some stormdamaged forests, frequencies of moderate windstorms have rarely been examined in an ecological context, making the search for a disturbance-linked turnover rate a difﬁcult one. Small gaps (40–130 m2 ) from individual tree deaths are the major sites of canopy turnover in North American and Japanese deciduous forests (Dahir and Lorimer, 1996). Many studies of gap dynamics include

attempts to estimate forest turnover rate from the area covered by gaps. Several problems plague the use of gap conﬁgurations to estimate turnover speciﬁcally caused by windstorms. One problem is that turnover in gaps omits that component of forest turnover not occurring in discrete openings – that is, it omits wind damage that does not create light gaps (Lieberman et al., 1985, 1989; Webb, 1988). The other problem is that other sources of tree mortality also produce gaps; thus, gaps cannot be attributed to wind unless speciﬁc windstorms are observed or uprooting is in evidence. Some gap studies explicitly examine results of nonwind mortality – for example, by insects and pathogens (Ehrenfeld, 1980; Huenneke, 1983; Runkle, 1990a; Pauley and Clebsch, 1990; Kimball et al., 1995), by drought (Clinton et al., 1994), and by timber harvesting (McClure and Lee, 1993; Kimball et al., 1995; see also Binkley and White, Chapter 18, this volume;). Close scrutiny of gaps along an elevational transect in New Hampshire showed that wind stress and wind-throw caused the majority of gaps and were especially important at higher elevations; but at least 20 other agents also formed gaps, with multiple factors sometimes contributing to the same gap (Worrall and Harrington, 1988). An invaluable approach is

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Table 7.2 Experimental treefall/blowdown studies in the U.S.A. Location

Forest type

Northwest Pennsylvania

northern hardwoods

Central Massachusetts

Start date

Gap sizes and numbers

References

1981

3 single-tree (33–36 m2 ); 3 multitree (51, 69, 151 m2 )

Collins and Pickett (1987, 1988a,b)

northern hardwoods

1987

3 (75 m2 ); 3 (300 m2 )

Sipe and Bazzaz (1994, 1995); Carlton and Bazzaz (1998)

Central Massachusetts

northern hardwoods

1990

0.8 ha

Boose et al. (1994); Foster et al. (1997)

Central Oregon

coniferous (Douglas-ﬁr)

1990

0–2000 m2 ; 5 gap sizes

Spies et al. (1990b); Gray and Spies (1992)

periodically to census newly formed gaps, where there is no mystery about the cause of mortality or the effects of disturbance events (Romme and Martin, 1982). Besides the problems of uncertain gap causation, other problems complicate the use of gap studies to understand forest turnover, from windstorms or other causes. Neither the time of gap formation nor the longevity of a gap in the landscape is known for certain. Both parameters can be estimated, but the ultimate turnover rate is extremely sensitive to these numbers (Stewart et al., 1991) and thus accurate predictions are elusive. Within one forest area, work on gap turnover by two different scientists produced very different estimates (Barden, 1989), apparently because of different gap deﬁnitions and sampling procedures. Citing concerns about methods of estimating turnover, Stewart et al. (1991) wisely chose not to estimate turnover but instead to initiate long-term studies. Harvest-created and experimental tree-falls Understanding of windstorms is hampered by their unpredictability and the consequent inability to characterize the pre-disturbance forest. Another challenge is that management practices include salvage logging after windstorms, even on public lands. For more controlled study of windstorm responses, several ecologists have created experimental tree-falls. Where trees are pulled over as by wind and coarse woody debris is left intact, and where control areas are also monitored, these projects can provide a more realistic look at post-disturbance regeneration than do many studies of gap dynamics. Other studies have utilized timberharvest clearings as gap proxies (McClure and Lee, 1993; Parsons et al., 1994a,b; Kimball et al., 1995), again with the advantage of control over time and

size of openings, but with differences from wind-throw gaps: woody debris is removed, no mounds or pits are created, and soils are scariﬁed and compacted by logging equipment. Table 7.2 details four projects with ecologically-oriented experimental gaps that more fully simulate windstorm effects (Collins and Pickett, 1987, 1988a,b; Spies et al., 1990b; Gray and Spies, 1992; Boose et al., 1994; Sipe and Bazzaz, 1994, 1995; Foster et al., 1997). High winds not associated with storm events Some forests are shaped by chronic wind even without major storm events (Coutts and Grace, 1995; Ennos, 1997). Steady but strong winds can inﬂuence plant physiology (Grace, 1977) and tree morphology, as in the formation of reaction wood (Wilson and Archer, 1979). High-elevation winds impose physical force and desiccation on subalpine trees, and might play a role in the dwarfed “krummholz” growth form of trees near the timber line, and also perhaps in structuring the timber line itself (Scott et al., 1993). Wind probably also contributes to the phenomenon called “ﬁr waves.” Fir (or other conifer) waves are striking patterns within subalpine conifer forests, where mortality and subsequent reproduction occur in wavelike lines that shift position over time much as a wave moves shoreward. Fir waves have been observed in mountains of the northeastern United States (Sprugel, 1976, 1984; Maloney, 1986), Argentina (Rebertus and Veblen, 1993a), and Japan (Iwaki and Totsuka, 1959; Takahaski, 1979; Kohyama and Fujita, 1981). Initially diagnosed as blow-down areas from storms, these ﬁr waves include strips of standing dead trees whose mortality is still not fully understood. The proximity of the adjacent opening seems to be a contributing factor,

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perhaps by inﬂuencing wind gustiness, pathogens, ice damage, and winter desiccation (Sprugel, 1976). Marchand et al. (1986) hypothesized that mortality results from exposure to wind and rime ice along the edge of the opening, causing foliage loss, accompanied by root damage incurred when trees sway in the wind from which they were previously sheltered. Heavy sustained winds blow on these upper mountain slopes (Rebertus and Veblen, 1993a); that wind plays a direct or contributing role is increasingly clear.

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crown volume and weight, are potentially important but rarely examined. Smaller understory trees often sustain heavy damage as well, if toppled by falling neighbors. Thus the size distribution of damage can be bimodal (Brokaw and Walker, 1991; Nakashizuka et al., 1992). Tree size can be confounded with fungal infestation (most common amongst older trees, which are often but not always the tallest) and with tree species (if species differ in size distribution). Species

TREE MORTALITY AND DAMAGE

Although much research scrutinizes regrowth in gaps, far less is known about patterns of mortality and damage caused by windstorms. Such mortality information is essential if one is to understand how disturbance events inﬂuence forest dynamics. An important review paper by Everham (1995) summarized 26 storm events reported in 51 papers, including 9 papers about temperate-zone forests. That review showed that, while species differ in vulnerability to damage, no generalizations can be made on the basis of successional class or the conifer/angiosperm dichotomy. Many factors inﬂuence mortality patterns, with great variation between stands and sites. Some canopy trees survive even the most intense (highest wind-speed) windstorms (Everham, 1995; Glitzenstein and Harcombe, 1988; Duever and McCollom, 1993; Peterson and Rebertus, 1997). Other trees will resprout. More moderate windstorms leave a matrix of undamaged trees around scattered tree-falls. Beyond mortality and turnover, windstorms cause a wide range of damage which trees survive, although not without consequences: defoliation, branch loss, and bending, especially of secondary victims of falling trees. What determines which trees will be wind-thrown? Some major risk factors are listed below; well-designed studies should tease these and other factors apart with appropriate statistical analyses. Size Taller trees are often most vulnerable (Falinski, 1978; Foster, 1988b; Kauhanen, 1991; Peterson and Pickett, 1991; Foster and Boose, 1992), particularly when the canopy structure is uneven such that tall trees protrude (Webb, 1989). Size dimensions besides height, such as

In most windstorms, some tree species are hit more heavily than others (De Seze, 1987; Webb, 1989; Matlack et al., 1993), although it can be difﬁcult to distinguish the effects of species from effects of tree size (Falinski, 1978). Most vulnerable are species with weaker wood (Webb, 1989) or shallower root systems (Lorimer, 1977; Whitney, 1986; Gresham et al., 1991; Putz and Sharitz, 1991). Species also vary in survivorship of storm damage: in ability to resprout, either from the base or higher up (Glitzenstein and Harcombe, 1988), and in ability to survive being knocked into a leaning position (Akachuka, 1993). Fungal infections Wood-rotting fungi increase the risk of wind-throw by weakening either bole or roots (Hubert, 1918; Webb, 1989; Hytteborn et al., 1991; Matlack et al., 1993). Such fungal infestations can help explain unexpected mortality patterns: heavy damage to otherwise strongwooded species, high variance in damage within a species when fungi are present only in some populations, and equal damage to small and large stems when infection is equally prevalent across size classes. Genetic variability Like fungal infections, genetic variability is a factor that can generate a range of vulnerability within species. More studies are needed like that of Silen et al. (1993), who found strong parental inﬂuence on wind-throw susceptibility for an experimental forest of Pseudotsuga menziesii. Topography Ridge-tops and storm-facing slopes are most vulnerable

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to hurricane damage (Foster and Boose, 1992, 1995) whereas tornadoes tend to follow valleys (Gallimore and Lettau, 1970); but tornado winds can become gustier over rough terrain (Dessens, 1972). Hurricane Hugo (in South Carolina) damaged more trees at higher elevations (81% damaged) than in ﬂoodplain bottomlands (44%; Duever and McCollom, 1993). A windstorm in Finnish Lappland reduced canopy cover by 85% on wind-facing slopes, by 42% on hilltops, and by smaller percentages elsewhere (Kauhanen, 1991). However, lee-slope trees were also most vulnerable in New Zealand (Jane, 1986). In Finland, an analysis of background non-storm winds showed that higher elevations and wind-facing aspects experienced 42%–86% increases in wind-speeds as compared with valleys (Mansikkaniemi and Laitenen, 1991). Similarly, Irish plantations sustained highest damage on northeast slopes facing prevailing storm winds (O’Cinn´eide, 1975). Sometimes characteristics of topography and tree species are confounded, for example when certain species are more common on ridges or in swales. Thus, the prevalence of blow-downs in low areas in Maine (Lorimer, 1977) and Michigan (Whitney, 1986) resulted from weak wood and shallow roots in the species (conifers in the genera Abies, Larix and Picea) typically found there. The possibility that wind is a ﬁltering factor in structuring topographic gradients of vegetation deserves consideration (Charles Cogbill, pers. commun.). Soils Soil type and aspect combined to inﬂuence wind-throw risk in the Ukraine (Kalinin, 1991). In west-central Germany, rooting depth and wind resistance were least on mesic fertile soils (Heinrich, 1991). However, wind-throw vulnerability did not vary across a steep gradient of soil type and depth in Canterbury, New Zealand (Jane, 1986). Topography and soils are often confounded factors, however, with shallow soils on exposed ridges and deeper soils in protected valleys. Load of climbers and epiphytes The woody vines known as lianas made host trees more susceptible to hurricane damage in a South Carolina swamp (Allen et al., 1997). This and similar ﬁndings in tropical forests suggest a more negative than neutral effect of climbers and epiphytes on their host plants (Strong, 1977).

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Silvicultural treatment When trees are thinned, remaining trees often succumb to high winds in their newly opened canopy. Foresters have documented increased wind-throw in thinned stands, especially where topographic risk factors are high and where thinned stands are also fertilized (Gloyne, 1968; Hutte, 1968; Valinger and Lundquist, 1992; Valinger et al., 1994). Early thinning may help to build wood strength, while planting at low densities initially can minimize the need for later thinning (Becquey and Riou-Nivert, 1987; Debourdieu, 1987). Complex mortality patterns Taken together, these risk factors generate a variety of mortality patterns in windstorms. When different tree species sustain different levels of damage, a shift in community composition can occur. The following examples illustrate this complexity of damage and its consequences for forest dynamics. In wet forested sloughs in South Carolina, Hurricane Hugo (at 155 km hr−1 ) imposed a mortality pattern suggesting depletion of diversity (Putz and Sharitz, 1991). The two dominant species (Nyssa aquatica and Taxodium distichum) sustained little damage while mortality was heavier for less common species, perhaps because they grew on drier, less stable microsites. As for the wind-resistant dominants, Taxodium distichum has strong wood and deep sinker roots, while Nyssa aquatica tends to lose its leaves and then form less of a target for wind. The overall damage rate of 19% of trees in the slough community was low compared with 49% of trees damaged in adjacent bottomland hardwood stands, resulting in part from heavy damage to Pinus taeda whose large crowns brought down neighboring trees (Putz and Sharitz, 1991). Uneven mortality was also seen among species in both upland and bottomland forests somewhat farther (97 km) from the eye of Hurricane Hugo (87 km hr−1 with gusts to 138 km hr−1 ; Gresham et al., 1991). Here, 24% of stems were undamaged, 49% lightly damaged, 16% moderately damaged, and 11% heavily damaged. Again, survivorship was high. Although larger trees were most susceptible, considerable variation in wind resistance between species resulted from autecological differences, including root systems, wood strength, buttress formation, and foliage retention (Gresham et al., 1991), as well as soil type and topographic afﬁnities (Gardner et al., 1992). At this

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coastal locality, the hurricane also caused extensive tree mortality indirectly by intrusion of salt water and later by infestation of weakened trees by pine bark beetles (Dendroctonus sp. and Ips sp.) (Gardner et al., 1992), illustrating the complexity and interactions of disturbance factors. Indirect effects of salt transport during hurricanes have been seen elsewhere. Differences in tolerance of salt mist explained differences in hurricane-induced mortality across species in Taiwan plantations (Chen and Horng, 1993). Sometimes the dominant tree species are the most heavily damaged by windstorms. A Texas tornado imposed heaviest mortality to the most abundant canopy species in two different vegetation types; this was primarily a size effect, species of lower canopy strata sustaining less damage (Glitzenstein and Harcombe, 1988). These same heavily-damaged canopy species also had poor reproduction, particularly a lack of resprouting. In one of two blow-downs, the mortality/regeneration proﬁle suggested little change in vegetation; in the other blow-down, heavy damage to non-resprouting pines suggested a possible decline in the pine populations and thus a decline in diversity, although the authors called for long-term study to test these predictions. These results also illustrate that similar (or the same) windstorms can have quite opposite consequences depending upon forest composition. Two Texas forests, one with pines and one without, may follow different post-disturbance trajectories (Glitzenstein and Harcombe, 1988). Two Minnesota pine forests, one with wind-ﬁrm Acer saccharum under its pines and one with an understory of weak-wooded Abies balsamea, diverge in response to the same windstorm. One forest with A. balsamea had discrete light-gaps where diversity might be maintained, while the other forest had an Acer saccharum understory which progressed into the canopy where shade-intolerant pines and Populus had blown down (Webb, 1989). These two Minnesota stands also differed in mortality proﬁles across species; older and larger Pinus resinosa and Betula papyrifera in one stand sustained higher than average mortality, while both species sustained relatively little mortality in the other stand where typical stems were younger and shorter (Webb, 1989). Another moderate windstorm struck a Quercus– Pinus forest in southern New Jersey and damaged trees of only two of the seven species present, both oaks with a high incidence of fungal rot from past ﬁre damage (Matlack et al., 1993). Neither tree size nor

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species wood strength inﬂuenced wind-throw risk in this stand. In many cases, mortality patterns will tend to advance succession, when early-successional, lightdemanding trees are most vulnerable to wind-throw (Glitzenstein and Harcombe, 1988; Webb, 1989; Battaglia et al., 1999). This pattern is common because of the higher canopy stature of these species in a successional forest, and because of typically weaker wood associated with fast growth in openings. The 1938 hurricane in New England imposed heaviest damage to Pinus strobus, advancing successional turnover to hardwoods which were in some cases already present in the understory (Foster, 1988a,b; Foster and Boose, 1992). However, some shade-tolerant trees are not wind-ﬁrm (for example Abies balsamea; Webb, 1989), complicating the search for generalizations. When shade-tolerant trees are already in the canopy, they sustain damage alongside shade-intolerant species. A Kentucky tornado imposed heaviest damage to Acer saccharum and Fraxinus americana, apparently because these were the most abundant trees in the forest (Held and Bryant, 1989). When mortality was placed in context of abundance and tree size, Acer saccharum sustained relatively low rates of mortality in Minnesota thunderstorms because of its extremely strong wood (Webb, 1989). Mortality from windstorms is probably minimal for smaller understory plants and shrubs, although studies of damage are generally lacking. For lianas (woody epiphytes), heavy mortality was documented in a South Carolina swamp community following Hurricane Hugo (Allen et al., 1997). Damaged shrubs normally maintain some undamaged stems, and the typical clonal herbs of woodlands can also persist even if a few ramets are buried in wind-throw debris. Two methodological imperatives emerge from the literature on wind mortality and damage. First, if global comparisons are to be made between storm types or forest types, methods of describing mortality and damage must be standardized (Everham, 1995). A second need is for each damage/mortality survey to include a sampling of the background forest composition, as context in which to place the damage proﬁle. In the absence of pre-storm vegetation analysis, sampling of undamaged trees within the blow-down area or in adjacent undisturbed forests is better than no contextual framework at all.

202 RESPONSES OF THE FOREST: REGENERATION OPPORTUNITIES

The community response to windstorm disturbance depends not only upon the mortality and damage proﬁles, but also upon the use of newly available resources by existing plants or new colonists. When trees are wind-thrown, a pulse of new resources becomes available, with increased light, moisture, and nutrients that were previously pre-empted by the windthrown tree. In addition, microsites on the forest ﬂoor (wind-throw mounds, pits, stumps, and logs) provide new substrates in a matrix of soil covered with litter and full of preexisting plant roots; interacting animals respond to the change in forest structure as well. Light gaps Wind damage to the canopy releases light over large tracts or in smaller patches called light gaps. The fate of light gaps has attracted much ecological attention, in forests both of the tropical zone (Whigham et al., Chapter 8, this volume) and the temperate zone (Williamson, 1975; Barden, 1979, 1981; Runkle, 1981, 1985, 1990b, 1998; Hibbs, 1982, 1983; Foster and Reiners, 1986; Worrall and Harrington, 1988; Poulson and Platt, 1989; McClure and Lee, 1993; Lertzman et al., 1996). In this section, the inﬂuence of light on windstorm responses is examined, including some gap studies not linked with wind but potentially analogous to the type of light regime produced by moderateintensity windstorms. Light-gap measurement and characteristics Light gaps can be measured in a variety of ways, using actual light levels (Canham, 1988; Canham et al., 1990), canopy photography (Canham et al., 1990; Easter and Spies, 1994), or (most commonly) physical dimensions of a structural opening (Runkle, 1981, 1992; Brokaw, 1982). In some cases the presence of shade-intolerant plants is among the criteria used to deﬁne a gap (Naka, 1982), a problematically circular procedure. Even in undisturbed forests, light has a heterogeneous distribution (Chazdon, 1988; Canham et al., 1994; Percy et al., 1994). Direct light measurements are difﬁcult to obtain with a spatial and temporal distribution sufﬁcient to characterize the entire gap area both diurnally and across the seasons. Thus, to detect gap effects adequately requires more than single-time, single-point light measurements. Existing studies that characterize light in gaps are

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not linked to windstorm events, but nevertheless offer insight into possible consequences of wind-throw for light within the forest. Collins and Pickett (1987) found light levels to be elevated only at noon and only north of gap centers (in the northern hemisphere), in experimental gaps measuring 51–151 m2 . However, in India, gaps (>20 m2 ) had higher light levels and lower humidity than the surrounding subtropical broadleaf climax forest (Barik et al., 1992). When Canham et al. (1990) examined gaps (size 78 m2 ) in ﬁve different forest types, they showed that light penetrates to the understory in and around the gap, particularly along the north side of gaps in the northern hemisphere. Light zones increased with latitude, reﬂecting the angle of sunlight reaching the adjacent understory. This peak of light availability along the north edge of gaps was recognized earlier (Minckler and Woerheide, 1965; Minckler et al., 1973). Poulson and Platt (1989) reported strong growth responses along north edges of gaps, but little along south edges. However, the daily duration of direct sunlight is brief even in sizeable gaps. In taller deep-crowned Pseudotsuga menziesii forests in Oregon, a canopy gap produces a much smaller light zone than in four other forest types (Canham et al., 1990). Physical gap measurements are made in a variety of ways. Most widely used are Brokaw’s (1982) deﬁnition of a gap as an opening extending down to a height of 2 m, and Runkle’s (1982) similar deﬁnition of gap as the area beneath a canopy opening, used in tandem with the expanded gap concept which extends from the opening outward to the bases of surrounding trees. However, widely differing gap criteria hamper efforts to compare studies. Runkle (1992) reviewed the options and problems, and provided a gap-sampling protocol developed through workshops with participants who study diverse forests (see also Christensen and Franklin, 1987). One key point of difference is how tall the understory can be within a gap; another is the lower size limit for recognized gaps. When small gaps and understory-ﬁlled areas are excluded, this poses particular problems for attempts to utilize gap dynamics for insight into windstorms. The same features used to delineate gaps are also measured as gap consequences, an appropriate approach for scrutinizing gap dynamics but misleading and somewhat circular when all wind-related turnover is of interest. When light gaps favor shade-tolerant species Contrary to expectation, existing plants are the usual

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

beneﬁciaries of light gap formation. Especially when small, gaps are ﬁlled by upward growth of understory trees and saplings (Watt, 1947; Bray, 1956; Brewer and Merritt, 1978; Runkle, 1984, 1990b; Canham, 1985, 1989, 1990; Spies et al., 1990a; Cho and Boerner, 1991) or simply by lateral growth of adjacent trees (Gysel, 1951; Hibbs, 1982; Runkle and Yetter, 1987; Frelich and Martin, 1988; Spies and Franklin, 1989). Even in tropical forests, new species have only sparse gap-related opportunities because so few gaps are large; the existing understory normally controls the response (Brokaw and Scheiner, 1989), as in temperate-zone forests. When a layer of shade-tolerant understory trees advances toward the canopy, succession is in effect accelerated by disturbance (Lorimer, 1980), especially if mortality is heaviest among shade-intolerant trees. Similarly, disturbance by logging advances succession in Michigan forests, despite the large size of openings (Abrams and Scott, 1989). In some forests the loss of individual trees does not create measurable light gaps at all (Lieberman et al., 1985, 1989; Spies and Franklin, 1989; Webb, 1989). The non-gap component of forest turnover and of windstorm dynamics is generally overlooked despite these aforementioned studies, but recent ﬁndings of Hubbell (1999) have drawn attention to the limits of gap dynamics at last. Discrete, measurable gaps are most likely to form in forests with a monolayered canopy (Fig. 7.2), broad tree crowns, relatively short trees, (Spies et al., 1990a), and an open or weak-wooded subcanopy (Webb, 1989).

Fig. 7.2. Diagram showing how a monolayered forest (a and b) forms larger, more discrete light-gaps than a multilayered forest with understory layers (c and d).

In both old-growth and successional forests of western Oregon and Washington, most tree deaths are not followed by gap formation (Spies et al., 1990a). If gaps form, they are often restricted to the canopy

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layer. The tall narrow conifers form small gaps if any, well below the gap-size threshold (estimated at 700– 1000 m2 ) for shade-intolerant Pseudotsuga menziesii (Douglas-ﬁr) to regenerate (Spies and Franklin, 1989). However, gaps in earlier stages of succession have larger effective sizes because of the lower stature of the canopy. Even in successional forests, however, the principal response to tree-falls is release of a shadetolerant understory (Spies and Franklin, 1989; Spies et al., 1990a). Even large-area, high-intensity windstorms can fail to set back succession in cleared patches. Glitzenstein et al. (1986) in Texas concluded that disturbance by a tornado caused accelerated succession in at least one (Pinus–Quercus) of the two forest types studied. Peterson and Pickett (1995) in Pennsylvania found that average species richness increased over a six-year period following tornado disturbance, but that this increase occurred both in openings and in surrounding forest. Despite the large blow-down size, shade-intolerant tree species did not predominate, perhaps because of an absence of propagules (Populus and Prunus might be expected) in a large old-growth tract; instead the shade-tolerant taxa were the chief beneﬁciaries (Acer saccharum, Fagus grandifolia, Tsuga canadensis). However, yellow birch (Betula alleghaniensis) also beneﬁted; this intermediate-tolerance species probably would not regenerate without openings. In some cases, shade-intolerant species that initially appear in light gaps do not survive long and are replaced quickly by shade-tolerant species (Clinton et al., 1994; Kupfer and Runkle, 1996). Resprouting of damaged trees can constitute a major component of the windstorm response, as seen after hurricane disturbance in South Carolina (Putz and Sharitz, 1991). Likewise, rapid and dense resprouting of Populus tremuloides minimized establishment opportunities for other species in a recent blow-down (1995) in northwestern Minnesota (S. Webb, pers. observ.). Many temperate-zone forests have one or more shade-tolerant tree species that establish beneath an intact canopy and respond to light gaps. In the original vision of gap dynamics, gaps are ﬁlled not by patches of light-demanding species but clumps of shade-tolerant saplings (Bray, 1956). In eastern North America’s deciduous forests, such shade-tolerant gap ﬁllers include Acer saccharum and Fagus grandifolia (Brewer and Merritt, 1978; Barden, 1980, 1981; Canham, 1990; Runkle, 1990b; Poulson and Platt,

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1996), and also Nyssa sylvatica (Orwig and Abrams, 1994), all of which grow slowly under the canopy and respond to release when canopy trees die overhead or nearby. Runkle found that gap regeneration was dominated by shade-tolerant Acer saccharum and Fagus grandifolia in forests both in Ohio (Runkle, 1990b) and in North Carolina (Runkle and Yetter, 1987). Brewer and Merritt (1978) reported the same situation in Warren Woods, southeastern Michigan [but see Poulson and Platt (1996) who anticipated shadeintolerant trees in large gaps within this same forest]. Canham (1985) showed that Acer saccharum grows into the canopy slowly, having an average of three periods of suppression interspersed with periods of release corresponding to increased light from nearby canopy gaps. Acer saccharum appears in study after study as a major beneﬁciary of windstorms (Held and Bryant, 1989; Webb, 1989; Mladenoff, 1990; Cho and Boerner, 1991), on account of its shade tolerance and advanced regeneration poised to capture newly available resources, and also due to its wind-ﬁrmness and ﬂexibility (King, 1986). Many gap studies in other parts of the world reveal similar dynamics that favor shade-tolerant trees established before gap formation. In Chilean rainforests, gaps were ﬁlled not by the existing canopy species of disturbance origin (Nothofagus dombeyi) but by shade-tolerant Aextoxicon punctatum (Veblen, 1985a,b). Existing emergent canopy species were not reproducing and will likely disappear if only small gaps form (Veblen, 1985a). In old-growth forests of Argentina, Nothofagus betuloides formed a welldeveloped understory capable of responding to release, except where excluded by an understory of the subcanopy tree Drimys winteri; where Nothofagus pumilio was present, its seedlings and saplings also expanded upward into gaps from pre-gap establishment (Rebertus and Veblen, 1993b). In one of two xeric localities in Chile also examined by Veblen (1989a), Nothofagus dombeyi was regenerating in the understory and thus increasing at the expense of a canopy species, Nothofagus antarctica, that does not regenerate in tree-falls. The second Chilean site also had a predisturbance understory, but in this case the understory more closely matched the canopy, suggesting that gap dynamics maintain the forest’s existing composition (Veblen, 1989a). In extensive light gaps within a cool temperate Fagus crenata forest in Japan, only 2.5% of all stems within 36 gaps were of pioneer species (Nakashizuka, 1984); suppressed saplings of Fagus

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crenata were abundant and likely to replace most windthrown canopy trees (Yamamoto, 1989). In a more species-rich warm temperate Japanese forest, nearly all predicted gap successors were from ﬁve species widely present as suppressed saplings (Yamamoto, 1992). In subalpine old-growth forests of southwestern British Columbia, the many small, persistent gaps are ﬁlled primarily by lateral-growth of neighboring trees or by shade-tolerant Abies amabilis (Lertzman and Krebs, 1991; Lertzman, 1992). Likewise, in central Sweden, even the largest gaps ﬁlled with the shadetolerant Picea abies, while little regeneration was seen for more light-demanding species: Betula spp., Pinus sylvestris and Populus tremula (Qinghong and Hytteborn, 1991). Light gaps can play a key structural and developmental role in the forest without admitting shade-intolerant species, as in several coniferous forest examples. In a Colorado subalpine forest, Abies lasiocarpa and Picea engelmannii are both shade-tolerant dominants, but Abies lasiocarpa reproduces under the closed canopy whereas Picea engelmannii only begins to regenerate when canopy gaps form (Aplet et al., 1988). In this case, new colonization occurs in gaps but the colonist is a shade-tolerant species already present; here gaps interact with other life-history characteristics (such as greater longevity for Picea engelmannii; Veblen, 1986) to reorganize and maintain the community. Similarly, in old-growth Picea abies forests in Sweden (Leemans, 1991), saplings and understory trees were concentrated in gaps but scarce under closed canopies, demonstrating the importance of gaps for regeneration; but the species of gap saplings and trees was the same (Picea abies) as that dominating the canopy. A similar dynamic apparently operates in subalpine forest of central Japan, where advanced regeneration by shadetolerant Abies mariesii might be accompanied into canopy gaps by occasional shade-intolerant codominants that sustain less mortality (Yamamoto, 1995). Shade-intolerant trees lacked colonization opportunities in a Nothofagus forest in New Zealand, where gap beneﬁciaries were trees and saplings already present in the understory (Stewart and Rose, 1990; Stewart et al., 1991). The two dominant species differed in life history traits, but both responded favorably to windstorms, to the exclusion of other trees. Shade-intolerant plants in gaps In contrast, there are examples of gaps (created by windstorms and other causes) where shade-intolerant

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

trees are supported in a diversity-enriching patchdynamics pathway. Large clearings from hurricanes, tornadoes, and blow-downs seem most likely to favor shade-intolerant species, but smaller light gaps can also have this effect under some (but, as has been shown, not all) circumstances. One component of diversity enrichment comes from the buried pool of viable seeds, whose germination is triggered by light or other environmental changes when the canopy opens up. In eastern North America, Prunus spp. (Marks, 1974; Dunn et al., 1983; Held and Bryant, 1989) and Rubus spp. account for the majority of the woody plant seed bank. A dense thicket of Rubus brambles is common in large blow-downs shortly after the storm. Figure 7.1 shows such a thicket in the Flambeau Tract in Wisconsin eight years after a 1977 down-burst struck an old-growth forest. Rubus allegheniensis increased for three years following one tornado disturbance in Pennsylvania but then declined as tree saplings closed in (Peterson and Pickett, 1995). However, the same locality had little growth of Prunus, apparently because its seeds were sparse in the soils of this isolated old-growth forest (Peterson and Pickett, 1995). Most of the wind-dispersed Betula species (birches), which are light-demanding to various degrees, depend upon disturbances to regenerate in sizeable blowdowns. Six years after the Pennsylvania tornado, a large clearing of 400 ha2 supported high densities of Betula alleghaniensis and B. lenta, which rarely regenerate in closed forest (Peterson and Pickett, 1995). A similar increase for Betula alleghaniensis developed throughout the large Wisconsin blow-down (Fig. 7.1) (Dunn et al., 1983). Following the 1938 hurricane in New England, birches germinated on many wind-throw mounds (including Betula alleghaniensis, B. lenta, B. papyrifera, B. populifolia; Spurr, 1956; Henry and Swan, 1974), where they still persist today (S. Webb, pers. obs.). Both B. alleghaniensis and B. lenta show rapid lateral growth into small gaps (Hibbs, 1982). In a Vermont montane spruce–ﬁr forest, Betula papyrifera thrived in gaps (of unknown cause) but was scarce under closed canopies (Perkins et al., 1992). A spruce–ﬁr forest in the southern Appalachian Mountains (Tennessee) had ten times more Betula alleghaniensis in gaps than in background forest plots (White et al., 1985a,b). However, small tree-fall gaps in a Michigan forest were not large enough to beneﬁt Betula alleghaniensis (Mladenoff, 1990). Other trees that depend upon light can thrive

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in areas of wind-throw. By eleven years after a Kentucky tornado, several shade-intolerant trees had increased in density, including Prunus serotina, Quercus muehlenbergii and Q. rubra (Held and Bryant, 1989). A decline in density but an increase in basal area were seen for Fraxinus americana, another light-demanding species generally expected to beneﬁt from canopy openings. However, other shade-intolerant and midtolerant species had declined (Carya cordiformis, Celtis occidentalis, Gleditsia triacanthos and Juglans nigra) while the shade-tolerant Acer saccharum increased its numbers from 35% to 53% of relative density (Held and Bryant, 1989). Beyond such studies linked with known windstorms, one can draw inferences, with caution, from the forest light-gap literature. Even though gap-oriented studies tend to oversample large gaps, and sometimes use the presence of shade-intolerant trees as evidence for past gaps, there are surprisingly few cases in which shade-intolerant species predominate in gaps. In most investigations, shade-tolerant trees are the most important respondents in light gaps, as already discussed. Gap size is one key factor in controlling whether the existing ﬂora is simply restructured, or whether instead the gap patch supports a distinctive assemblage. The sizes of light gaps in one Indiana beech–maple (Fagus–Acer) forest corresponded to sizes of patches of seral species (Fraxinus spp., Liriodendron tulipifera, Quercus spp., Sassafras spp., etc.), suggesting that these species got established in light gaps (Williamson, 1975). The light-demanding tree Liriodendron tulipifera showed no pattern of suppression and release in its tree rings, like that seen in more shadetolerant trees (Orwig and Abrams, 1994). L. tulipifera depends on sizeable openings and is more common following agriculture than following disturbances that form small gaps (Clebsch and Busing, 1989). Only in the largest clear-cut patches was Liriodendron tulipifera reproducing ten years after logging in Illinois (Minckler and Woerheide, 1965). Nevertheless, this species is an important codominant within southern Appalachian forests, along with other shadeintolerant (Fraxinus americana) and shade-tolerant (Tsuga canadensis) trees, suggesting regular gapforming disturbance events of some magnitude in the past (Lorimer, 1980; Barden, 1981). In Warren Woods, southwest Michigan, Poulson and Platt (1989) concluded that Liriodendron tulipifera needed less light than Fraxinus americana but more than Prunus

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serotina; all three of these shade-intolerant trees were expected to reproduce in sufﬁciently large gaps. The abundance of oaks in eastern North America has long suggested a disturbance origin for the forests, given their lack of shade tolerance. Quercus rubra might maintain dominance after large-scale wind-throw if vigorous oak seedlings are already present (Hibbs, 1982), an uncommon situation today. In smaller gaps, Quercus rubra does not reproduce (Bray, 1956; Webb, 1989; Cho and Boerner, 1991). Other studies also suggest that light-demanding trees enter into the forest only in relatively large gaps. In an Ohio forest, Runkle (1990b) found shadeintolerant Liriodendron tulipifera in only the four largest of 36 gaps, and found another shade-intolerant tree, Fraxinus americana, at high densities in large, young gaps but not in older gaps, suggesting inability to persist until forest closure. However, this 12-year resurvey showed most gap takeover by pre-existing shade-tolerant stems (Acer saccharum and Fagus grandifolia), as was also true for gaps in more speciesrich hardwood forests in the Southern Appalachian Mountains (Runkle and Yetter, 1987). These Southern Appalachian cove forests have received much scrutiny, often with conﬂicting conclusions about the role of gap dynamics in maintaining shade-intolerant trees. Clinton et al. (1994) reported a general increase in richness with increasing gap size, and Barden (1981) found shade-intolerant trees more prevalent in multiple-tree gaps than in single-tree gaps. Although 97% of all predicted gap successors were shade-tolerant Acer saccharum and Fagus grandifolia, shade-intolerant trees were expected to persist through occasional capture of large gaps (Barden, 1981). A similar correlation was found between gap area and tree species richness in a subtropical broadleaf climax forest in India (Barik et al., 1992). These gap size/diversity correlations reﬂect in part an increase in light availability as gap size increases (Canham et al., 1990). One mechanism by which light gaps theoretically could enrich diversity is through gap partitioning, where gaps of different sizes, or different zones within a gap, favor different trees (Denslow, 1980). Runkle et al. (1995) sought but did not ﬁnd differences suggesting gap partitioning by two coexisting species of Nothofagus in southern New Zealand. Sipe and Bazzaz (1994, 1995) examined three species of maple (Acer pensylvanicum, A. rubrum and A. saccharum) for gap partitioning, by transplanting seedlings into controlled experimental gaps. After three years, the

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three species showed differences in growth responses to large gaps but not to small gaps (nor in control areas). Linkages with environmental measurements showed little support for the gap-partitioning concept. Heterogeneity within gaps has not been widely examined in temperate forests; more studies with control areas like those of Sipe and Bazzaz (1995) would also be useful to elucidate dispersal constraints on patchy gap vegetation. In temperate forests in general, evidence for gap partitioning has not been forthcoming aside from specialized utilization of wind-generated substrates (Yamamoto, 1992; see also section “Mounds, pits, and woody debris”, pp. 208–211 below). Gap enrichment depends strongly upon what species are part of the system. For example, in Chile Nothofagus dombeyi has greater gap importance at high elevations where shade-tolerant competitors are absent (Veblen, 1985b). In a Minnesota Pinus–Abies forest, opportunities for shade-intolerant species result from the absence of the shade-tolerant layer of understory Acer saccharum so prevalent on other soils nearby (Webb, 1989). The importance of understory development is further discussed below. In search of more insights into wind disturbance, the analogy between harvest openings and windstorm gaps must be drawn with caution, because of the very different substrates that result when logged trees are removed, often by heavy machinery; wood substrates are removed, and wind-throw mounds do not form. In Maine, gaps created by forest harvest differed from natural gaps, the former having more herbaceous plants and Rubus, in part because of larger gap size but also because of soil disturbance during the harvest (Kimball et al., 1995). Gaps from logging were examined in a New Hampshire hardwoods forest; these large openings (292–1032 m2 ) had shade-tolerant trees dominating the smaller range, but admitted shade-intolerant and mid-tolerant trees in the larger openings, including Acer rubrum, Betula alleghaniensis, B. papyrifera and Prunus pensylvanica (McClure and Lee, 1993). The importance of the understory The forest understory is a major part of the equation for windstorm responses. A well-developed understory can preclude gap formation and can inhibit new colonists where gaps do form (Fig. 7.2). As already emphasized, many temperate forests include understory layers of shade-tolerant trees, saplings, and seedlings that are likely to beneﬁt from death of canopy trees (Canham,

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Fig. 7.3. Photographs illustrating well-developed understory layers of saplings (Acer saccharum) and shrubs (Corylus cornuta and others) in Minnesota Pinus forest (Webb, 1989). Similar shade-tolerant understories are the major beneﬁciaries of individual tree-falls in many forests around the world.

1985, 1990) and can continue to shade the forest ﬂoor against establishment of light-demanding species. The case of Acer saccharum was described above and is illustrated by Fig. 7.3. In one Japanese beech forest, an extensive pre-disturbance seedling bank of Acer mono and Fagus crenata might help to explain the absence of pioneer trees such as Betula spp. in gaps (Hara, 1983, 1985, 1987). If the understory comprises wind-ﬁrm species, it will beneﬁt from canopy damage; however, if understory species are damage-prone, then canopy gaps are more likely to extend deeper toward the forest ﬂoor (Webb, 1989). The importance of the understory ﬂora is illustrated by the contrasting effects of a moderately severe 1983 windstorm in two Minnesota stands (Webb, 1989). Discrete light gaps formed in a Pinus–Abies community because the shade-tolerant understory ﬁrs (Abies spp.) had weak wood and a growth form rendering them vulnerable to snagging by a falling tree; the ﬁrs rarely survived the fall of overstory trees. Meanwhile, in a nearby Pinus–Acer community, few tree-falls produced measurable, discrete light gaps. Here the understory was dominated by shade-tolerant species with very strong wood (Acer saccharum and

Ostrya virginiana), which survived and often beneﬁted from the fall of overtopping canopy trees. Thus the wind-ﬁrmness of shade-tolerant understory trees helps to explain why these two forests responded differently to the same disturbance event. Tall understory trees, a thicket of shrubs, or a growth of bamboo can all dominate the dynamics of a wind-damaged forest patch. Understory bamboos inﬂuence tree regeneration in some South American and Asian forests. Veblen (1989b) found that tall bamboos (Chusquea spp.) inhibited tree regeneration in gaps in Chilean and Argentinian Nothofagus forests, but that trees did regenerate in gaps farther south where bamboos were absent. In China, Taylor and Zisheng (1988) found different gap-replacement trends in two old-growth Abies faxoniana/Betula utilis stands, one with more bamboo (Sinarundinaria fangiana) and the other with regeneration by gap-favored Betula trees. In temperate broadleaf forests of Japan, a dense undergrowth of bamboo (Sasa nipponica) apparently inhibits seedlings of trees that elsewhere establish a seedling bank (Acer mono, Fagus crenata) but does not inhibit reproduction by sprouting for some other trees (Acer japonicum, A. sciadophylloides, Magnolia obovata;

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Nakashizuka et al., 1992; Yamamoto, 1995). In a dense matrix of bamboo, tree-fall gaps are generally not colonized by shade-intolerant trees, except where windthrow mounds and rotting logs provide bamboo-free substrates (Nakashizuka, 1989). Small trees that never play a canopy role are sometimes the major gap beneﬁciaries. These include several species in Japanese broadleaved-evergreen forests (Yamamoto, 1992), the understory tree Cornus ﬂorida in a New Jersey hardwood forest (Ehrenfeld, 1980), Acer pensylvanicum in a northern hardwoods forest (Hibbs et al., 1980), the understory tree Ostrya virginiana in northwestern Minnesota hardwood forests (Webb, 1989), and Drimys winteri in Argentina (Rebertus and Veblen, 1993b). Other types of understory vegetation can also play a role in windstorm dynamics. Tall shrubs characterize the hardwood and mixed forests of northwestern Minnesota (Fig. 7.3), contributing alongside Acer saccharum understories to a lack of light on the forest ﬂoor where trees blow down. Many shrubs can sprout in high densities in response to light gaps (Gysel, 1951; Dunn et al., 1983). In some areas of the southern Appalachians (North Carolina), an understory of Rhododendron, a tall broadleaved evergreen shrub, inhibits tree regeneration before and after gaps form (Runkle, 1985; Clinton et al., 1994). At one New Zealand site, heavy fern cover combined with scattered shrubs to limit densities of tree seedlings and saplings, compared with two otherwise similar stands of oldgrowth Nothofagus forest (Stewart et al., 1991). Few studies of post-storm colonization examine herbaceous plants (Dunn et al., 1983; Peterson and Campbell, 1993). Several surveys of light gaps and other microsites (mounds, pits, logs) not associated with windstorms have demonstrated diverse herbaceous response patterns across a range of forest types. Distinctive assemblages appeared in light gaps (Goldblum, 1997) and on old mounds and pits in eastern New York (Beatty, 1984) and on mounds and pits in Pennsylvania (Peterson and Campbell, 1993). These studies had large sample sizes and careful statistical analyses. Other studies have found patterns that do not suggest species enrichment within light gaps or on microsites. In an Illinois forest, the herbaceous plants on logs, mounds, and pits were the same species found within one meter on the undisturbed forest ﬂoor (Thompson, 1980). In a Tennessee forest, fallen logs supported lower total diversity than the soil, but those plants present had more vigorous growth on logs (Bratton, 1976).

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Increases in cover but not richness are sometimes seen in the understory layer (Moore and Vankat, 1986). For example, three years after experimental gaps were created, the fern Thelypteris novaboracensis had increased in density within gaps in northwestern Pennsylvania (Collins and Pickett, 1988b); the vernal Erythronium americanum responded more immediately with more stems in gaps than non-gap areas, but did not subsequently increase in cover (Collins and Pickett, 1988a). Large gaps but not small gaps had higher densities of seedlings of the light-demanding small tree Prunus serotina. Thus, as with trees, the understory plants of winddisturbed landscapes are often rearranged and restructured without changes to the overall composition and richness of the forest. However, well-designed timetransgressive studies suggest that some windstormtype disturbances can help to explain the complex patterning observed in the understory of the temperatezone forest. Herbivory Herbivory can also inﬂuence gap dynamics and responses to windstorms. Many forested areas of eastern North America are subject to anthropogenically elevated levels of herbivory by white-tailed deer (Odocoileus virginianus). The composition and open structure of the shrub and seedling layer thus may be an artifact of human settlement and predator eradication. Deer browsing is thought to explain the disappearance, over 50 years, of a Pennsylvania forest understory (Whitney, 1986) and also to explain the scarcity of Tsuga canadensis regeneration in another Pennsylvania understory, with consequences for tornado responses (Peterson and Pickett, 1995). Herbivory also contributes to the paucity of tree establishment in Patagonian blow-downs (Veblen et al., 1996), while introduced herbivores in New Zealand are thought to inﬂuence forest regeneration there as well (Jane, 1986). Thus, the all-important understory with its strong inﬂuence on windstorm consequences may itself look very different from the understory from which today’s canopy trees emerged. Today’s deep light gaps in some temperate forests would have been broken up vertically by understory layers in the past. Mounds, pits, and coarse woody debris In addition to light gaps, another mechanism by which

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

windstorm disturbance might modify the community is by generating substrate heterogeneity: wind-throw mounds and pits when trees uproot, stumps when trees snap in the wind, and rotting logs in either case (Fig. 7.4). Such microsites differ from the surrounding forest ﬂoor in such features as temperature, moisture retention, nutrients (Stark, 1994), and freedom from competing plant roots; and thus might support different biotic assemblages. As with light gaps, such microsites do not always enrich diversity but instead play quite different roles in different forests.

Fig. 7.4. Illustration of microsites formed when trees are uprooted or snapped by windstorms. (a) Uprooted trees typically form a mound and a pit, although (b) pits are sometimes absent; (c) some tree-falls form no basal microsites; (d) snapped trees leave behind stumps and, as in all cases without salvage logging, a rotting log.

Mounds and pits The wind-throw mound (or knoll) and its associated pit (also called a cradle or crater) form only when trees are uprooted, a mode of damage nearly always caused by windstorms. In many old-growth forests the ground is distinctly uneven, a sign of past windstorms. Mounds and pits can cover 1.6–48% of the forest ﬂoor (Beatty, 1984; Webb, 1989). Long after the upturned roots have rotted away, mounds of soil and rock will persist, perhaps for 300–500 years (Denny and Goodlett, 1956; Stephens, 1956), if not leveled by agricultural or silvicultural activities. Soil proﬁles are rearranged by the uprooting process (Lutz and Griswold, 1939; Lutz, 1940; Lyford and MacLean, 1966; Vasenev and Targulyan, 1994) and various other soil features are modiﬁed (Stone, 1975; Beatty and

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Stone, 1986; Norton, 1989; Habecker et al., 1990; Semmel, 1993). In many cases, mounds support more trees than either pits or surrounding forest (Denny and Goodlett, 1956; Lyford and MacLean, 1966; Henry and Swan, 1974; Raup, 1981; Collins and Pickett, 1982; Nakashizuka, 1989). The exposed mineral soil on mounds is an ideal seed bed for small seeds (Hutnick, 1952), and mounds are free of competition from the extensive living roots elsewhere. The same species that proliferate in sizeable light gaps are also particularly successful on wind-throw mounds (Hutnick, 1952; Henry and Swan, 1974; Dunn et al., 1983), as is true for tropical pioneer species observed by Putz et al. (1983). Acer saccharum and Prunus serotina, both trees, were most abundant on mound tops in a Pennsylvania forest (Collins and Pickett, 1982). Several herbaceous plant species also had highest importance on mounds in this Pennsylvania forest, and the same was true in a New York forest, although here the distinctiveness of mound assemblages was greater in one forest type (without hemlock) than in another (Beatty, 1984). Tree ferns (Cyathea smithii and Dicksonia squarrosa) are specialists on wind-throwmounds in a New Zealand podocarp forest, where the dominant podocarps (Dacrydium cupressinum and Prumnopitys ferruginoides) are scarce on mounds (Adams and Norton, 1991). Conversely, plant diversity and cover can be low on wind-throw mounds in some forests (Denny and Goodlett, 1956; Peterson et al., 1990). Alongside the attraction of exposed soil, mounds pose challenges to would-be colonists: higher temperature extremes and lower levels of organic matter, litter cover, cation exchange capacity, nutrient content, and snow cover than pits (Beatty, 1984). Within a few years of a Pennsylvania tornado, Peterson et al. (1990) found lower species richness on mounds than in pits. In a Pinus/Acer forest in northwestern Minnesota, old wind-throw mounds supported lower densities and lower richness than the surrounding forest ﬂoor, with no unique species or assemblages on mounds (Webb, 1988). In a nearby Pinus/Abies forest, the characteristically small mounds were also ﬂoristically indistinguishable from control areas. This lack of signiﬁcance for microsite patches might result from the small size of light gaps and the dense pre-windstorm understory. In this location, most mounds have not yet been colonized 13 years after they formed in 1983 (S. Webb, pers. observ.); if this time lag is typical, then

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light gaps may close before the mound or pit substrates are suitable for colonization. Mounds also can differ from undisturbed microsites by impeding ingrowth by vegetative sprouts. Both mounds and pits had densities of Fagus grandifolia lower than the background values in two different Pennsylvania forests where this tree normally reproduces vegetatively (Collins and Pickett, 1982; Peterson and Campbell, 1993). However, root sprouts of Populus tremuloides were slightly more numerous on mounds than elsewhere in the Minnesota Pinus/Acer forest (Webb, 1988), perhaps deriving from the wind-thrown Populus trees themselves. Wind-throw mounds and pits also differ in size and conﬁguration depending upon the species of windthrown tree and the details of the uprooting event (Beatty and Stone, 1986; Mattheck et al., 1995). Colonization patterns can vary with the size of mounds (Webb, 1988) or pits (Peterson et al., 1990) and can also vary within a mound or pit (Hutnick, 1952; Collins and Pickett, 1982; Peterson et al., 1990). Pits (“craters”; Falinski, 1978) can be hostile or favorable microhabitats, depending upon the depth to the water table and the size of the wind-cleared openings; the role of pits also apparently shifts as they age over time. In many eastern North American forests, pits support low diversity and low densities of plants (Thompson, 1980; Raup, 1981; Beatty, 1984), with assemblages different from those on mounds or elsewhere (Beatty, 1984). A New Zealand temperate forest also had pits with low diversity, gradually ﬁlling with silt and then with a dense growth of bryophytes (Adams and Norton, 1991). The inhibiting factor in a central New York forest was the deep litter that accumulates in pits; when litter was removed, germination and establishment were enhanced for plants previously growing only on mounds (Beatty and Sholes, 1988). Litter depth was also negatively correlated with species richness in a northwestern Pennsylvania forest (Peterson and Campbell, 1993). Another problem in pits is moisture accumulation (Hutnick, 1952). In a Polish forest dominated by Picea abies, a shallow water table apparently caused pits to ﬁll with short-lived plants of aquatic and mud habitats (Falinski, 1978). Hydrophilic plants also colonized pits in a northern Pennsylvania forest, but in this case the pits had higher species richness than mounds (Denny and Goodlett, 1956). The higher moisture level in pits, if not excessive, can promote seedling survivorship whereas the exposed

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mound might be a microhabitat subject to drought. Acer rubrum, a tree of both uplands and swamps, was most common in pits following hurricane damage to a New Hampshire forest (Henry and Swan, 1974). After a Pennsylvania tornado (Peterson and Pickett, 1990), new pits were more rapidly colonized than mounds, and supported higher species richness. Even Betula alleghaniensis, common in other forests on mounds only, was abundant in pits at this site. Perhaps pits are favorable immediately following windstorms before litter accumulates, whereas on mounds mineral soil is exposed for a longer period of time. However, tree seedlings were more abundant as pits aged in Minnesota forests with a moderate-windstorm disturbance regime (Webb, 1988). Here pits of all ages supported lower plant density and diversity than mounds or control areas, with a ﬂora distinguished by an absence of otherwise ubiquitous Acer saccharum and the presence of otherwise uncommon seedlings of Quercus rubra and Tilia americana. These geographic differences in vegetation on pits and mounds probably result from the complex interplay of local climate, soil type, depth to the water table, and disturbance magnitude and intensity as they inﬂuence the size and duration of light gap conditions. Stumps and rotting logs Not all wind-thrown trees are uprooted; thus mound and pit microsites are not always created (Fig. 7.4). When Peterson and Pickett (1991) reviewed 17 windstorm studies, they found a wide range in the prevalence of uprooting. From 0% to 91% of wind-damaged trees were uprooted, but in the majority of examples uprooting accounted for less than 50% of the damage. In two Minnesota forests, 37% and 80% of wind-throws following a moderate thunderstorm resulted in snapped trees (excluding bent trees; Webb, 1988), and 33% of wind-thrown trees following a tornado in Pennsylvania were snapped (Peterson and Pickett, 1991). Twice as many trees were snapped as uprooted by a hurricane in South Carolina sloughs (Putz and Sharitz, 1991); 50% of damaged trees were snapped by the same hurricane elsewhere in South Carolina on a ﬂoodplain (Duever and McCollom, 1993). The stump formed by tree breakage represents a substrate different from mounds or pits but similar to rotting logs. Of course, stumps and logs are formed whenever trees die, not only as a result of windstorms. Undisturbed old-growth forests can have 45% (Maser and Trappe, 1984) or more of the surface covered with

DISTURBANCE BY WIND IN TEMPERATE-ZONE FORESTS

dead wood. A widespread policy of salvage logging hampers ability to probe the importance of such coarse woody debris in forests. Coarse woody debris can serve to conserve nutrients through gradual nutrient release during the decay process (Harmon et al., 1986), but can immobilize nitrogen in other forests (Zimmerman et al., 1995). These woody substrates can also be important sites for germination and establishment. Best known in this regard are the so-called nurse logs that support tree seedlings unable to establish elsewhere in forests of the Paciﬁc Northwest [Washington and Oregon (U.S.A.) and adjacent Canada; Minore, 1972; Franklin and Dyearsness, 1973; Christy and Mack, 1984; Harmon and Franklin, 1989]. Tree seedlings predominate on rotting wood in other regions as well, particularly for conifers in North American coniferous forests (Harmon et al., 1986; Webb, 1989; Pauley and Clebsch, 1990; Gibson and Brown, 1991; Hofgaard, 1993), but also in Nothofagus forests in Chile (Veblen, 1985b) and New Zealand (Stewart, 1986), and in an oldgrowth Fagus–Abies forest in Japan where bamboo inhibits seedlings on soil (Nakashizuka, 1989). In the New Zealand forests, the two dominant, shade-tolerant tree species (Nothofagus menziesii and Weinmannia racemosa) constituted most regeneration on fallen trees in both small tree-falls and within a larger area of windthrow (Stewart, 1986). Experimental work in the Paciﬁc Northwest indicates that these woody microsites are favorable primarily because of freedom from competition from herbs and mosses on the forest ﬂoor; these experiments ruled out waterlogging (on the forest ﬂoor), litter shedding (on logs), nutrient differences, and differential seed predation as alternative explanations (Harmon and Franklin, 1989). In other less moist environments than the temperate rainforests of that research, rotting wood might have other signiﬁcant advantages for tree reproduction: constancy of moisture availability, abundant mycorrhizae (Harvey et al., 1979), and thin layers of germination-promoting mosses (Place, 1955; St. Hilaire and Leopold, 1995). Stumps can differ somewhat from rotting logs in their ecological role. Only on stumps did Tsuga heterophylla become established within gaps that were otherwise ﬁlled by Abies amabilis (Minore, 1972). Stumps sometimes resprout, perpetuating the occupation of a site by the same species. Resprouting is common for some species (particularly angiosperms) but not possible for others (particularly conifers). Immediately

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after a Pennsylvania tornado, 25% of all snapped trees resprouted, but only 68% of these were still alive four years later (Peterson and Pickett, 1991). In a Japanese forest, Castanopsis cuspidata resprouts from stumps and thus can persist, despite its absence from the predisturbance understory (Yamamoto, 1992). Another contribution of coarse woody debris is as habitat for diverse fungal species including those involved with decomposition. Old-growth forests preserve a high level of fungal diversity that is not paralleled in second-growth forests or even in virgin forests where wind-thrown trees are removed (Hood et al., 1989). In experimental gaps within a Pseudotsuga menziesii forest in Oregon, the “skirt” zone at the bases of stumps, snags, and even living trees supported higher densities of mycorrhizal root tips than fallen logs (Vogt et al., 1995). A key question about the importance of rotting wood is how quickly the stump or log will become suitable for seedling establishment. Wood decomposition rates vary widely among localities [from decades in subalpine forests (Lambert et al., 1980) to within a few years in tropical rain forests (Lieberman et al., 1985)], and also among tree species (Hood et al., 1989) and sizes (larger trees decay more slowly; Maser and Trappe, 1984). Where portions of fallen trees are propped above the ground, decomposition is slowed in eastern and midwestern forests in the U.S.A. (pers. observ.), but in temperate rainforests a standing dead tree will decompose more quickly than a fallen tree (Cline et al., 1980). Below-ground gaps: soil moisture and nutrient changes As foresters have long recognized, light is not the only resource for which plants compete in forests. When a tree blows down, the pulse of new resources previously usurped by that tree includes not only light but also below-ground resources: water, nutrients, and even space itself (McConnaughay and Bazzaz, 1991). Thus, a below-ground “gap” is created, an opportunity like the above-ground light gap for increased growth by existing plants or possibly for establishment by new colonists. Here I review studies of change at and below the soil level, beyond those changes involved with formation of microsites (mound, pit, stump, log) as already discussed. Evidence for below-ground gaps comes from several

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types of research. Trenching experiments have demonstrated release from competition when neighboring roots were excluded, even without canopy clearance (Toumey and Kienholz, 1931; Korstian and Coile, 1938; Horn, 1985). Comparative photographs show a lush growth of tall shrubs and grasses eight years after trenching, in contrast with unvegetated untrenched plots (Toumey and Kienholz, 1931). In a more manipulative study, soil moisture levels were higher and transplanted seedlings of Acer rubrum and Cornus ﬂorida had better survivorship or growth where neighboring roots were severed within a North Carolina forest (Horn, 1985). Low root densities can also suggest the presence of below-ground gaps. The biomass of ﬁne roots was signiﬁcantly less in small tree-fall gaps than in surrounding forest soils in a Pennsylvania study (Wilczynski and Pickett, 1993). In contrast, small gaps in a Pinus contorta forest in southeastern Wyoming differed little from undisturbed areas in abundance of ﬁne root tips and ectomycorrhizal roots; however, increasing size of the clearing up to a 30-tree gap was correlated with a decrease in root growth and ectomycorrhizae. Thus, the large gaps had detectable underground components (Parsons et al., 1994a,b). Total root biomass, and especially biomass of small roots, were also lower in openings within New Jersey Pinus rigida forests (Ehrenfeld et al., 1995). Measurements of ﬁne-root biomass and turnover are increasingly utilized as indicators of environmental variation and change (Vogt et al., 1993). Nutrient availability has also been examined in gaps. In a Pinus contorta forest in Wyoming (Parsons et al., 1994b), the gradient of increasing gap size was paralleled by increasing nitrogen mobilization, suggesting a below-ground gap with respect to the consumption of nitrogen. These ﬁndings might be different in windstorm-caused gaps where, unlike the logged openings of that project, fallen trees remain on the ground. Nitrogen cycling was also modiﬁed by small tree-fall gaps in Acer saccharum/Tsuga canadensis forest in Michigan, but in different directions under the two dominant trees (Mladenoff, 1987). Nitrogen mineralization and nitriﬁcation were greater in gaps within evergreen, acid-soil Tsuga areas but were less in gaps in Acer saccharum areas, as compared with paired intact control (non-gap) areas. In the soils of New Jersey pine barrens with very low nitrogen content, openings of apparent ﬁre origin had virtually no organic matter and ﬁve times less litter compared

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with the forested matrix, perhaps because of a scarcity of ericaceous shrubs with their abundant surface roots. Nitrogen availability in these openings was low because of low inputs, while nitrogen availability in the forest was low because of rapid cycling (Ehrenfeld et al., 1995). With the recent explosion of inquiry into belowground ecological processes, there is a growing realization that the forest ﬂoor is highly heterogeneous and that analysis should be stratiﬁed by patch type (Vogt et al., 1995). Windstorms contribute to spatial patterning via light gaps, coarse woody debris, mound/pit microtopography, and diminished root competition and thus enhanced resource availability where trees are killed. Seed dispersal, seed banks, and seed predation The future of a wind-disturbed area depends not only upon resource pulses and understory development but also upon seed ecology: seed banks, seed dispersal, seed predation, and seed pathogens. Pathogenicity is lower in light gaps for a tropical forest (Augspurger, 1984), while the other seed-related interactions can be stronger or weaker after windstorms under various circumstances. Because, as has been seen, windstorm responses are usually controlled by pre-existing vegetation, seed dynamics may not be of universal importance. Seed ecology is thus perhaps most important in large blow-downs, on mounds and other wind-created microsites, and for shade-tolerant forest species in areas with small gaps and/or well developed understory vegetation. Seed availability itself depends upon the history of the site, the size of the clearing as it inﬂuences distances to seed sources, and the behavior of seed dispersers and seed predators as modiﬁed by the presence of fallen trees and light-gaps. The seed bank is a pool of buried, viable, but dormant seeds. For most dormant forest seeds, germination is triggered by environmental signals linked to light (V´asquez-Yanes and Orozco-Segovia, 1994). As already noted, in North America the seed bank is most conspicuously the source of light-demanding Prunus and Rubus, genera known for long persistence as seeds and for abundant growth in large blow-downs. However, the seed bank is actually a surprisingly diverse collection of propagules (Leck et al., 1989; Beatty, 1991). A high degree of spatial variability in seed pools (V´asquez-Yanes and Orozco-Segovia, 1994) reﬂects in part the history of the forest. Peterson and

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Carson (1996) found a minimal seed-bank response in a Pennsylvania old-growth forest without previous history of disturbance, whereas nearby second-growth forests responded to the same tornado with extensive germination from a rich seed bank of plants presumably present at an earlier stage of succession. Peterson and Carson reviewed other studies that also support the importance of stand history to the abundance of buried seeds. In contrast, Beatty (1991) found that buried seed pools were similar in composition to the existing vegetation in a wide range of habitats in central New York, thus reﬂecting recent seed rain rather than stand history. This was also true of seed banks of mounds and pits: most species present as seeds were already established in the vegetation. Mladenoff (1990) also found that seed banks were compositionally similar to the existing forest understory in small gaps in western Upper Michigan; interestingly, gap seed banks had larger densities of seeds than the seed bank of the undisturbed forest (Mladenoff, 1990). Seed dispersal can also inﬂuence windstorm responses (Schupp et al., 1989). If seed sources are too distant, then a plant that is perfectly capable of growing on disturbed ground will be excluded by the dispersal constraint. In tropical rainforests, vertebrates disperse seeds abundantly and over substantial distances (Denslow and Gomez Diaz, 1990). Dispersal distances seem more limited for many temperate-zone species. In a New Hampshire northern hardwoods forest, the quantity of wind-dispersed seeds of both Acer saccharum and Betula alleghaniensis declined exponentially with distance from a forest edge, while larger nut-like seeds of Fagus grandifolia were virtually absent from the open area (Hughes and Fahey, 1988). The seed rain might extend farther into tree-fall gaps than into the logged openings of that study, because coarse woody debris would provide cover and perches for vertebrate dispersal agents (McDonnell and Stiles, 1983) and could also introduce roughness and thus more wind turbulence which would promote deposition of wind-dispersed propagules. Other studies document substantial dispersal distances for beechnuts and acorns (Johnson and Adkisson, 1986), and paleoecological tracking of postglacial tree migrations also shows that most temperate forest trees have a surprising capacity for dispersal over long distances (Davis, 1981b, 1983; Webb, 1986, 1987). Seed predation rates can also change when trees fall, because neither vertebrate nor invertebrate seed predators utilize all microhabitats equally. For example,

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uprooted trees inﬂuence the movements of rodents in a Polish forest (Olszewski, 1968) and presumably elsewhere. Ant diversity and abundance were enriched in wind-fall areas in a Picea/Fagus forest in Bavaria (Theobald-Ley and Horstmann, 1990), including some species of seed-consuming ants. In a Minnesota study, post-dispersal predation on seeds of the tall shrub Prunus virginiana was much heavier in closed forest than in adjacent open ﬁelds or small tree-fall gaps (Webb and Willson, 1985), perhaps reﬂecting avoidance of exposed sites by vertebrate foragers. Such predators caught in traps baited with P. virginiana seeds included Clethrionomys gaperi, Eutamias minimus, Peromyscus leucopus, Spermophilus tridecemlineatus, Tamias striatus, and Zapus hudsonius. Meanwhile, smaller ant-dispersed seeds of the woodland herb Uvularia grandiﬂora faced heavier post-dispersal predation (after elaisome removal) in open ﬁelds than in small tree-fall gaps or closed forest (Webb and Willson, 1985). Similar studies by Whelan et al. (1991) with Cornus drummondi and Prunus serotina also showed spatial heterogeneity in predation intensity depending upon the microhabitat, but furthermore demonstrated high variance in predation patterns between years. In both disturbed and intact forests, a better understanding is needed of seed predation, which could help explain the rarity of some seedlings and the abundance of others.

CONCLUSIONS

Generalizations are elusive when one considers the myriad scenarios that have been documented for windstorm consequences in temperate forests. Windstorm disturbance can enrich diversity or deplete it at the landscape scale, and can set back succession or accelerate it within patches where trees have blown down. Most temperate forests around the world have understory layers of shade-tolerant species that are capable of responding to canopy openings. No general predictions about windstorm consequences can be made on the basis of forest type (conifer, evergreen, deciduous, xeric, humid) or successional stage. Each tree species has its own individualistic combination of shade tolerance, wind-ﬁrmness, and other life-history features that dictate its vulnerability to wind damage and capacity to regenerate following the disturbance. The type of response depends upon the details of the forest itself and can hinge upon understory

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composition, sometimes controlled by the presence or absence of a single key taxon (such as bamboo or Acer saccharum). A pre-storm understory of shade-tolerant saplings, shrubs, or forbs may inhibit new regeneration by light-demanding species, even where large canopy openings are created. A post-storm ﬂush of resprouting by trees and shrubs can have a similar effect. Thus, strong autogenic forces operate in the wake of a windstorm. Human activity has modiﬁed forest structure and understory development indirectly by changing patterns of herbivore abundance and distribution; deer in eastern North America and introduced mammals in New Zealand exert so much inﬂuence that it is difﬁcult to understand natural forest dynamics. The disturbance response also depends upon allogenic aspects of individual windstorm events and of the disturbance regime in a given locality. Windstorms do not vary along a single-dimensional continuum; for example, hurricanes disturb larger areas (higher magnitude) but with lower wind-speeds (lower intensity) compared with tornadoes. Blow-downs and tornadoes that create large clearings also have fringe zones with scattered tree-falls. Surveys of mortality and damage show that tree species differ in vulnerability to wind damage but with site-speciﬁc inﬂuence of confounded factors such as topography, tree size, and fungal infection. Some trees survive even the most severe windstorms. Two strong concerns hamper the utility of much winddamage research: methods of assessing damage must be standardized (Everham, 1995), and the undamaged component of the forest must be sampled as context for interpreting the proﬁle of mortality. Wind-throw mounds, pits, stumps, and rotting logs each play important roles in windstorm responses in some but not all forests. Thus, large tracts need to be protected from salvage logging for the sake of tree regeneration, which is dependent on dead wood in many temperate forests. The natural patchiness of vegetation makes causal patterns difﬁcult to tease apart without long-term study over large areas. Unfortunately, fallen and damaged trees are usually removed following windstorms, leaving only small, fragmented nature preserves where dynamics of coarse woody debris can be studied. More generally, our understanding of windstorm disturbance regimes is hampered by extensive forest clearance and conversion to agriculture of most portions of the temperate forest biome. For example, of virgin forest present in the United States before European

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settlement, only a tiny fraction remains (0.4%–1.6% for eastern regions; Frelich, 1995; Davis, 1996). Yet nearly all sizeable remnants in eastern North America have been impacted by catastrophic windstorms in recent decades. Knowledge of windstorm consequences necessitates the preservation of extensive areas so that the full range of patch types and the full range of disturbance responses are preserved (White, 1987; Foster et al., 1996). Long-term studies are needed, for several reasons: to monitor windstorm return times, damage patterns, and responses; and to examine interactions among windstorm events. Recent projects that began after speciﬁc windstorms should be continued not only to track changes in blow-down areas but also to track subsequent windstorm occurrences and consequences. Remarkably few studies actually examine dynamics that are known to relate to windstorms rather than to unspeciﬁed gap-forming causes and chronologies. Such gap-dynamics studies provide only limited elucidation of windstorm consequences, because they exclude turnover not occurring in gaps but include gaps created by non-wind causes. In the future, every effort should be made to identify causal agents and timing of gap formation, not only for insights into disturbance but also to understand the extent to which the present gap conﬁguration is in equilibrium, and to assign ecological mechanisms to the dynamics. Future gap research should also incorporate “control” nongap patches in which parallel sampling is done, to permit distinguishing aspects of the gap composition and structure which are related to the gap and others which simply mirror the background context. The extensive literature cited in this chapter provides a framework in which to place additional research on gap dynamics in temperate forests. Early detailed studies of forests in the eastern United States led to generalizations that patch dynamics promote landscape-level diversity, generalizations which turned out to be contradicted by studies in other regions of the United States and of the world. Further progress in disturbance ecology requires a global perspective and more research outside of North America. ACKNOWLEDGEMENTS

This chapter was supported in part by research grants from the Minnesota Department of Natural Resources and from Drew University. For helpful suggestions I thank Charlie Cogbill, Sara Cooper-Ellis, David Foster, Lawrence Walker and Dennis Whigham.

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Chapter 8

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

INTRODUCTION

Disturbance has been deﬁned in various ways (e.g., Sousa, 1984; Rykiel, 1985; van Andel and van den Bergh, 1987; Pickett et al., 1989), in part because disturbance theory must deal with a wide variety of phenomena under that name (Pickett et al., 1989). A useful deﬁnition of disturbance for forests is “a relatively discrete event causing a change in the physical structure of the environment” (Clark, 1990). In forests, this change in physical structure refers primarily to damage and removal of aboveground biomass (Grime, 1979), and removal of the litter layer and mixing of surface soil layers (e.g., Putz, 1983). Changes in the physical structure of the environment (Clark, 1990) are concomitant with changes in environmental variables below the main canopy (e.g., soil resources, light quantity and quality, and temperature: Chazdon and Fetcher, 1984; Bellingham et al., 1996). For individual disturbance events, the size of the canopy opening is often correlated with most of the direct and indirect changes in environmental variables and resource levels (Brokaw, 1985b). Even small gaps (Canham and Marks, 1985) provide opportunities for regeneration and adult growth (Sousa, 1984; Oliver and Larson, 1990). Consistent with the above deﬁnition, in this chapter we focus on forest-canopy disturbances that occur as discrete events in time. We are primarily concerned with those relatively small canopy disturbances (background canopy gap disturbance) caused by a variety of agents and larger disturbances caused by major windstorms (catastrophic wind disturbance). A modiﬁcation of Clark’s (1990) deﬁnition of the scale of wind-generated disturbances, which represents a continuum of disturbance (Everham and Brokaw, 1996; Lugo and Scatena, 1996), is useful because it

can be used to divide disturbances into those that we will consider to be background canopy disturbances (<103 m2 : Table 8.1), the scale most often used to investigate tree-fall dynamics in forests, independent of the causes of the tree-falls (Brokaw, 1985a) and catastrophic wind disturbances (>103 m2 : Table 8.1). We only indirectly deal with major disturbance events (e.g., ﬁre, ﬂoods, river meanders, landslides, shifting agriculture) that both open the canopy and more substantially disturb the soil, understory, and litter layer (see Uhl, 1982a; Lugo et al., 1983; Johns, 1986; Oliver and Larson, 1990; Lugo and Scatena, 1996). Some types of canopy disturbances have been described by the concept of “patch dynamics” (White and Pickett, 1985), in which the disturbance effects are relatively discrete spatially, and where there is a “shifting mosaic” of patches of different ages. After disturbance, the patch changes through time and regains characteristics of the older patches. This concept applies well to canopy gaps formed by the fall of one to a few trees and blow-downs where many trees are felled together (see Sousa, 1984; White and Pickett, 1985; Whitmore, 1989; Clark, 1990). The concept is problematic in that gap edges are not always Table 8.1 Scale of wind-generated disturbances in neotropical moist forests 1 Type of disturbance

Approximate area of effect (m2 )

Hurricanes/typhoons

105 –107

Blow-downs/wind-throws

103 –105

Single tree-falls

102 –103

Branch falls

101 –102

1

223

Modiﬁed from Clark (1990).

224

Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

obvious, openings do not always extend through all canopy layers (Lieberman et al., 1989), small gaps may be closed from the side (Runkle and Yetter, 1987), and patches may also be internally heterogeneous (White and Pickett, 1985). In contrast to discrete gaps, hurricane disturbance may often, though not always, be better described as “diffuse” (Pickett and White, 1985), particularly where the canopy is thinned and no discrete patches are formed (Everham and Brokaw, 1996). In this chapter we focus on background canopy disturbances that create canopy gaps extending to the understory [see Brokaw (1982a) and Popma et al. (1988) for deﬁnitions of canopy gaps] and catastrophic wind disturbances of both discrete and diffuse nature. One can view disturbances as having causes and effects (Rykiel, 1985) to which ecological systems respond (van Andel and van den Bergh, 1987). Three largely similar dichotomous groupings of causes of disturbance have been used (but see Lugo and Scatena, 1996): endogenous versus exogenous (White and Pickett, 1985), internal versus external (Vooren, 1986; van der Meer and Bongers, 1996b) and biotic versus abiotic. Endogenous causes (e.g., disease, competition, epiphyte and vine loading of canopies, and herbivory) originate from within the biotic community while exogenous causes (e.g., drought, wind, lightning, and rainfall) originate from without (White and Pickett, 1985). Endogenously caused disturbances may largely occur through the gradual fall of trees that died standing, hereafter referred to as standing mortality (Vooren, 1986; Krasny and Whitmore, 1992; van der Meer and Bongers, 1996b). Exogenously caused disturbances can occur gradually (e.g., from standing mortality owing to drought) or suddenly (e.g., from wind-throw). There is interaction between these two groups of causes (White and Pickett, 1985). For instance, though major wind disturbance obviously has a largely exogenous component, trees weakened by disease are more prone to damage (Putz and Sharitz, 1991; see also Everham and Brokaw, 1996). Wind is an important cause of canopy disturbance, and probably the primary cause of canopy disturbance in forests where large tropical storms or other strong, though localized, wind storms are common occurrences (see Lugo et al., 1983; Shaw, 1983; Brokaw, 1985b; Walker et al., 1991; Lugo and Scatena, 1996) and where other causes of major disturbance are not operative (Leighton and Wirawan, 1986; Foster, 1990). Vegetative response to disturbance depends on characteristics of the disturbed site (e.g., severity of damage), species availability

(e.g., propagule availability and survival through the event), and post-disturbance species performance such as growth rates, sprouting ability, and survival (Pickett et al., 1987; Pickett and McDonnell, 1989; Lugo and Scatena, 1996). Our objectives in this chapter are to build on the work of Everham and Brokaw (1996) and other reviewers by describing the causes and effects of background canopy disturbance (Denslow, 1980; Hartshorn, 1980; Brokaw, 1985a,b) and catastrophic wind disturbance (Lugo et al., 1983; Glitzenstein and Harcombe, 1988; Brokaw and Walker, 1991; Tanner et al., 1991; Lugo and Scatena, 1996; Zimmerman et al., 1996) in tropical and subtropical forests, and comparing vegetative response to catastrophic wind events with response to background canopy disturbance. Finally, we describe ecosystem response to catastrophic wind disturbance.

CAUSES AND EFFECTS OF CANOPY DISTURBANCE

Wind storms in the tropics – distribution and patterns Within the tropics, winds that are not associated with catastrophic events usually impact forests during periods of rainfall that occur chieﬂy in areas of atmospheric disturbance and persist for days, moving irregularly around the landscape, but often toward the west. The distribution, size, and intensity of tropical rainstorms varies temporally and spatially (Schwerdtfeger, 1976; Lauer, 1983), but storms that are potentially strong enough to generate winds causing canopy disturbances can generally be organized into three types: (1) The Intertropical Convergence Zone produces winds, that are rarely strong, over the oceans and widely over continents during rainy seasons. Forest disturbances associated with these types of windstorms are most often associated with individual or small groups of rain squalls. (2) Northern cold fronts (nortes) in certain longitudes (e.g., in the northern Caribbean) penetrate into the tropics and become part of the trade-wind system. Strong winds and showers or thunderstorms are commonly associated with these systems. (3) Larger-scale easterly-winds produce squall-line systems. These occur in West Africa, across the tropical Atlantic to the Caribbean, and infrequently in the central and western Paciﬁc. Wind velocities

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

of more than 20 m s−1 are associated with these systems. These three types of wind disturbances occur throughout the tropics, but their occurrence and intensity are difﬁcult to predict and their impacts are poorly known. Richards (1996) suggested that all tropical areas are subject to wind disturbances before and during convectional rainstorms and thunderstorms, and he cited several examples to demonstrate the range of impacts that wind disturbances can have. In Sarawak, winds associated with squalls created canopy gaps as small as 0.04 ha (equivalent to felling a single tree) but they can also cause catastrophic damage; areas as large as 80 ha, for example, were damaged by wind in swamp forests dominated by Shorea albida. Other authors have also documented the importance of wind as an agent of background canopy disturbance (e.g., Whitmore, 1989). Winds that cause background canopy disturbance can also create catastrophic disturbances that would be difﬁcult to predict under any circumstances. Uhl et al. (1988a), for example, found that local rain and windstorms mostly accounted for the background rate of canopy disturbance in the S˜ao Carlos terra ﬁrme forest in Venezuela, but that windstorms also “occasionally knock down whole sections of forest”. Kellman and Tackaberry (1993) found that “extreme winds of atypical orientation” caused trees to fall in a direction that was not uniformly distributed. Numerous accounts of blow-down and wind-throw disturbances can be found in the literature, but almost all of them are qualitative descriptions (e.g., Richards, 1996; Whitmore, 1989; citations in Clark, 1990; Nelson et al., 1994). Bruenig (1989) used aerial photographs taken in different years to demonstrate the importance of windstorms in creating wind-throws in forested wetlands in Borneo. Bruenig found that the rates of single-tree gaps varied among forest types from 0.2–3.0% of the area per year, but that the rate of disturbance in one forest type appeared to be about 1% for single tree-fall gaps and for openings created by wind-throws. Bruenig suggested that forests with taller trees and trees with a high height/diameter ratio are more susceptible to lightning and wind-throw. Forested wetlands appear to be particularly susceptible to blowdown and wind-throw disturbances because the trees have very shallow roots. Clearly, there have been too few quantitative studies to characterize the importance of blow-downs and wind-throws in tropical forests, even though their impacts on forest structure and

225

species composition are long-lived (Hubbell and Foster, 1986). Most catastrophic wind disturbances are associated with tropical cyclones, the general term for hurricanes and typhoons, which are the most destructive types of windstorms in the tropics and subtropics; their damaging effects can extend far into the temperate zone (Encyclopædia Britannica, Inc., 1992; Boose et al., 1994). They are intense storms (as large as 150–250 km across) with maximum wind speeds that are, at least, 32.7 m s−1 . Heavy rain is also usually associated with tropical cyclones. The number of tropical cyclones per year, world-wide, varies between approximately 30 and 100, and they occur primarily near Southeast Asia, the Caribbean, and adjacent waters (Vega and Binkley, 1993), and in the southwest Paciﬁc and Australian waters (Fig. 8.1). In addition to disturbing natural ecosystems, tropical cyclones also cause enormous economic destruction and human suffering (Encyclopædia Britannica, Inc., 1992). Background canopy disturbance Overview Canopy disturbance is now viewed as integral to an understanding of tropical forest ecology (Hartshorn, 1978; Whitmore, 1975), and much of the data cited in this chapter were obtained as part of efforts to characterize what we have deﬁned as the “background” pattern of disturbance in tropical forests, disturbances that are usually caused by branch-falls, standing dead trees, and the fall of one to several adjacent trees. The average rate of gap formation in humid lowland tropical forests has been reported to be 1 ha−1 yr−1 with a range of 0.7–2.6, opening ~1–2% of the forest per year with a range of 0.5–3.6% yr−1 (Denslow, 1987; Swaine et al., 1987; Whitmore, 1989; Hartshorn, 1990; Jans et al., 1993; Yavitt et al., 1995; Lugo and Scatena, 1996). Average gap size, shortly after gap formation, ranges from 54 to 120 m2 (see review in Jans et al., 1993). Small gaps are much more common than large gaps (Lawton and Putz, 1988; Uhl et al., 1988a; Chandrashekara and Ramakrishnan, 1994). In lowland moist and wet forests, the largest gaps range considerably in size. Brokaw (1982a,b) measured gap sizes of 232 m2 and 342 m2 in an old forest on Barro Colorado Island (BCI), Panama. Yavitt et al. (1995) measured a gap size of 604 m2 in a young forest there. Sanford et al. (1986) measured gaps as large as 781 m2 at La Selva, Costa Rica. Gap turnover is a function of

226

Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

Fig. 8.1. Major tracks (arrows) and frequency (shaded areas) of hurricanes and typhoons. Source: Encyclopædia Britannica, Inc. (1992).

the size of gaps and their rate of occurrence and can be thought of as the number of years it would take to cover 100% of the forest in gaps (assuming there is no gap overlap). Estimated gap turnover rates range from 60 to 145 years in humid lowland forests (see review in Jans et al., 1993), and rates of gap formation can vary considerably from year to year (Mart´ınez-Ramos et al., 1988; Lawton and Putz, 1988; Dickinson et al., 1999). The variation is even greater when rare and large disturbance events are included in the estimates of turnover rates (Brokaw, 1985b; Leighton and Wirawan, 1986; Whigham et al., 1990). Variation among forests Similarities among forests in gap size, frequency, and turnover have been emphasized (Lawton and Putz, 1988; Hartshorn, 1990; Kapos et al., 1990; Jans et al., 1993; Yavitt et al., 1995). The similarities, however, are likely owing to a bias towards studying humid lowland forests in the neo-tropics (P. Hall, pers. commun., 1997) not obviously affected by major periodic disturbance (see summary data above). There are fewer data on background canopy disturbance rates for forests subject to major periodic disturbances, dry tropical forests, montane forests, wetland forests, and paleotropical forests, but differences among forest types are to be expected. Several examples follow. In a wet forest in Puerto Rico periodically disturbed by periodic catastrophic wind events, mean gap size (76 m2 ) and gap formation rate (0.8 gaps ha−1 yr−1 ) are below the midrange values for humid forests not impacted by periodic

major disturbance (Scatena and Lugo, 1995). The gap formation rate in a semideciduous forest in the Mexican Yucatan (1300 mm annual rainfall) periodically hit by major disturbance is exceedingly low at 0.2 gaps ha−1 , resulting in 0.07% of the forest being disturbed in an average year (Dickinson et al., 1999). The average gap size in the Yucatan forest (<50 m2 ) is similar, after correcting for a lower minimum gap size, to an elﬁn montane forest in Costa Rica (Lawton and Putz, 1988), and at the low end for humid lowland tropical forests. Similarly, maximum gap sizes also are lower for the Yucatan site (>180 m2 ; Dickinson et al., 1999), a montane forest site (>135 m2 ; Lawton and Putz, 1988), and in a West African forest with a long dry season (244 m2 ) studied by Jans et al. (1993). Jans et al. also reported a correspondingly long gap turnover time (244 years) for that site. Compared with humid lowland neo-tropical forests, gap turnover times also were long (154–375 yr) in two equatorial Southeast Asian forests (Jengka Forest Reserve, Peninsular Malaysia: Poore, 1968; Samarinda, East Kalimantan, Borneo: Riswan et al., 1985). Long gap turnover times can be partly explained by small gap size in the Jengka (Whitmore, 1975). East Kalimantan is periodically affected by severe droughts, which may explain long turnover times there (Leighton and Wirawan, 1986). In the following section and in Table 8.2, we describe what we expect to be the main sources of variation among sites in background canopy disturbance (also see Brokaw, 1985b). These include: wind regime, forest

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

227

Table 8.2 Predictions about variation in background canopy disturbance regimes #

Prediction

1

Dry, high-altitude forests, and extremely nutrient-poor forests should have the lowest average gap size owing to small crown sizes.

2

Forests dominated by tall trees with thin crowns will have linear gaps, with less effect on understory light levels.

3

Forests subject to natural and anthropogenic major disturbance (ﬁres, blow-downs, hurricanes, agricultural clearing, logging, landslides, ﬂooding events, and river meanders) will have smaller gaps and a lower rate of gap formation than forests that do not experience such events or that have not experienced such an event in >200 years.

4

Shallow rooting of trees on fertile, extremely nutrient-poor, or waterlogged soils will increase the rate of tree-fall and prevalence of uprooting (and soil disturbance), while seasonally dry forests and forests on moderately infertile and well-drained soils will exhibit low rates of tree-fall and more stem-snapping.

5

Forests on slopes will have a higher rate of tree-fall than forests on ﬂat terrain only if soils and parent material are unstable and rooting is shallow.

6

Forests with high abundance of woody vines should exhibit larger gap sizes and more frequent gaps.

7

High rates of standing mortality will result in a smaller average gap size and lower overall rates of gap formation.

8

Except where continuous strong winds have large effects on structure, forests with higher frequencies of strong, gusty winds (e.g., where rainstorms are more frequent) should have higher rates of gap formation than forests where winds are less gusty.

9

Dry tropical forests should have smaller gaps and lower rates of gap formation than humid forests, owing to smaller crown sizes, a more stable tree architecture, higher rates of standing mortality, periodic ﬁre (if applicable), and a shorter period in which most gap causal factors operate.

structure (as affected by site conditions, biogeography and species composition, and prior catastrophic disturbances), tree anchorage, standing mortality, and vines and epiphytes. The effects of wind are often mediated by the other factors. We also compare dry and wet tropical forests, which appear to differ in several of the above factors, and we predict that they will prove to vary considerably in background canopy disturbance. Wind regime Forests subject to more frequent strong and gusty winds should have more canopy disturbances than forests where winds are calm (Grace, 1977; Richards, 1996; see also Table 8.2, Prediction 8). In terms of damage to trees, high mean wind speeds, strong gusts, and abrupt changes in wind direction are probably most likely to cause damage (Gloyne, 1968; Grace, 1977). Within a given forest, sites that are most exposed to gusty winds have higher tree-fall rates (see review in Brokaw, 1985b). Strong gusts (from 780 to over 100 km hr−1 ) have been reported to occur on multiple occasions in a given year in several sites (Bultot and Grifﬁths, 1972; Lawton and Putz, 1988; Uhl et al., 1988a). Whitmore (1975) suggested that variation among forests in the frequency of convectional storms and squall lines would explain variation among forests in disturbance. Squall lines along the Malaysian Peninsula are reported to be the

most important cause of large gaps (Whitmore, 1975). The generally calmer winds of equatorial Southeast Asia have been hypothetically linked to high rates of standing mortality and low overall gap formation rates in forests there (P. Hall, pers. commun., 1997). In forests with distinct dry seasons, a peak in gap formation occurs during the middle of the wet season [see review in Brokaw (1985b), and also Matelson et al. (1995)], although not all wet season gaps can be ascribed to wind (see below). Ninety percent of the gaps sampled at La Selva in Costa Rica were formed during the 6 wettest months (Hartshorn, 1989), and in a cloud forest in Monteverde (Costa Rica) 70% of the gaps formed in one year were caused during one severe, but common type of storm (see Lawton and Putz, 1988). Dry tropical forests are also likely to have more wind-induced damage during the wet season, as most of the strong wind gusts occur during the wet season (Bultot and Grifﬁths, 1972). In addition to causing individual tree-fall gaps, wind damage can enlarge existing gaps by toppling trees at the gap edge (Bruenig, 1989; Whitmore, 1984), or cause clustering of gaps owing to increased turbulence around existing gaps where the canopy is uneven (Poorter et al., 1994). Wind is often the coup de grace to a tree made vulnerable by such things as disease, poor rooting conditions owing to waterlogged soils, and rain-loading

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Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

(Gloyne, 1968; Whitmore, 1975; Brokaw, 1985b; Richards, 1996). Though rain-loading of the canopy can cause large, live trees to fall in calm, rainy weather (Strong, 1977), standing mortality cannot generally be directly attributable to wind, although extreme winds may lead to eventual standing mortality (Shaw, 1983; Whigham et al., 1991). Where gap formation by standing mortality is minimal, we expect that the largest proportion of gaps would be caused by uprooting, snapping, and live limb-fall owing to wind. Trees fall in the direction of the prevailing winds in a subalpine temperate forest, strong evidence of the importance of wind in those environments (Wooldridge et al., 1995). Wind speeds vary in several ways, most importantly perhaps with geographic setting and the frequency of storms (Whitmore, 1975). Average coastal wind speeds are generally higher than inland wind speeds at the same altitude (Grace, 1977). Windstorms may often be more common somewhat inland from the coast owing to orographic rain storms (Dale, 1959; Whitmore, 1975). Average wind speed increases with altitude (Grace, 1977), although canopy gaps do not seem to become more frequent at higher altitudes (Lawton and Putz, 1988; Matelson et al., 1995). Trees whose physiognomy is shaped to a large degree by strong and continuous winds, such as trees in exposed cloud forests, are not as prone to wind damage as trees that are not so conditioned (Lawton and Putz, 1988; Brokaw, 1985b; Richards, 1996). This is likely in large part owing to the increased turbulence associated with rough rather than smooth canopy topographies (Gloyne, 1968; Grace, 1977). Topography signiﬁcantly affects wind speed and direction above canopies (Grace, 1977). When wind speeds are high, forests on the lee side of ridges can experience severe turbulence (Grace, 1977). Wind is responsible for the formation of disturbances over a wide range of scales, from single tree-falls (e.g., Denslow, 1987) to blow-downs (Dunn et al., 1983; Uhl et al., 1988a; Clark, 1990; Nelson et al., 1994), and swaths of hurricane-disturbed forest (Everham and Brokaw, 1996; Richards, 1996). In forests that are not typically affected by major wind disturbance, the incidence of multi-tree wind-throws may be the most important factor in disturbance regimes and resulting patterns of species abundance differentiating one forest from another (Denslow, 1987). Forest structure Effects of site conditions: Variation among forests in

above-ground structure (shoot characteristics, including crown size, average tree height, and allometry) should be reﬂected in the size distributions of tree-fall gaps, their effect on light conditions at ground level, and on the frequency of gap creation. A positive relationship between tree size (measured as trunk diameter) and gap size [i.e., area of the hole in the canopy that extends to ground level; see Brokaw (1982a) and Popma et al. (1988)] has been documented (Brokaw, 1982a; Lawton and Putz, 1988; Clark and Clark, 1996). Small gap sizes in montane and drier forests are a reﬂection of this (see p. 226 above). It is not clear whether bole height, bole thickness, or crown size most inﬂuences gap size. However, the ﬁndings that the height at which the bole gave way has little effect on gap size (Lawton and Putz, 1988) and that gaps formed by uprooting trees are no larger than gaps formed by snapped trees (Jans et al., 1993) indicate that crown size is most important (Richards, 1996; see also Table 8.2, Prediction 1). All else being equal, large crowns are on tall thick boles (King, 1991) and are poised to do maximum damage to surrounding trees when they fall. Tree architecture varies with wind regime, rainfall, soils, topography, altitude and latitude [see review of hypotheses regarding physiological limits on tree size by Stevens and Perkins (1992)]. Rainfall patterns are primary determinants of above-ground forest structure (Beard, 1955; Holdridge et al., 1971; Ellenberg, 1979; Lieberman et al., 1996). Forests with the highest densities of emergents that have large crowns should occur in lowland tropical wet and moist forests on well-drained soils (Holdridge et al., 1971; Grubb, 1989), especially on lower slopes or in areas with little topographic relief (Ashton and Hall, 1992). Forest height in a wet Puerto Rican forest is highest on better drained areas and lowest where soils are often waterlogged (Lugo et al., 1995). Kira (1978) found that the tallest forests at a number of Southeast Asian sites occur where rainfall was around 2000 mm yr−1 and was evenly distributed throughout the year. Tree height drops both in wetter sites and in sites with decreasing amounts and increasing seasonality of rainfall. In concert, the height to diameter ratio decreases in drier and more seasonal forests, that is, a tree of 1 m dbh becomes shorter (Kira, 1978). Thus, crown size may not decrease as quickly as tree height with decreasing rainfall. High stand turnover rates (as at La Selva) should lead to low densities of very large trees (Lieberman et al., 1985; Clark, 1996). Stand and basal area turnover appear to increase generally in

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

productive sites (Phillips et al., 1994), consequently, such sites may have more, but smaller, gaps. Forests on extremely nutrient-poor soils should also have small tree-fall gaps as the trees are typically of low stature, although the effects of low nutrient status are often confounded with periodic moisture deﬁcits (Whitmore, 1975; Grubb, 1989). Kapos et al. (1990) found larger gaps, and larger trees, on a relatively nutrient-rich site than on a nutrient-poor site. Periodic drought in eastern Borneo killed the largest and tallest trees, leading to a broken emergent stratum (Leighton and Wirawan, 1986; Richards, 1996; see below). Large trees on ridges are particularly vulnerable to drought-related mortality, since the soils dry quickly (Leighton and Wirawan, 1986; Ashton and Hall, 1992; Richards, 1996). This effect may have contributed to lower rates of gap formation and lower gap sizes on ridges compared with slopes in West African seasonal forests (Poorter et al., 1994). Rooting was impeded on hill crests (Poorter et al., 1994), perhaps compounding an effect of periodic drought. Tree height and crown size decrease with increasing altitude (Holdridge et al., 1971; Whitmore, 1975; Ellenberg, 1979; Lieberman et al., 1996). Consistently stronger winds at higher altitudes (Grace, 1977) lead to shorter trees with thicker boles, primarily due to the mechanical effects of continuous winds on tree growth (Lawton and Dryer, 1980; Telewski, 1995). Also, there appears to be a temperature effect on forest height along altitude gradients (Pendry and Proctor, 1996). A hypothesis that there should exist a pattern of increasing crown width from the temperate to the tropical zone was proposed by Terborgh (1985), but King (1991) did not ﬁnd such a pattern. Forests dominated by tall trees with small crowns produce small gaps that are often linear in shape, providing little opportunity for regeneration of lightdemanding species (Putz and Appanah, 1987; see also Table 8.2, Prediction 2). The distribution of forests in which tall trees with small crowns predominate is unclear, but examples include certain zones in Sarawak peat swamps (Whitmore, 1975) and Pasoh Forest Reserve in Peninsular Malaysia (Putz and Appanah, 1987). Both of these sites are nutrient-poor, and trees also experience periodic moisture stress (Whitmore, 1975; Putz and Appanah, 1987). Canopy trees in the tallest forests may generally show a reduction in crown expansion as compared with canopy trees in somewhat shorter forests (King, 1996).

229

Biogeographic and species-composition effects: Biogeographic effects on forest structure may occur; they should not be confused with site effects, and the effects of past disturbance on species composition. Variation in species composition within and among biogeographic regions may affect disturbance regimes when different species have different stem architecture (Beard, 1945a; Wadsworth and Englerth, 1959; Jans et al., 1993), wood properties (Putz et al., 1983), rooting patterns (Everham and Brokaw, 1996), and modes of death (Brokaw, 1985b). The high density of very tall, emergent trees in some dipterocarpdominated forests may be related to unique aspects of the reproductive biology and ecology of this Southeast Asian family (Ashton, 1988). Forests in Borneo often have a towering emergent canopy of dipterocarps, and on certain sites 80–100% of all individuals in the upper canopy are dipterocarps. Forests east of Wallace’s Line (where dipterocarps are less species-rich), in Africa and in the Neotropics typically have a sparse cover of emergent trees above the main canopy (Ashton, 1988). Higher densities of large emergents may lead to larger gaps when emergents fall. High rates of standing mortality in certain equatorial forests in Southeast Asia (Table 8.3) may also have a biogeographic component (Hall, 1991). Effects of prior catastrophic disturbances: While most studies on background canopy disturbance have been done in forests not subjected to major disturbance, major disturbances have large and long-term effects on canopy structure (Johns, 1986; Foster, 1988; Saldarriaga et al., 1988) and subsequent background canopy gap disturbance (Spies and Franklin, 1989; Lorimer, 1989). Any disturbance that results in high mortality of large canopy trees would be expected to lead to a period of lower rates of formation of large tree-fall gaps and smaller gap sizes, as the large dead canopy trees are replaced by individuals from smaller size classes (Hartshorn, 1978; Brokaw, 1982b; Denslow and Hartshorn, 1994; Dahir and Lorimer, 1996; see also Table 8.2, Prediction 3). Such disturbances (see Johns, 1986) include blowdowns (Whitmore, 1975), ﬁre (Leighton and Wirawan, 1986), landslides (Guariguata, 1990), agricultural clearing (Saldarriaga et al., 1988), logging, ﬂooding events (Mori and Becker, 1991; Gullison et al., 1996), severe droughts (Leighton and Wirawan, 1986; Woods, 1989) and river meanders (Foster, 1990). As the frequency of large-scale disturbances increases, tree-fall gaps may

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Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

Table 8.3 Standing mortality of tropical forest trees Region

Climate 1

Standing mortality 2 (%)

Soil fertility 3

Bako, Sarawak, NW Borneo

4167 [0]

84

low

Hall (1991)

Mersing, Sarawak, NW Borneo

3905 [0]

68

high

Hall (1991)

Lambir, Sarawak, NW Borneo

2874 [0]

65

moderate

Hall (1991)

Lower Montane Venezuela

1650 [2]

64

−

Carey et al. (1994) 4

Lowland Venezuela

2725 [1–2]

60

−

Carey et al. (1994) 5

Pasoh, peninsular Malaysia

1900 [1]

45

low

Putz and Appanah (1987)

Source

Paracou, French Guiana

3000 [3]

44

low

Durrieu de Madron (1994)

La Selva, Costa Rica

4000 [0]

40

high

Lieberman and Lieberman (1987)

Amazonas, Brazil

2186 [4]

26

low

Rankin de Merona et al. (1990)

Barro Colorado Island, Panama

2656 [3]

14

high

Putz and Milton (1982)

Amazonas, Venezuela

3500 [0]

¾ 10

low

Uhl (1982b)

1

Annual rainfall in mm. The number of consecutive months with precipitation below 100 mm is shown in brackets. All stems that died after excluding stems for which the cause of mortality was unknown. Consequently, estimates in the Table differ somewhat from the original citation. 3 Soil fertility, often estimated by the current authors. 4 Data are averaged over 9 plots in the State of Merida. 5 Data are averaged over 8 plots in the States of Merida, Delta Amacuro, and Bol´ıvar. 2

occur less and less frequently, and eventually become inconsequential to tree population dynamics (Lorimer, 1989). In the Luquillo Experimental Forest in Puerto Rico, where hurricane return times are about 60 years, periodic hurricanes open more of the canopy than do background canopy gaps, in all topographic positions apart from the often waterlogged riparian valleys (Lugo and Scatena, 1996). Extreme cases of a reduction in the importance of background canopy gaps may occur in the “hurricane scrubs” of North Queensland, Australia (Webb, 1958) and in forests in the Caribbean that are damaged frequently and severely (Beard, 1945b). Even minor hurricanes would be expected to reduce tree-fall rates in the years after the hurricane event, as wind disturbance would fell trees that had rot in their trunk or were poorly rooted in the substrate (Putz and Sharitz, 1991). Major canopy disturbances also tend to reduce canopy height owing to the felling of the tallest trees during severe wind events (Wadsworth and Englerth, 1959; Foster, 1988). In forests subjected to frequent major canopy damage, tree heights should remain lower as a result of bole thickening and reduced height growth in response to increased lateral illumination (see Holbrook and Putz, 1989). These effects on tree size should tend to reduce gap size when trees fall. Recovery following complete canopy removal may

provide a maximum bound on the time to recovery of forest canopy structure. The gap disturbance regime (frequency and size) of seasonally dry forests receiving around 2600 mm of rainfall on Barro Colorado Island, Panama, following abandonment of agriculture, was similar to nearby primary forest after 70–80 years (Yavitt et al., 1995). In upland, nutrient-poor, wetforest sites in the upper Amazon basin, stems 40– 60 cm dbh were prevalent in a stand 60 years old, but after clearing it took 190 years for these forests to regain the previous basal area and biomass (Saldarriaga et al., 1988). After approximately 200 years, forests which formed on the trailing edge of river meanders were still undergoing structural changes in species composition towards larger emergents, which would continue to impact background canopy disturbance (Foster, 1990). Recovery following catastrophic wind disturbance, measured in a variety of ways, may take from decades to centuries, depending in part on severity of damage (Everham and Brokaw, 1996). Catastrophic wind events in Puerto Rico recur at a frequency less than the time to forest recovery (Lugo and Scatena, 1996). Tree anchorage Deeply rooted species (Touliatos and Roth, 1971) and forests in which most individuals are deeply rooted

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

(Brokaw, 1985b) or otherwise ﬁrmly anchored by root grafting (Basnet et al., 1992) or root penetration into soil parent material on shallow soils (Wadsworth and Englerth, 1959; Basnet et al. 1992) suffer less damage in general and less uprooting in particular from strong winds (Everham and Brokaw, 1996; see also Table 8.2, Prediction 4). Deep rooting is impeded in waterlogged soils (Hartshorn, 1978) or where impermeable soil horizons occur (Richards, 1996). Rooting depth may be superﬁcial in soils where deeper layers are exceedingly nutrient-poor or highly acid (Richards, 1996). Relatively shallow rooting also appears to occur in soils with high fertility in wet forests (Brokaw, 1985b). On the other hand, deep rooting appears to occur on soils of moderately low fertility, with somewhat open nutrient cycles, and where seasonal drought confers a premium on deep rooting (Richards, 1996). In support of the hypothesized relationship between rooting patterns and gap formation, Kapos et al. (1990) found lower rates of tree-fall on a more infertile soil. Hartshorn (1978) found higher rates of tree-fall on waterlogged soils where deep rooting was impeded (see also Scatena and Lugo, 1995; Lugo et al., 1995). Firm anchorage may result in a lower frequency of tree-fall and a higher proportion of bole-snapping relative to uprooting (Richards, 1996). Uprooting, as opposed to bole snapping, creates disturbance in the soil and litter layer, which is important in vegetation response. Whether a tree was uprooted or snapped may have little effect on gap size, however (Lawton and Putz, 1988; Jans et al., 1993), because uprooting or snapping may often depend more on wood characteristics than on tree size (Putz et al., 1983). In contrast, in Nouragues, French Guiana, relatively small uprooted trees tended to create gaps that were larger than would be expected from their size (van der Meer and Bongers, 1996a). The increase in gap size was owing to slightly higher numbers of fallen trees in gaps caused by uprooted trees on shallow soils where, compared with the parts of the plot with deeper soils, there was poor root-system development and anchorage. Everham and Brokaw (1996) in a review of uprooting during catastrophic wind events, found that the highest rates of uprooting occurred on the wettest sites. Wadsworth and Englerth (1959) reported high rates of uprooting on deep soils and high rates of stem breakage on shallow soils. Uhl et al. (1988a) reported a blow-down in which 80% of trees were uprooted on nutrient-poor soils where tree roots were shallow.

231

However, most stems that die outside of these blowdowns do so after snapping (Uhl, 1982b). Of the forests listed in Table 8.4, few had percentages of uprooting above 50%, after excluding standing mortality and other modes of tree damage and gap formation that do not involve uprooting or trunk breakage. The highest rate was at La Selva, a wet site with relatively nutrient-rich soils. In most forests (Table 8.4), the majority of tree-fall and gap-making events cause little or no soil disturbance. Although there are no strong patterns evident in the Table, several dry and moist forest sites (e.g., Amazonas, Pasoh, Tai and Zagne) had low to moderate rates of uprooting. As expected if relative soil fertility affects uprooting, increasing soil fertility from Tai to Zagne and Para coincides with increasing rates of uprooting. Jans et al. (1993) explained the higher percentage of uprooting at Para (compared with nearby sites Zagne and Tai) to impeded rooting owing to waterlogging of soils and higher gravel content at Para. In contrast, the high rate of stem snapping at Tai (Vooren, 1986; Jans et al., 1993) was attributed to good rooting conditions (Jans et al., 1993). The relatively large size of gaps caused by uprooting in parts of the sample plot with shallow soils appears to account for the increase in the proportion of uprooting among gap-makers over what might be expected from the proportion of uprooting among fallen trees as a whole (compare the 1996a and 1996b data of van der Meer and Bongers in Table 8.4). We have no explanation as to why the leeward cloud forest at Monteverde had a relatively low percentage of uprooting (Matelson et al., 1995) compared with all other sites and with Monteverde windward cloud forest (Lawton and Putz, 1988). In considering the data in Table 8.4, it must be borne in mind that high rates of standing mortality should generally lead to low overall rates of gap formation (e.g., the Sarawak and Pasoh sites in Tables 8.3 and 8.4). Topographic position inﬂuences anchorage and treefall rates, but its effects appear to depend on the instability of the soil and underlying parent material, and on waterlogging or other factors that lead to shallow rooting (Table 8.2, Prediction 5). There are few data on the effects of slopes on canopy gap formation. Canopy gaps may often be more frequent on slopes than on ﬂat terrain (Oldeman, 1978; Berner, 1992; Lugo et al., 1995), but not at all sites (Kapos et al., 1990; Hubbell and Foster, 1986; Scatena and Lugo, 1995). In a Costa Rican montane oak–bamboo forest, an increase in gap-formation probably involved

232

Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

Table 8.4 Relative percentage of gap-makers or fallen trees that were either uprooted or snapped across the main stem Site

Sample 1 Climate 2

Soil fertility 3

La Selva, Costa Rica

GM

4000 [0]

high

90

−

Hartshorn (1980)

Lower Montane Venezuela

FT

1650 [1–2]

−

69

31

Carey et al. (1994) 4

Uprooted (%)

Snapped (%)

Source

Bako, Sarawak, Borneo

FT

4167 [0]

low

56

44

Hall (1991)

Lambir, Sarawak, Borneo

FT

2874 [0]

moderate

54

46

Hall (1991)

Lowland, Venezuela

FT

2725 [1–1.5]

−

52

48

Carey et al. (1994) 5

Monteverde,6 Costa Rica

GM 7

>2500 [−]

−

48

52

Lawton and Putz (1988)

Mersing, Sarawak, Borneo

FT

3905 [0]

high

47

53

Hall (1991)

Nouragues, French Guiana

GM

3000 [3]

low

45

55

van der Meer and Bongers (1996a)

Para, Ivory Coast

GIST

2100 8 [4]

low 9

43

57

Jans et al. (1993)

Zagne, Ivory Coast

GIST

1650 8

low 9

40

60

Jans et al. (1993)

Pasoh, Malaysia

FT

1900 [1]

low

40

60

Putz and Appanah (1987)

Amazonas, Brazil

FT

2186 [4]

low

36

64

Rankin de Merona et al. (1990)

Nouragues, French Guiana

FT

3000 [3]

low

34

66

van der Meer and Bongers (1996b)

Monteverde,6 Costa

Rica

[4]

FT

>2500 [−]

−

26

74

Matelson et al. (1995)

Barro Colorado Island, Panama

FT 7

2656 [3]

high

25

75

Putz et al. (1983)

Tai, Ivory Coast

FT 7

1875 8 [4]

low 9

23

77

Vooren (1986)

Barro Colorado Island, Panama

FT

2656 [3]

high

22

78

Putz and Milton (1982)

Noh Bec, Quintana Roo, Mexico

GM 7

1500 [6]

moderate

18

82

Dickinson et al. (1999)

Tai, Ivory Coast

GIST

1875 8 [4]

low 9

15

85

Jans et al. (1993)

1 Data sources differed in whether the sample included gap-makers (GM), gap-makers from single tree-fall gaps (GIST), or fallen trees (FT) – whether they caused gaps or not. 2 Annual rainfall in mm. The number of consecutive months with < 100 mm of precipitation is shown in brackets. 3 Often estimated by the current authors. 4 Data are averaged over 9 plots in the State of Merida. 5 Data are averaged over 8 plots in the States of Merida, Delta Amacuro, and Bol´ıvar. 6 Leeward (Lawton and Putz, 1988) and windward (Matelson et al., 1995) cloud forest. Rainfall does not include cloud deposition. 7 Trees that fell over at ground level, often owing to a rotten base, are included in the “snapped” category because they cause minor soil disturbance. 8 Estimated by M. Dickinson. 9 F. Bongers, pers. commun., 1997.

higher growth and mortality rates on slopes, downslope leaning of boles, and larger crown volumes on the down-slope side of trees (Berner, 1992). Asymmetric crowns may increase the probability of tree-fall even on ﬂat sites (Young and Hubbell, 1991; Richards, 1996). Soil instability on slopes ampliﬁes the effects of other conditions that would predispose a tree to being toppled (Denslow, 1987), but soil types and their parent materials are differentially resistant to slippage (Guariguata, 1990) and differentially suitable for anchorage. For instance, in palm forests in the Luquillo Experimental Forest in Puerto Rico, tree falls are more frequent on steep slopes with shallow clayey

and often waterlogged soils (Lugo and Scatena, 1996; Lugo et al., 1995). In contrast, Wadsworth and Englerth (1959), also in Puerto Rico, noted that uprooting during hurricane winds was prevented in shallow soils on slopes by root penetration into cracks in stable parent rock, as compared with deeper soils in valleys. As may be the case with the palm forests noted above, waterlogging appears to exert a stronger effect than slope in the Bisley catchments of the Luquillo Experimental Forest where background canopy gap formation rate is highest on waterlogged soils in riparian valleys, but lower on slopes and in upland valleys (Scatena and Lugo, 1995).

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

Standing mortality Standing dead trees tend to form smaller gaps (Brokaw, 1985b; Putz and Appanah, 1987; Hall, 1991; Krasny and Whitmore, 1992; Jans et al., 1993; Midgley et al., 1995) than tree-falls, and high rates of standing mortality should translate into lower overall rates of gap formation because many trees that die standing will not form discernable canopy gaps (Table 8.2, Prediction 7; see also Putz and Appanah, 1987). The depressive effect of high rates of standing mortality on rates of canopy gap formation and on gap sizes should be particularly acute where the trees that die standing are larger than the trees that are uprooted or snapped, as is the case in certain lower montane and lowland sites in Venezuela (Carey et al., 1994) and at Mersing, Sarawak (Hall, 1991), but not in La Selva, Costa Rica (Lieberman et al., 1985). The depressive effect of standing mortality of large trees on gap size is exempliﬁed by a study of three West African sites, among which gap-size distribution and rate of formation did not differ, owing to a preponderance of gaps formed by branch-fall from standing dead emergents in the site with the largest trees (Vooren, 1986; Jans et al., 1993). If rates of uprooting and snapping had been equal, gap sizes should have been larger at the site with the larger trees (Jans et al., 1993). The proportion of canopy gaps caused by dead or dying trees varies considerably: 72% in a temperate forest in New York (U.S.A.) (Krasny and Whitmore, 1992); 70% in a South African forest (Midgley et al., 1995); at least 50% in a West African forest (Jans et al., 1993); 27% in a semi-deciduous forest in the Yucatan (Dickinson et al., 1999); and at least 2% in a cloud forest at Monteverde, Costa Rica (Lawton and Putz, 1988). Standing mortality often leads to gaps that may best be described as gradual gaps (Krasny and Whitmore, 1992), formed by slowly dying trees that drop limbs one by one or by uprooting or snapping after the tree has died. Accordingly, some of the percentages above are probably underestimates of the role of standing mortality in gap creation. Although there are few sites for which data are available, the highest rates of standing mortality appear to be in equatorial Southeast Asia (Table 8.3). Comparably high rates of standing mortality have also been reported in two temperate forests (Krasny and Whitmore, 1992; Midgley et al., 1995). Data suggest that gap formation rates are also lower in equatorial Southeast Asia (see p. 226 above; see also Putz and

233

Milton, 1982). Rates of standing mortality should be highest where wind gusts are not frequent and strong trees are not architecturally prone to structural failure, trees are deeply rooted, and rare and severe droughts occur. From Table 8.3, it appears that soil fertility has little effect, although the site with the highest rate of standing mortality has low soil fertility (Hall, 1991). Average annual rainfall appears to have no effect, although Hall (1991) attributed high rates of standing mortality to periodic moisture deﬁcits in freely draining soils. Such moisture deﬁcits can occur quickly under certain soil conditions (Richards, 1996). The locations of sites (Table 8.3) examined by Hall (1991) did not appear seriously affected by the 1982– 1983 drought that caused such severe mortality in east Bornean forests. Lack of frequent strong gusty winds may also help explain high rates of standing mortality in equatorial Southeast Asia (P. Hall, pers. commun., 1997), while high rates of tree snapping in the Venezuelan Amazon (Uhl, 1982b) may account for the lowest proportion of standing mortality in Table 8.3. Wind appears to be an important cause of tree-fall in Barro Colorado Island and La Selva (p. 227; see also Brokaw, 1985b), forests with low rates of standing mortality. Different rates of standing mortality among forests may also be related to differences among species in propensity to die standing (Seth et al., 1960; Brokaw, 1985b; Jans et al., 1993; Condit et al., 1995). In a French Guianan forest dominated by species with pyramidal crowns, standing death was more common, while in forests with predominately umbrella-shaped crowns, wind-throw was more common (Ri´era, 1995). Senescence (Swaine et al., 1987), lightning strikes (Putz and Appanah, 1987; Lawton and Putz, 1988; Bruenig, 1989; Smith et al., 1994; Magnusson et al., 1996), defoliation by herbivores (Whitmore, 1975), fungal pathogens (Whitmore, 1975; Swaine et al., 1987) and termites (Putz and Appanah, 1987) also cause standing mortality. Vines and epiphytes High abundance of woody vines may increase gap frequency, while high frequencies of vine interconnections among adjacent trees may increase gap size (Putz, 1984; see also Table 8.2, Prediction 6). Woody vines appear to increase the likelihood that a tree will fall or die due to increased mechanical stress on the shoot and to shading of the tree’s foliage (Putz, 1984). In a somewhat different way, epiphytes may increase rates of gap formation by adding weight to the

234

Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

canopy of trees in wet montane forests, especially in the wet season (Strong, 1977; Matelson et al., 1995). Vines may increase gap size by pulling down trees that are either linked with the gap-making tree or are hooked by woody vine tangles on the falling tree (Putz, 1984). There is ample evidence for this effect from vine-cutting experiments preceding logging operations (Putz, 1985). In contrast, woody vines may also have a dampening effect on tree-fall frequency and gap size by stabilizing trees through their connections with neighbors (Putz, 1984). Vine abundance is highest in areas subject to disturbance (Webb, 1958; Whitmore, 1974; Hegarty and Caball´e, 1991) and may be least in forests on nutrient-poor soils (Grubb, 1989). Dry to wet forest gradient It has been suggested that tropical dry forests should be less dynamic than tropical wet forests (Hartshorn, 1978; S.H. Bullock, pers. commun., 1996). But there have been few studies in which gap size and disturbance frequency have been described for a forest receiving less than 2000 mm of rainfall. Jans et al. (1993) and Dickinson et al. (1999) provide evidence suggesting that rates of gap formation and gap sizes are lower in forests with long dry seasons (see above, p. 226). The pattern of decreasing gap disturbance rates along the wet to dry forest gradient is obviously not well established, but several hypotheses have been advanced as to why drier forests should have lower rates of gap disturbance than wetter forests (see Table 8.2, Prediction 9): (1) Trees in drier forests have smaller crowns which produce smaller canopy gaps (Holdridge et al., 1971). However, as forest height decreases, it takes less of a gap to create an equivalent light gap, as gap aperture is the key variable (Canham et al., 1990; Lawton, 1990). Also, light limitation in the understory of the shortest forests may not be the most serious problem for tree regeneration (Lieberman and Li, 1992) so that light gaps lose their importance. Below-ground gaps may become effectively smaller for a given gap aperture, because of higher root/shoot ratios in drier forests (Cuevas, 1995). (2) Trees become shorter for a given trunk diameter as annual rainfall decreases and seasonality increases (Kira, 1978). Short trees with thick boles should be less subject to snapping (Putz et al., 1983), and shorter boles present a shorter moment arm that should lead to decreases in uprooting. Accordingly, rates of gap formation should be lower.

(3) Trees more often die standing in drier forests, leading to smaller tree-fall gaps (S.H. Bullock, pers. commun.). High rates of standing mortality would be expected on drier sites because of better anchorage, relative low height/diameter ratios, shorter periods during which gap-causing agents operate, and relatively low rain-loading of the tree canopies. Although standing mortality has been found to be important in forests that receive 2000 mm rainfall per year or less (Vooren, 1986; Putz and Appanah, 1987; Jans et al., 1993), it is also a feature of wet and moist forests, and there are too few data from drier forests to assess this hypothesis (Table 8.3). Standing mortality is associated with droughts in dry and moist tropical forests (Whigham et al., 1990; Swaine, 1992; Condit et al., 1995). Evidence from the 1982– 1983 El Ni˜no Southern Oscillation event support the notion that forests experiencing a yearly dry season may be less affected than wetter forests (Richards, 1996). Condit et al. (1995) reported 3% mortality (versus 2% during a non-drought period) on Barro Colorado Island, Panama, a forest where there are three months with less than 100 mm of rainfall. In contrast, mortality rates were substantial in forests in which rainfall does not typically drop below 100 mm in any given month. Woods (1989) reported 12–28% mortality in previously logged wet forest in Sabah, Malaysian Borneo. Similarly, Leighton and Wirawan (1986) reported 37% and 71% mortality of large trees (¾ 60 cm dbh) on two ridge plots, and 39–60% mortality of large trees (¾ 50 cm dbh) on mostly dry ridges at another site, in East Kalimantan on the island of Borneo. On wetter alluvial soils, mortality was approximately half as great as on ridges (Leighton and Wirawan, 1986; see also Ashton et al., 1995). Severe droughts of the magnitude of the 1982–1983 event recur in east Borneo at intervals of ﬁfty to several hundred years (Leighton and Wirawan, 1986). (4) Rates of tree mortality are lower in drier forests, therefore rates of gap formation are lower. In contrast to this expectation, Lugo and Scatena (1996) found no relationship between relative mortality (% yr−1 ) and rainfall in a large sample of sites (see also Swaine, 1992). As disturbance-regime data are usually area-based, we used data from Phillips and Gentry (1994) to test for a relationship between rainfall (range 1500–4746 mm yr−1 ) and the number of stems that died per hectare per

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

year. As with relative mortality, no relationship was found (F (2,28) = 0.35; R2 = 0.02; P = 0.71). (5) Drier forests may burn periodically (Uhl et al., 1988b; Swaine, 1992; Snook, 1993), leading to long periods without appreciable formation of treefall gaps as forest structure recovers (Dahir and Lorimer, 1996). (6) Rates of gap formation are lower in drier tropical forests owing to a shortened wet season, while gap-causing agents (such as gusty wind, rainfall, and lightning strikes) are operative over longer periods in forests that receive high rainfall and are aseasonal. (7) Less rain-loading of tree canopies occurs in dry forests because of lower epiphyte biomass (Holdridge et al., 1971). Lower rates of gap formation should result.

Catastrophic wind disturbances Blow-downs and wind-throws in which a number of trees are damaged or killed are larger disturbances than those created by the death of single trees (Table 8.1). Art (1993) has deﬁned blow-downs as “an extensive toppling of trees by wind within a relatively small area, greatly altering the small-scale climate within the ecosystems”. Blow-downs and wind-throws appear to be fairly common in tropical and subtropical forests, but clearly there have been too few quantitative studies to characterize their importance in tropical forests, even though their impacts last for a long time and clearly inﬂuence forest dynamics (Hubbell and Foster, 1986). Hurricanes and typhoons are the most destructive types of tropical windstorms. They occur in all regions of the world, and their impacts can be devastating (e.g., Yih et al., 1991). Many areas with vast expanses of tropical forest (e.g., South America, Africa, large areas in Asia and Australasia), however, are not impacted by hurricanes and typhoons (Fig. 8.1). In recent years, there have been a large number of articles in which the impacts of hurricanes and typhoons are considered and several authors have written summary articles (e.g., Brokaw and Walker, 1991; Tanner et al., 1991; Smith et al. 1994; Everham and Brokaw, 1996). While it is possible to predict an average return time for hurricanes within tropical and subtropical areas (Everham and Brokaw, 1996), there are few examples where it has been shown that the structure and dynamics of the forest are strongly inﬂuenced

235

by periodic hurricane events (Roth, 1992; Zimmerman et al., 1994) as much as or more than by background canopy gaps. It has been predicted, however, that forest structure and ﬂoristics would be inﬂuenced by repeated hurricane or typhoon damage (e.g., Odum, 1970; Fraver et al., 1998). Coastal mangroves may represent one of the few types of tropical and subtropical forests in which periodic wind disturbance has a dominant inﬂuence on the physical structure of the canopy (Roth, 1992). Predictions about the distribution and intensity of hurricanes and typhoons may mean little when one is considering a particular area of tropical forest, because the damage effects are greatest near the center of the storm, and there are few examples to demonstrate that speciﬁc forests have been heavily damaged by hurricanes at a frequency equaling the regional return frequency. Three examples demonstrate this point. Hurricane Gilbert heavily damaged a forest in the northeast Yucatan Peninsula in 1988 (Whigham et al., 1991; Harmon et al., 1994; Whigham et al., 1998). The northeast Yucatan was struck by 50–60 hurricanes between 1886 and 1968 (Alaka, 1976) which would represent a return time of approximately 1.6 years for the region. Our studies have been ongoing since 1984. Hurricane Gilbert has been the only storm during that 12-year period to do any signiﬁcant damage to the forest. Long-term residents could not remember any storm that damaged the local forests as much as Hurricane Gilbert did. Records of hurricanes passing through the Sian-Ka’an Biosphere Reserve, southcentral Yucatan Peninsula, show the same pattern. The area encompassed by the reserve has been struck by 11 hurricanes between 1893 and 1982 and each of them followed a different path (L´opez Ornat, 1983), indicating that few areas would have been heavily damaged more than once or twice during the period of approximately 90 years, even though the return interval, based on long-term records for the area, would be about nine years. The Luquillo Experimental Forest in Puerto Rico is one of the few areas where long-term climatological and vegetation data are available to evaluate the occurrence and impact of hurricanes. The northeastern portion of Puerto Rico was struck by approximately 45 hurricanes between 1886 and 1968 (Alaka, 1976) for a return frequency of 1.8 years. Heavy to moderate damage to the Luquillo Forest occurred three times between 1928 and 1932 (a 1.3-year return interval) but the next damaging storms did not occur until 24 years (1932–1956)

236

Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

and 33 years (1956–1989) had passed (Zimmerman et al., 1994). Since Hurricane Hugo in September 1989, four hurricanes have passed near the Luquillo Forest, but none has caused any signiﬁcant damage (J. Zimmerman, pers. commun., 1996). The extent of hurricane and typhoon impacts can range from very small to very large, and the impacts in any particular location will be inﬂuenced by a number of factors, both biotic (e.g., stem size, stem condition, species, presence of pathogens, structural complexity of the stand) and abiotic (e.g., storm intensity, topography, soil characteristics, disturbance history) (Everham and Brokaw, 1996; Imbert et al., 1996). Topography, for example, has been shown to be important in the Luquillo Forest in Puerto Rico where the landscape is rugged compared to the more uniform landscape of New England (Lugo, 1995). The factors which mediate the effects of wind as a cause of background canopy disturbance may often be obscured in intense hurricanes and typhoons (Everham and Brokaw, 1996).

COMMUNITY AND ECOSYSTEM RESPONSES TO CANOPY DISTURBANCE

Background canopy gaps, blow-downs, and catastrophic wind events in tropical and sub-tropical forests create a continuous range of conditions that set the stage for community and ecosystem response (Lugo and Scatena, 1996; Vandermeer et al., 1996). We ﬁrst describe community-level responses by focusing on both changes in species composition and pathways of initial response to mainly catastrophic wind disturbance. We then compare responses between background canopy disturbances and catastrophic wind disturbances. We close this section by discussing changes in ecosystem processes, associated primarily with catastrophic wind disturbance events that cause large changes in the physical structure of forests. Community-level responses A framework for understanding responses The interplay between characteristics of the disturbed site (see p. 228 above), species availability, and post-disturbance species performance determine changes in species composition and relative abundance immediately following the disturbance, and set the stage for long-term successional changes (Pickett et al.,

1987; Pickett and McDonnell, 1989; Everham and Brokaw, 1996; Vandermeer et al., 1996). Species availability (Pickett and McDonnell, 1989) after a wind disturbance includes propagules (seeds for most species) that are dispersed into a site at the time of the disturbance event, occur in a seed bank (generally in the soil in tropical forests; Garwood, 1989), or are dispersed to the site after the event from reproductive individuals that either occur outside the area of disturbance or were present in the disturbed area and recovered from damage incurred during the disturbance (Noble and Slatyer, 1980; Everham and Brokaw, 1996). Species availability also includes intact or damaged adults and juveniles that were present at the time of the disturbance. Sprouting of persistent meristems from roots and stems of intact or damaged plants is an important component of recovery (Everham and Brokaw, 1996). Species performance (i.e., establishment and growth of individuals that originated from seed or persisted through the disturbance event) is determined by the autecology of the species (e.g., germination, growth, and assimilation patterns), environmental conditions (e.g., light and soil moisture) prior to and following the disturbance, and species interactions (e.g., disease, herbivory, and competition; Pickett and McDonnell, 1989). Throughout the literature on responses to catastrophic wind disturbances and background canopy gaps, species have been grouped by relative shade tolerance (Brokaw, 1985b; Clark and Clark, 1987; Denslow, 1987; Swaine and Whitmore, 1988; Brokaw and Scheiner, 1989; Everham and Brokaw, 1996). The grouping of plants into shade intolerant and shadetolerant species necessarily involves breaking up a continuum of conditions and responses, as well as lumping species with different life-history attributes. Thus there is room for much reﬁnement (Clark and Clark, 1992; Grubb, 1996) in categorizing species responses. We believe that the most useful approach for comparing species responses to wind disturbances is to differentiate species that have relatively abundant seedlings and saplings below a closed canopy (shadetolerant) from those that may persist for relatively short periods below a closed canopy, but grow very little, if at all, beyond their seed reserves (shade-intolerant: Clark and Clark, 1987; Swaine and Whitmore, 1988). Pioneer species (Whitmore, 1989; Clark and Clark, 1992; Kennedy and Swaine, 1992) are a subset of the shade-intolerant group of species, and are short-lived, grow quickly in high light, and have light wood.

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

237

Table 8.5 Predictions about absolute change in species composition and the prevalence of different pathways of response to catastrophic wind disturbance in tropical forests #

Prediction

Species composition 1

Absolute change in patterns of species relative abundance (species composition) should be greatest where catastrophic wind events have the longest return interval

2

Large changes in species composition should be associated with catastrophic wind events that damage forests established on old ﬁelds or plantations

3

Little change in species composition should occur in forests frequently hit by catastrophic winds, because of dominance by species that resist and are resilient to damage (i.e., survive well and sprout readily)

Response pathways 4

Regrowth from damaged stems predominates after catastrophic wind events in the tropics

5

The importance of regrowth decreases with increasing mortality rates, damage, and uprooting

6

The importance of regrowth should be greatest in wet sites within humid tropical forests, in high-altitude tropical forests, and in dry tropical forests

7

The importance of release of understory trees after catastrophic wind disturbance is currently underestimated, primarily owing to a lack of long-term studies of regrowth

8

Recruitment of short-lived shade-intolerant species with dormant seeds should be greatest where mortality and damage is greatest and uprooting more common

9

Recruitment of short-lived shade-intolerant species should be minimal in forests with a long return time for catastrophic events

10

Recruitment of shade-intolerant species with non-dormant seeds should be minimal in catastrophic wind disturbances owing to reductions in populations of seed dispersers and nonexistent or inconsequential post-event fruiting

11

Repression by vines of tree release and recruitment should occur where damage is severe and frequent

12

Repression by vines should affect species composition of recovery owing to differential abilities among tree species to shed and avoid vines

Shade-intolerant species are most likely to regenerate from seeds that germinate at the time of gap formation (Lieberman and Lieberman, 1987; Kennedy and Swaine, 1992), while the shade-tolerant species are more likely to colonize a gap from established seedlings or saplings (Brokaw, 1985b; Brokaw and Scheiner, 1989; Connell, 1989; Brown and Whitmore, 1992) or from sprouts of damaged seedlings or saplings (Putz and Brokaw, 1989). This may often be primarily a function of mortality rates below a closed canopy (see Lieberman et al., 1990). The response of shadeintolerant species to wind disturbances is further inﬂuenced by whether the species have or do not have a dormant seed bank (e.g., Garwood, 1989). Species that do not maintain a dormant seed bank can only respond to a wind disturbance if seeds are recruited to the site from individuals at the site that produced seed shortly before or after the disturbance, or from reproductive individuals not at the disturbance site but close enough for seeds to be dispersed by wind or animals.

Responses to catastrophic wind disturbance Major pathways of response to catastrophic wind disturbance: Responses to catastrophic wind disturbances have been discussed in terms of recovery pathways (Everham and Brokaw, 1996), including: (1) regrowth from sprouting of damaged stems; (2) release of established seedlings and saplings of primarily shade tolerant species that were present in the understory at the time of disturbance; (3) recruitment of seedlings of primarily shade-intolerant species from dormant seeds in the soil seed bank, or from seed dispersal to the site just before or following the disturbance; and (4) repression of vegetation by fast-growing species of vines, herbs, and shrubs. Table 8.5 lists 12 generalized predictions about community responses to catastrophic wind disturbance. In this section we discuss the basis for each of the predictions. We expect exceptions to the generalizations, because they are based on a review of the literature which includes ﬁndings from catastrophic wind events of widely different intensities and from forests that differ in disturbance history, soils, and

238

Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

type (e.g., wet and dry forests, lowland and montane, temperate and tropical). Change in species composition: Absolute change in relative abundance patterns (species composition) in response to a catastrophic wind event should depend, in large part, on the characteristics of the disturbed site and on the species composition of the forest before the event, both of which are related to the severity, frequency, and recency of past events (Richards, 1996; see also Table 8.5, Predictions 1–3). The greatest change in species composition has been predicted to occur in forests that are infrequently impacted by catastrophic wind disturbances. This response is mostly owing to the increased importance of shade-intolerant species which often have low abundance prior to the disturbance and have increased opportunities for recruitment following the disturbance (Richards, 1996). Wind damage in forests with a low frequency of disturbance should be greater because canopy trees are taller and the topography of the forest canopy is more irregular and thus more prone to wind damage (Foster and Boose, 1992; Poorter et al., 1994; Everham and Brokaw, 1996; Richards, 1996). An exception to this prediction occurred in Nicaragua where storms are infrequent, hurricane damage was severe, and recruitment was limited (Boucher et al., 1994; Vandermeer et al. 1995, 1996). The absolute change in species composition would not be expected to be great in forests that suffer frequent wind disturbance, where species are better adapted to resist and respond to damage (Frangi and Lugo, 1991; Richards, 1996). Large changes in species composition have been shown to occur in the temperate zone when a canopy of relatively shade-intolerant species is heavily disturbed and replaced by shade-tolerant species that were present in the understory (Spurr, 1956; Webb, chapter 7, this volume). This general process has been termed accelerated succession (Spurr, 1956; Abrams and Scott, 1989). In contrast, catastrophic wind events in the tropics are often characterized by little change in species composition (Everham and Brokaw, 1996). Differences between tropical and temperate forests may be due, in part, to the fact that many of the temperate forests in which studies have been done developed on sites that had been previously disturbed by logging or had been cleared for agriculture and then abandoned (Spurr, 1956; Foster, 1988). The same type of response may be anticipated in the tropics where plantations (Lugo, 1992; Parrotta, 1995; Fu

et al., 1996) and secondary forests (Lugo, 1992) often have a diverse understory differing in composition from the canopy trees and where plantations often are more severely damaged during catastrophic events than natural primary and secondary forests (see Everham and Brokaw, 1996; Fu et al., 1996). Plantations or natural forests dominated by fast-growing, soft-wooded species would be highly susceptible to wind damage and resulting mortality (Putz et al., 1983; Zimmerman et al., 1994; Fu et al., 1996) and should show low levels of regrowth (Zimmerman et al., 1994). Relatively little change in species composition would be expected where species are resistant and resilient to damage. The importance of resistant and resilient species should increase in forests that are frequently damaged by wind (Brokaw and Walker, 1991; Frangi and Lugo, 1991; Walker, 1991; Bellingham et al., 1994; Zimmerman et al., 1994; Matelson et al., 1995; Everham and Brokaw, 1996; Scatena and Lugo, 1995). Hard-wooded, shade-tolerant species have been found to survive damage well and to sprout more readily than soft-wooded species (Putz et al., 1983; Zimmerman et al., 1994). Also, palms appear to resist wind damage and readily recover from defoliation (Frangi and Lugo, 1991). Forests that experience frequent windstorms, but that are not severely damaged owing to the dominance of resistant and resilient species, should be dominated by shade-tolerant species with seedling and sapling pools (Richards, 1996). Limited opportunities for recruitment in gaps would result from both reduced damage to established trees and a pulsed pattern of disturbance with many years between events, rendering inviable a life-history involving short life-span and recruitment from seed in gaps (Noble and Slatyer, 1980). Also, as we argue below (pp. 239–241), there are several general barriers to recruitment in tropical cyclonic-storm disturbances that would appear to make recruitment a less than ideal regeneration strategy. Regrowth by sprouting: In their review, Everham and Brokaw (1996) suggested that regrowth from surviving stems predominates following catastrophic wind disturbance in the tropics because most events cause low to moderate damage and mortality (Table 8.5, Predictions 4–6). Sprouting occurs after mechanical damage to individuals that were damaged by wind, by falling debris, or by the fall of large individuals. Several variables appear to inﬂuence the importance of recovery by regrowth. Everham and Brokaw (1996) predicted that regrowth would become less important as direct

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

or indirect mortality from wind disturbance increased. Also, the importance of recovery by regrowth may decrease as the proportion of uprooted trees increases; uprooted stems are, in some cases, less likely to sprout than snapped trees (Putz et al., 1983; Bellingham et al., 1994; Everham and Brokaw, 1996). On moister sites, sprouting appears to be a common tree response, but increases in density due to sprouting may be negated by increased mortality of uprooted trees, which appears to be common on wet sites (Everham and Brokaw, 1996). The importance of stem sprouting as a major pathway of recovery following wind disturbance also appears to be greater in drier forests and in forests at higher elevations (Ewel, 1977; Ewel, 1980; Murphy and Lugo, 1986), although this tendency has not been evaluated in terms of forest response to hurricane damage. Release: Our predictions regarding release as an important recovery mechanism are based on only a few studies of this phenomenon in tropical forests (Table 8.5, Prediction 7). There are at least four reasons why the importance of release after catastrophic wind events is likely to be underestimated. First, for seedlings and saplings, the difference between release and regrowth is trivial; to be in either response category, the plants had to be present before the disturbance event. Damaged seedlings and saplings would fall into the regrowth category, while individuals that were not damaged would arbitrarily fall into the release category. Second, we expect that short-term studies of regrowth seriously overestimate the eventual importance of sprouts, because of the development of disease associated with wounds and a concomitant decrease in wind-ﬁrmness (Roth and Hepting, 1943; Shigo, 1984; Putz and Sharitz, 1991). Large trees often cannot recover effectively after severe damage (Oldeman, 1978), and may be very likely to die, even long after the event that damaged them (Shaw, 1983; Walker, 1995). Third, catastrophic wind disturbance may initiate changes in canopy dominance, beginning with an increase in the importance of short-lived pioneers which, after several decades, are replaced by shade-intolerant species that persist until the next disturbance event (Weaver, 1986). The fourth reason why release may be underestimated comes from evidence that catastrophic events appear to lead to enhanced growth of suppressed seedlings and saplings. Defoliation and structural damage to canopy trees results in an increase in the amount of light in the understory, at least for a couple

239

of years (Fern´andez and Fetcher, 1991; Bellingham et al., 1996). In Nicaragua, although the majority of stems were sprouts, released seedlings and saplings appeared to account for a large proportion of the stems encountered in post-hurricane inventories (Yih et al., 1991; Boucher et al., 1994; Vandermeer et al., 1995). In Puerto Rico, falling litter and debris from Hurricane Hugo killed 60% of the seedling pool of one species, but the remaining seedlings of that species responded with a strong increase in growth (You and Petty, 1991). An increase in growth in smaller trees following hurricane disturbance was observed in the Yucatan following Hurricane Gilbert (Whigham et al., 1991). The abundance of the long-lived Shorea parvifolia in the storm forest of Kelantan disturbed in the late 1800’s may be an example of an important episode of release (Wyatt-Smith, 1954). Release has also been found to be important in temperate forests recovering from catastrophic wind disturbance (e.g., Spurr, 1956; Foster, 1988; Platt and Schwartz, 1990; Merrens and Peart, 1992; Webb, chapter 7, this volume). Even if regrowth of damaged stems is more important than release among the set of individuals that successfully captures space in the canopy after a given disturbance event, a generalized release response in past disturbance events will still have played a central role in that prevalence of regrowth (Foster, 1988; Connell, 1989; Clark and Clark, 1992; You and Petty, 1991). Recruitment: Recruitment rarely appears to be the dominant recovery pathway following catastrophic wind disturbances (Everham and Brokaw, 1996; Table 8.5, Predictions 8–10). Recruitment, nonetheless, is generally a component of recovery (Weaver, 1986; Frangi and Lugo, 1991; Walker, 1991; Bellingham et al., 1994). Recruitment should be greatest following wind disturbances that cause extensive damage and mortality, particularly in forests that are infrequently impacted by major storm events (Everham and Brokaw, 1996; Richards, 1996). Recruitment is, however, likely to vary spatially within the area impacted by wind disturbances. The greatest damage is more likely to occur near the center of the storm than at the periphery (Lugo et al., 1983; Richards, 1996), and recruitment would thus be more important near the storm center. Recruitment would also be expected to be more important in topographically exposed sites that are more heavily damaged (Bellingham, 1991; Frangi and Lugo, 1991; Foster and Boose, 1992); particularly

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Dennis F. WHIGHAM, Matthew B. DICKINSON and Nicholas V.L. BROKAW

at higher-elevation sites where damage is more severe and less spatially variable (Beard, 1945a,b; Brokaw and Grear, 1991). Several examples in which high levels of recruitment have occurred seem to counter the general perception that regrowth is the primary recovery mechanism following wind disturbance. Whitmore’s (1974) 21-year study of the role of catastrophic wind disturbance in tropical forests contrasted sites on Kolombangara, in the Solomon Islands, that were subject to severe damage by past hurricane events with those that had been more sheltered from hurricane damage. The canopies of the more hurricane-prone areas were dominated by long-lived shade-intolerant species while the sheltered sites had a high representation of nonpioneer species. The shade-intolerant species suffered heavy damage during disturbance events, yet recruited well afterwards from seed. Snook (1993) also reported successful recruitment by long-lived shade-intolerant species after hurricanes in the Yucatan Peninsula of Mexico. Lack of recruitment by shade-intolerant species following catastrophic wind disturbances may be related to the presence of few large gaps, lack of exposed soil, the presence of a thick litter layer, and lack of propagules (Everham and Brokaw, 1996; Walker, 1999). Catastrophic wind events create a range of gap sizes dominated by small gaps, although the size distribution shifts to a larger average gap size as storm intensity increases (Everham and Brokaw, 1996). Except for patches that are severely damaged, the canopy may not remain open long enough for pioneers to avoid suppression, as regrowth following defoliation and minor branch damage can be rapid (Bellingham et al., 1994). Perhaps more important than a lack of large gaps is the high deposition of litter and lack of disturbance to the substrate and understory necessary for germination and establishment of pioneer seedlings (Weaver, 1986; Everham and Brokaw, 1996). “Litter gaps”, where the litter layer is removed and mineral soil is exposed, are thought to be indispensable for high rates of germination and establishment of small-seeded species in general and pioneers in particular (Putz, 1983; Putz and Appanah, 1987; Raich and Christensen, 1989; Kennedy and Swaine, 1992; Molofsky and Augspurger, 1992; Grubb, 1996). Litter gaps are often associated with uprooted trees, nurse logs, and steep slopes (Grubb, 1996). Uprooting mixes the soil, which may enhance germination rates (Putz, 1983; Putz and

Appanah, 1987). Experimental litter removals after a hurricane were particularly beneﬁcial to pioneer species (Guzm´an-Grajales and Walker, 1991). The importance of litter gaps is supported by the ﬁnding that ﬁres that follow hurricanes, by removing litter, lead to high levels of establishment of pioneer species, particularly if ﬁre intensity is high enough to kill vegetation that would otherwise sprout profusely (Oliver and Larson, 1990; Snook, 1993; Everham and Brokaw, 1996). Even if conditions are right for germination, growth and survival of shade-intolerant seedlings, dormant or newly dispersed seeds may not be available (Schupp and Fuentes, 1995; Everham and Brokaw, 1996). Lack of a response to major disturbance by short-lived shadeintolerant species has been attributed to the absence of both a seed bank and reproductive individuals owing to a long disturbance return time or the spatial pattern of disturbance (Noble and Slatyer, 1980; Uhl et al., 1988a; Brand and Parker, 1995; Peterson and Carson, 1996). One possible example of lack of response by short-lived shade-intolerant species owing to lack of propagules were the severely hurricane-damaged Nicaraguan forests described by Boucher et al. (1994) and Vandermeer et al. (1995). The hurricane return time for the area is long, perhaps too long for adults and seed banks to persist between events. Because major wind disturbances should signiﬁcantly reduce background rates of canopy gap formation (Dahir and Lorimer, 1996), there may be few opportunities for regeneration between events. Unique problems of seed input may be associated with catastrophic winds. When defoliation is extensive, seed fall just after the event can be minimal (Lindo, 1968), although defoliation has been reported to trigger a large ﬂowering response, at least in some species, and lead to an increase in seed fall compared to normal levels (Everham and Brokaw, 1996; Lugo and Scatena, 1996). This later seed rain may often be irrelevant, however, because of a premium on early establishment (Brokaw, 1985a; Brown and Whitmore, 1992; Kennedy and Swaine, 1992) and because the seed bank is likely to be the predominant source of viable seeds in background canopy gaps (Garwood, 1989) where defoliation does not occur. In large gaps created by catastrophic winds, seed rain may be further constrained among wind-dispersed pioneer species by the distance to potential seed sources. A large proportion of species have animal-dispersed seeds (Swaine and Whitmore, 1988; Levey et al.,

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

1994), and the dispersers often avoid large gaps in the canopy (Schupp et al., 1989). Seed and fruit dispersers may leave hurricane-damaged areas because of a lack of seeds and fruit (Ackerman et al., 1991; Yih et al., 1991). One might expect that, where hurricane damage is most severe, ﬂowering and fruiting would be most attenuated. This would particularly affect those species that do not have dormant seeds, a seeding and sapling bank, or a strong sprouting response, but might have little effect on pioneers. If this is the case, it could partially explain the lack of an overwhelming recruitment response in some severely damaged forests (Boucher et al., 1994; Vandermeer et al., 1995, 1996). Repression: Long-term repression of recovery as a result of the proliferation of herbs and vines (Everham and Brokaw, 1996) appears to be infrequent in the tropics (Table 8.5, Predictions 11, 12). In high-altitude dwarf forests in Puerto Rico, recovery was dominated by ferns and grasses (Weaver, 1986; Walker et al., 1996), although the cause of a lack of tree recovery was not clear. In gaps in Samoan lowland forest created by catastrophic winds, dense growth of grasses and ferns suppressed tree regeneration (Wood, 1970). These would seem to be examples of the arrestedsuccession effect, where shrub and herb communities delay invasion by trees in the temperate zone (Putz and Canham, 1992). The climbing habit of vines, however, creates a different situation. Repression by vines has been documented in several severely and frequently disturbed forests. Frequent hurricanes in wet forests of northeastern Australia can lead to the formation of “cyclone scrub”, a short-statured forest with emergent climber towers and abundant vines throughout the canopy (Webb, 1958). These forests occur where winds from frequent storms are locally intensiﬁed by topography (Webb, 1958). A similar situation occurs in Nigeria as a result of frequent tornados (Jones, 1955a,b). Recruitment and release can be suppressed by vines in large blow-downs (Lindo, 1968; Wood, 1970) and after intensive logging (Putz, 1991) where forest structure is conducive to vine proliferation (Hegarty and Caball´e, 1991; Putz, 1985). Soil disturbance owing to logging promotes high levels of vine establishment (Putz, 1985, 1991). Proliferation of vines appears to delay forest structural development, but, except for frequently disturbed forests (Webb, 1958), the effect is temporary (Wyatt-Smith, 1954; Webb, 1958; Whitmore, 1974; Whitmore, 1975). Putz (1980) listed several characteristics that enable trees

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to avoid and rid themselves of vines, these include: fast height and diameter growth, regular self-pruning of branches, large compound leaves, and vine-removing symbionts. The ability to avoid and shed vines provides a mechanism by which vine repression may alter the species composition of trees following severe damage (Table 8.5). In support of this general expectation, vine tangles in big gaps (¾ 1 ha) in Costa Rica exclude “all but the very fast growing pioneers” (Hartshorn, 1980). Also, palms are “climber shedders/vine tangle surmounters par excellence” owing to their methods of leaf production and shedding (Putz, 1980). It is apparent that not all severely damaged forests are prone to repression by vines (Walker et al., 1996). One possible example of this is the “hurricane forests” of the Caribbean island of St. Vincent, where damage is frequent and uprooting predominant, but where vine proliferation is apparently absent (Beard, 1945b). If blow-down is not prevalent, severe winds and major branch loss may rid trees of vines. In the forests damaged by Hurricane Gilbert in northeastern Yucatan (Whigham et al., 1991), almost all vines were eliminated from canopy trees, but they survived the disturbance and within a month most began to branch profusely. For the ﬁrst year after the disturbance it appeared that vines would play an important role in recovery, but within ﬁve years almost all of the vine branches that were marked in 1988 had died, and most of the trees were still free of vines even though some had heavy vine loads prior to the hurricane (D. Whigham and E. Cabrera, pers. observ.). This could potentially explain the lack of vine repression in some relatively wind-resistant forests. Contrasts between responses to catastrophic wind and other canopy disturbances It should be apparent that much of the literature upon which an understanding of response to catastrophic wind disturbance is based comes from studies of background canopy disturbance, and response to background canopy gaps is often a good model for understanding response to catastrophic wind events (Ackerman et al., 1991; Everham and Brokaw, 1996). However, Ackerman et al. (1991) suggested that response to background canopy disturbance is an imperfect analog for response to both the least and the most severe damage by tropical cyclonic storms, while moderate damage should most closely approximate background canopy disturbance. Comparisons are made difﬁcult, in part, because recovery in background

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canopy gaps is typically studied only where gaps extend into the understory (Brokaw, 1982a; Popma et al., 1988; Lieberman et al., 1989). In contrast, in studies of catastrophic wind disturbance, gaps (or heavily disturbed patches) are not generally distinguished from areas where only upper canopy damage occurs (Frangi and Lugo, 1991; Walker, 1991). We are not aware of any studies which explicitly compare regeneration in sites dominated by background canopy disturbance with that in forests damaged by catastrophic events [(the closest would be that of Whitmore (1974)]. Forests whose disturbance regimes are dominated by catastrophic wind events should differ from forests dominated by background canopy disturbance in the frequency of large gaps, in the lack of seed sources for various reasons, and in poor recruitment even where seeds are present. Foster and Boose (1992) and Everham and Brokaw (1996) conceived of a gradient of disturbance from small gaps, created by standing dead trees or branch falls, to larger gaps formed by the uprooting or breakage of one or many trees, and, at the extreme, to large gaps with indistinct edges created by damaged and defoliated trees during a catastrophic wind event. Less severe catastrophic wind events may cause minimal upper-canopy damage, and there may only be a few small gaps that extend to the understory (Everham and Brokaw, 1996). When compared with background canopy disturbance, catastrophic wind events can create equally small gaps as well as much larger gaps (Foster and Boose, 1992; Everham and Brokaw, 1996). For both types of disturbance, small gaps are much more frequent than large gaps, although, despite their low frequency, large gaps may cover a comparatively large area (see Lawton and Putz, 1988; Brokaw, 1985b; Foster and Boose, 1992). Timing is obviously different, with background canopy disturbances occurring continuously whereas catastrophic events occur in a pulsed and infrequent manner (Oldeman, 1989). The general assumption is that release, sprouting, and ingrowth of gap-edge trees dominate in small gaps, while recruitment becomes more prevalent in large gaps (Dunn et al., 1983; Brokaw, 1985a,b; Denslow, 1987; Raich and Christensen, 1989; Whitmore, 1989; Everham and Brokaw, 1996) in both disturbance types. Similarly, predictions and data from forests dominated by background canopy gaps (Hartshorn, 1978; Denslow, 1980; Putz and Appanah, 1987; Uhl et al., 1988a; Brokaw, 1985b; Ashton, 1981; Jans et al., 1993) and from forests damaged by catastrophic wind

events (Weaver, 1986; Runkle, 1990; Walker, 1991; Bellingham et al., 1994; Richards, 1996) suggest that it is the frequency of large gaps and severely damaged patches that should determine the relative abundance of shade-intolerant species in a forest. Where background canopy disturbance dominates in the tropics, shadetolerant species comprise by far the largest group (Clark and Clark, 1987; Whitmore, 1989; Welden et al., 1991; Lieberman et al., 1995). Few authors have found (Whitmore, 1974) or predict (Richards, 1996) a dominance by shade-intolerant species in forests damaged frequently by catastrophic wind disturbances. This may be owing to the ﬁnding that most catastrophic wind events cause only minor to moderate damage (Everham and Brokaw, 1996). Even where gap sizes are large, relatively poor recruitment responses have been shown to occur (Uhl et al., 1988a; Boucher et al., 1994; Vandermeer et al., 1995) owing to lack of propagules and post-dispersal phenomena. In a large blow-down in the Venezuelan Amazon in which uprooting predominated, Uhl et al. (1988a) ascribed a lack of pioneer response to a lack of seed sources, which, in turn, may be due to poor regeneration of these species in the prevailing small gaps (see also Peterson and Carson, 1996). Putz and Appanah (1987) have suggested, however, that a lack of propagules should not be expected to be more strongly associated with forests damaged by catastrophic wind events (Putz and Appanah, 1987). Long intervals between events in forests whose disturbance regimes are dominated by catastrophic winds would tend to reduce the relative abundance of both adults and seed pools of short-lived pioneers (Noble and Slatyer, 1980). In cyclonic storm disturbance, lack of a recruitment response may be due to propagule limitation related to defoliation and branch damage of canopy trees (Lindo, 1968) and to the lack of seed dispersers in heavily damaged areas (Ackerman et al., 1991; Yih et al., 1991). Even where propagules are present, recruitment may be reduced in catastrophic wind disturbances by a lack of disturbed and litter-free soil (Guzm´anGrajales and Walker, 1991; see review in Everham and Brokaw, 1996) and by rapid vegetative recovery that would compete with new seedlings (Everham and Brokaw, 1996). Similarly, both lack of litter-free soil in forests where uprooting is uncommon (e.g., Putz and Appanah, 1987) and competition from established individuals (Kennedy and Swaine, 1992) have been related to reduced recruitment in background canopy

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

disturbances. Lack of disturbed soil in catastrophic wind disturbances, even where mortality and structural damage rates are high, may not be as generally important as previously suggested (Everham and Brokaw, 1996): the proportion of uprooting among fallen trees ranges from 15% to 88% in tropical forests hit by catastrophic winds (Everham and Brokaw, 1996) while similar rates of uprooting (15%–90%) are reported for tree-falls in forests dominated by a background canopy disturbance regime (Table 8.3). In this section, we have provided a series of predictions (Table 8.5) regarding changes in species composition and recovery pathways following canopy wind disturbance. No clear patterns emerge, because there have not been enough long-term studies of tropical forests to elicit broad geographic patterns or determine whether or not predictions, such as those that we provide, adequately describe recovery processes following background and catastrophic wind disturbances. A general model that might be useful in designing future investigations to test these and other predictions comes from succession research on prairie pothole wetlands in the United States (van der Valk, 1981). Prairie pothole wetlands occur over a broad range of hydrogeomorphic settings in which water quality (fresh or saline) and disturbances, both natural (wet and dry cycles) and anthropogenic (agriculture, ﬁre), play important roles in determining the status of vegetation at any moment in time. Typically, vegetation undergoes dramatic changes in response to short- and long-term changes in water levels. Van der Valk based his model on “a Gleasonian approach” to understanding vegetation change at the ecosystem level, based on the ecological characteristics of individual species. He suggested that species or groups of species with similar life-history traits respond similarly to disturbance, whether of natural or anthropogenic origin. In applying this conceptual approach to tropical forests, the ﬁrst step would be to characterize the disturbance regime of the forest. Once the disturbance regime was adequately characterized, the responses of individual species (or groups of species such as shade-tolerant and shadeintolerant) to the range of disturbances would be evaluated. In prairie pothole wetlands, for example, the interactions between water levels and germination characteristics of seeds of species in the soil seed bank were useful in predicting vegetation responses. In tropical forests, the soil seed bank appears to be less important, but species responses over the range of sizes of wind disturbances (e.g., ingrowth in small

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canopy gaps, and sprouting in larger openings) appears to be most important. Ongoing long-term research in the Luquillo Experimental Forest in Puerto Rico (Zimmerman et al., 1996) should provide the types of data that will be required to test predictions such as those that we have provided in Table 8.5. Ecosystem-level responses Ecosystem-level processes (e.g., rates of decomposition, rates of nutrient uptake and release, primary production) should be inﬂuenced by wind-generated disturbances, especially in larger disturbances where changes in the physical structure of ecosystems are greater. Many changes in ecosystem-level processes are mediated through alterations of microclimate, changes in site water balance, and additions of nutrients and carbon associated with the destruction of biomass (Frangi and Lugo, 1991; Lodge and McDowell, 1991; Sanford et al., 1991; Lugo and Scatena, 1996; Scatena et al., 1996). Microclimate modiﬁcations primarily involve increases and changes in the quantity and quality of light (Raich, 1989; Brown, 1996) and associated variables such as soil and air temperatures and relative humidity (Brown, 1993). Reductions in leaf biomass, and destruction of branches and boles increase amounts of photosynthetically active radiation (Bellingham et al., 1996), decrease ﬁne-root biomass in the disturbance site (Parrotta and Lodge, 1991; Silver et al., 1996), and increase soil moisture owing to decreased evapotranspiration (Vitousek and Denslow, 1986). The degree of change in microclimatic variables following canopy disturbance is related to disturbance size (Chazdon and Fetcher, 1984; Canham et al., 1990; Bellingham et al., 1996), but there is often considerable spatial variation within the disturbance site (Brown, 1993). The microclimate changes as vegetation and the ecosystem recover from disturbance. Bellingham et al. (1996) found that photosynthetically active radiation increased with the severity of hurricane damage (i.e., amount of defoliation) but that light levels had returned to almost pre-hurricane levels within 33 months. Fern´andez and Fetcher (1991) found similar results in Puerto Rico 14 months after Hurricane Hugo. Increased light levels in single tree-fall gaps also decrease with time, but the causes of the changes would be different from those measured in hurricane-damaged forests because of differences in ecosystem recovery responses

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(e.g., ingrowth versus sprouting in the understory: Bellingham et al., 1996). There have been few studies of root biomass, root turnover, and root growth in gaps. Silver et al. (1996) found that root biomass in the Luquillo Experimental Forest (Puerto Rico) was only beginning to recover six years after Hurricane Hugo; though most of the soil nutrient pools had returned to pre-hurricane levels. Wind-generated disturbances of all sizes signiﬁcantly increase the necromass of fallen leaves, branches, trunks and upturned roots (Frangi and Lugo, 1991; Harmon et al., 1994; Smith et al., 1994). Not only does the amount of dead biomass increase in hurricaneimpacted forests (Harmon et al., 1994), but the standing stocks of nutrients associated with the dead biomass increase tremendously because of the relatively high nutrient content of the leaves and wood brought down by the disturbance (Tanner et al., 1991). Frangi and Lugo (1991) found that the amounts of nitrogen, phosphorus, potassium, calcium, and magnesium in aboveground necromass had increased by 19%, 18%, 17%, 23%, and 16% respectively, 7 months after Hurricane Hugo. Whigham et al. (1991) measured an increase of approximately 50% in coarse woody debris in a dry tropical forest following Hurricane Gilbert. Standing stocks of calcium, potassium, magnesium, and nitrogen increased between 22% and 36% and phosphorus and magnesium increased by more than 55%. The amount of leaf material deposited by Hurricane Gilbert was 29–98% higher than the total annual amounts of leaf litter-fall measured for the four years prior to the hurricane (Whigham et al., 1991). The concentrations and total amounts of nutrients in leaf litter-fall associated with Hurricane Gilbert were in many instances 100–300% higher than they had been in the four previous non-hurricane years (Table 8.6). Release of nutrients through leaching and decomposition of leaves and wood would potentially increase nutrient availability, as well as losses of nutrients to groundwater and surface water, and to the atmosphere by gas emission. There have been, however, few studies in which nutrient cycling has been examined at the ecosystem level in tropical and subtropical forests following disturbances similar to those caused by wind damage (Parker, 1985; Vitousek and Denslow, 1986; Uhl et al., 1988a; Lodge and McDowell, 1991; Walker, 1999). Hartshorn (1978), Whitmore (1978) and Orians (1982) each predicted that soil properties would change following canopy disturbance, and that tree uprooting in single tree-fall gaps would also change

the characteristics of exposed soils (Schaetzl et al., 1989). Vitousek and Denslow (1986) found higher soil moisture in tree-fall gaps, but no signiﬁcant increase in soil nutrients between gap and non-gap habitats, even though soils associated with uprooted trees had lower nitrogen, phosphorus and carbon, and higher rates of nitrogen mineralization. Parker (1985) found similar results in disturbed and undisturbed areas in the same forest type as that studied by Vitousek and Denslow. Uhl et al. (1988a) found no evidence to support the hypothesis “that treefall gaps might represent zones of high nutrient leakage” in humid tropical forests in Venezuela. There is some evidence for increased nutrient availability in larger wind-disturbed areas (Lodge and McDowell, 1991) but not as much as occurs in response to forest clear-cutting (Steudler et al., 1991). In addition, many of the changes in nutrient cycling in response to large-scale disturbances such as hurricanes appear to have little long-term impact on ecosystem function. Table 8.6 Biomass and nutrients (totals and concentrations) in leaf litter-fall for Hurricane Gilbert compared to the range of values measured in the 4 years prior to the hurricane 1 Leaf litter

Percent. increase 2

Biomass

29–98

Phosphorus Total

133–250

Concentratiion

60–166

Potassium Total

92–189

Concentration

143–382

Calcium Total

43–133

Concentration

8–17

Magnesium Total

33–100

Concentration

0–(−6)

Manganese

1

Total

50–200

Concentration

2–11

Data compiled from Whigham et al. (1991). Percent increase = 100 (x1 − x0 )/x0 ; where x0 is the mass or concentration before the hurricane, and x1 during it.

2

BACKGROUND CANOPY GAP AND CATASTROPHIC WIND DISTURBANCES IN TROPICAL FORESTS

As indicated earlier, many of the ecosystem responses related to nutrient cycling can be ascribed to large inputs of coarse woody debris and the release of nutrients through leaching and decomposition (Frangi and Lugo, 1991; Harmon et al., 1994; Smith et al., 1994). In this context, the Luquillo Experimental Forest in Puerto Rico has been the most intensively studied site (Zimmerman et al., 1996). Sanford et al. (1991) used simulation modeling to predict that an initial decrease in nitrogen mineralization would be followed by higher rates following hurricane damage. Their predictions with respect to nitrogen dynamics were veriﬁed by Lodge et al. (1991), who measured elevated soil ammonium concentrations and higher rates of nitriﬁcation 17 and 7 months, respectively, after Hurricane Hugo, and increased mineralization was later suggested by increased nitrate levels in stream water. Sanford et al. (1991) also predicted that a period of phosphorus immobilization after the hurricane would be followed in about two years by an increase in phosphorus availability. Soil phosphorus availability was not measured at the Luquillo site following Hurricane Hugo, but Lodge and McDowell (1991) used results from a long-term fertilization study at Luquillo to suggest that responses to increased amounts of phosphorus occur, but only after several years of fertilizer addition. This conclusion is supported by long-term studies of phosphorus addition to a tropical dry forest in the Yucatan (Whigham and Lynch, 1998). There was no detectable response in tree growth, leaf litter-fall biomass and nutrients to four years (1984–1988) of phosphorus fertilization of the forest (Whigham et al., 1998). One year after Hurricane Gilbert (1988), phosphorus concentrations in leaf litterfall were higher in each of four years after the hurricane than they had been during four pre-hurricane years. It is unclear, however, whether the increased nutrient levels in leaf litter-fall were due to increased phosphorus availability following four years of fertilization and/or to responses to Hurricane Gilbert. Reduced ﬁne-root biomass and added labile carbon in disturbed sites (Parrotta and Lodge, 1991) seem to inﬂuence the rates of emission of gases containing nitrogen and carbon. Steudler et al. (1991) consistently found lower rates of loss of carbon dioxide, decreased rates of emission of methane, and higher rates of nitrous oxide production in hurricanedisturbed areas than in undisturbed sites. All these trends are consistent with the hypothesis that lower oxygen levels in the soil alter microbially-mediated

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processes such as denitriﬁcation, methanogenesis, and soil respiration. We know of no data in which disturbance sites have been monitored until there is no evidence of the past disturbance. The studies of Bruenig (1989) and Whitmore (1989) suggest, however, that tropical and subtropical forests are very dynamic and that few areas remain in equilibrium for long periods of time. Zimmerman et al. (1996) provided a conceptual view of temporal patterns of recovery following catastrophic disturbance. Several authors have noted that it is possible to identify large-scale disturbance sites decades after the event occurred (e.g., Richards, 1996). Even though there have been few long-term studies of ecosystem-level processes in single treefall gaps or in areas impacted by blow-downs or hurricanes, the responses that have been measured are consistent enough to suggest that the impacts last for periods from a few months up to three years, and that ecosystem recovery occurs rapidly (Whigham and Lynch, 1998). We believe that largescale and long-term impacts to wind damaged forests, particularly those damaged by hurricanes, would only occur following massive destruction of biomass by ﬁres. Harmon et al. (1994) examined several sites in the Yucatan Peninsula, and found that sites that had been impacted by the hurricane and by ﬁre had the highest amounts of coarse woody debris, and they estimated that it would take between 30 and 150 years to return to pre-hurricane levels.

CONCLUSIONS

There is ample evidence that wind inﬂuences tropical forests, and that the loss of biomass that results directly or indirectly from wind occurs over a range of scales. In this chapter we have chosen to divide the continuum of wind disturbances into background disturbances that result in the death of one or a few trees, and catastrophic events that open larger areas of forest (e.g., wind-throws and damage caused by hurricanes). The division that we have made is, of course, arbitrary, and was done primarily to separate wind disturbances into those that received the most scientiﬁc attention (background disturbances) from those that have not been adequately studied (catastrophic disturbances). One obvious conclusion is that there is still little known about the distribution of and long-term responses to catastrophic wind disturbances, which appear to be

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widespread. Except for a series of recent studies of hurricanes (see articles in Biotropica, 23, 1991 and 28, 1996), most of the information that has been compiled about recovery from catastrophic events has been anecdotal or primarily qualitative. Recovery from catastrophic wind events can, however, be quite rapid even after extensive damage (Whigham and Lynch, 1998). With the advent of a more complete coverage of tropical areas using satellites and of improved techniques for evaluating satellite images, it should now be possible to evaluate more adequately the distribution and frequency of occurrence of catastrophic wind disturbances in tropical forests and monitor recovery from them. We focussed our discussion on elements of wind disturbance regimes that have important effects on species, community and ecosystem responses. The following general conclusions are offered: (1) Gap sizes should be largest where there is an abundance of large trees and a low rate of standing mortality, where vines are abundant, and where wind knocks down gap-edge trees. (2) Forests of small stature (dry forests, high-altitude forests, and nutrient-poor forests) should have the smallest gaps, while forests on slightly seasonal sites with adequate soil nutrient levels should have the largest gaps, because the trees are large. (3) Rates of gap formation should be highest where tree architecture is least vulnerable to wind, there is poor anchorage, and standing mortality rates are low, and in otherwise calm regions where gusty winds are most frequent. (4) Soil disturbance should be greatest where uprooting rates are highest. This should occur where anchorage is poor owing to shallow rooting and unstable soils and parent material, and where overall rates of gap formation are highest. (5) Wind is an important cause of background canopy disturbance, and appears to be the primary cause in some, although not all, forests not frequently affected by catastrophic wind and other major disturbances. (6) High rates of background canopy disturbance, as with major disturbance, should lead to a reduction in the susceptibility of the forest to damage, because when a large or vulnerable tree is felled, it is replaced by a younger individual that is less likely to be prone to damage. (7) The long-term impacts of catastrophic wind disturbances are poorly understood, especially

ecosystem-level responses, but recent longer-term studies in the Caribbean are showing a wide range of responses at species, community, and ecosystem scales. (8) Wind disturbances may be less important in parts of the Paleotropics, but many more studies are needed to verify this conclusion. (9) There is a need for long-term studies of species, community, and ecosystem recovery from wind disturbances. Most studies, to date, have focussed on the effects of wind disturbances more than on the recovery. (10) The effects of catastrophic winds, because of their long-sustained winds with speeds that are hard to predict, often cannot be predicted from what is known of the factors that inﬂuence background canopy-gap disturbance regimes.

ACKNOWLEDGMENTS

The authors would like to thank Lawrence Walker, David Foster, and Sarah Webb for reviewing the manuscript and providing many useful comments and suggestions. We would also like to note the assistance of Pamela Hall and Neil Gale. Funding for some of the research reported by Dickinson and Whigham was provided by the Smithsonian Institution Scholarly Studies Program, U.S. Man and the Biosphere Program, and the National Science Foundation. N. Brokaw’s participation in writing this chapter was supported by grant BSR8811902 from the National Science Foundation to the Institute for Tropical Ecosystem Studies, University of Puerto Rico, and to the International Institute of Tropical Forestry, as part of the Long-Term Ecological Research Program in the Luquillo Experimental Forest. Additional support for Brokaw came from the U.S. Forest Service (Department of Agriculture) and the University of Puerto Rico.

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Chapter 9

FOREST HERBIVORY: INSECTS T.D. SCHOWALTER and M.D. LOWMAN

INTRODUCTION

Herbivory is the process of feeding on any plant parts, including foliage, stems, roots, ﬂowers, fruits, or seeds. It is important in that it reduces the density of plants or plant materials, opens the canopy, transfers mass and nutrients to the forest ﬂoor, stimulates plant growth under certain conditions, and affects habitat and resource conditions for other organisms. Many studies have addressed the effects of disturbance, including ecosystem-management practices, on herbivorous insect species that are economically important (e.g., Schowalter et al., 1986; Mattson and Haack, 1987; Paine and Baker, 1993) and the effects of herbivore outbreaks on plant growth and mortality (reviewed by Schowalter et al., 1986). However, measurements of the magnitude of herbivory in different ecosystems and under different environmental conditions have used various non-comparable techniques, which hinders interpretation (reviewed by Lowman, 1995). Few studies have assessed the relationship between herbivory and disturbances or environmental changes, or the effects of herbivory on other ecosystem processes. Although loss of plant material through herbivory generally is negligible, or at least inconspicuous, periodic outbreaks of herbivorous insects can denude or kill plants of selected species over many square kilometers. This capacity to alter ecosystem structure dramatically has led to the widely-held view that herbivorous insects are biotic agents of disturbance (e.g., Veblen et al., 1994; D’Antonio et al., Chapter 17, this volume). However, insect herbivores (unlike abiotic disturbances) are an integral component of the ecosystem and respond (as do other organisms) to change in environmental conditions. Schowalter (1985) and Willig and McGinley (Chapter 27, this volume) have argued that integral ecosystem effects (e.g., stimulated

growth of non-host plants) should not be considered disturbance. We recognize that both perspectives contribute to understanding herbivory and its relationship to disturbance. Our challenge is to represent herbivores both as integral components of communities that adapt and respond to environmental change, and as agents of change that affect ecosystem structure and function in ways similar to abiotic disturbances. At issue is the threshold at which herbivory shifts from a normal trophic process to a disturbance. Normal levels of herbivory should not be considered disturbance, but, at some undeﬁned levels of intensity, scale, and frequency, outbreaks have effects on ecosystem structure and function similar to those of ﬁre, storm, drought, or ﬂood. A key aspect of herbivory is its selectivity with respect to plant species affected. Such speciﬁcity ensures a diversity of indirect effects on other components of ecosystems, in the same way that various species show different tolerances to other disturbances or environmental changes. Outbreaks typically are triggered by changing environmental conditions, especially disturbances (or their suppression), that alter the physiological condition or abundance of host plants or predators. However, outbreak populations can spread to neighboring patches, altering community composition, canopy coverage, and biogeochemical cycling processes. These herbivore effects may function as a regulatory mechanism to stabilize ecosystem processes following disturbances or during environmental changes (Schowalter, 1985). This chapter addresses relationships between disturbance and herbivory in forest ecosystems. We review measurements of herbivory, attributes of herbivore outbreaks as agents of disturbance, disturbance-related factors that trigger growth of herbivore populations, and the effects of herbivory on forest structure and

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ecosystem processes, including promotion of future disturbances. Comparison of data from temperate and tropical forests suggests general relationships between disturbance and herbivory. We note that our discussion focuses on herbivore relationships to discrete disturbances such as ﬁre, storm, drought, or ﬂood, rather than to longer-term environmental changes such as atmospheric pollutants or climate change, which are beyond the scope of this volume. Nevertheless, where appropriate, we consider these long-term environmental changes as they affect herbivore outbreaks. We also do not address exotic herbivores separately. Outbreaks of native or exotic herbivores depend on suitable environmental conditions (i.e., abundant and suitable hosts, and inadequate regulation by predators) and can have similar impact on ecosystem structure and function. Of course, exotic herbivores show a greater capacity to decimate their hosts, at least until their hostspeciﬁc predators are introduced.

TYPES AND PATTERNS OF HERBIVORY

Types of herbivory Plants are complex structures. They include nutritious parts (cytoplasm and ﬂuids) surrounded by cell walls composed of lignin and cellulose, which insects cannot digest directly. Exploitation of plant resources is inhibited further by various biochemical defenses that are toxic or interfere with digestion. Herbivorous insects have evolved morphological, physiological, behavioral, symbiotic, and other adaptations to counter these barriers. However, these adaptations typically restrict herbivorous insects to one host (monophagy) or a few species of hosts (polyphagy) with similar defensive characteristics. Herbivores affect ecosystem structure and function to varying degrees depending on the degree of speciﬁcity and the dominance status of hosts. Insects can be classiﬁed into different feeding guilds, or functional groups, based on their mode of exploiting plants for food. Speciﬁc groups of plantfeeders include chewers (consumers of foliage, stems, ﬂowers, pollen, seeds, and roots); miners and borers (which eat wood, bark, or one or more of the tissue layers between the intact upper and lower epidermis of leaves); gall-formers (which reside and feed within the plant and induce the production of abnormal growth reactions by the plant tissues); sap-suckers (which have

T.D. SCHOWALTER and M.D. LOWMAN

specialized mouthparts and survive by tapping into plant ﬂuids); and seed-eaters and fruit-eaters (which consume the reproductive parts of plants) (Romoser and Stoffolano Jr, 1994). Only seed eaters, seedlingeaters, and some tree-killing bark beetles are true plant predators; most herbivores are more correctly classiﬁed as parasites because they do not kill their hosts but feed on the living plant (Price, 1980). These different modes of consumption affect plants in different ways. For example, folivores (species that feed on foliage) directly reduce the area of photosynthetic tissue, whereas sap-sucking insects affect the ﬂow of ﬂuids and nutrients throughout the plant. Folivory is the best-studied aspect of herbivory. In fact, the term herbivory often is used even when folivory alone is measured, because loss of foliage is the most obvious and most easily measured aspect of herbivory. Folivory represents the direct consumption of photosynthetically active material. Consequently, the loss of leaf area by a tree can be used as a relative term to indicate the severity of an herbivore outbreak. In contrast, other herbivores such as sap-suckers or rootborers cause less conspicuous damage to trees which is more difﬁcult to measure, and also may have longerterm impacts (e.g., by disease transmission). Patterns of herbivory All plant species support characteristic associated insect herbivores, although some host a greater diversity of herbivores and suffer higher levels of herbivory than do others. Some plants can sustain high levels of herbivory and survive, whereas other species suffer mortality at signiﬁcantly lower levels. The consequences of herbivory vary signiﬁcantly, not just among plant species, but also in conjunction with different spatial and temporal factors (Huntly, 1991). For example, the combination of drought and herbivory or the effects of forest fragmentation on herbivore activity can signiﬁcantly affect the ability of the host plant to respond. The relative timing of herbivory and the intervals between attacks also have important effects on the overall impact on the forest ecosystem. Measurement of herbivory is usually expressed in temporal units (e.g., daily or annual rates), and ranges from negligible to several times the standing-crop biomass of foliage (Table 9.1), depending on forest type, environmental condition, and regrowth capability of the plant (Lowman, 1995). Lowman (1992)

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Table 9.1 Herbivory measured in temperate and tropical forests (including understory) 1 Location

Technique 2

Forest type

Level of grazing

Source

tropical rainforest

7.5% (new leaves)

1

Stanton (1975)

tropical rainforest

30% (old leaves)

1

Stanton (1975)

Panama

tropical rainforest

13%

1

Wint (1983)

Barro Colorado Island (Panama)

tropical rainforest

8% (6% insect; 1–2% vertebrates)

2,5

Leigh and Smythe (1978)

15%

2,5

Leigh and Windsor (1982)

Tropical Costa Rica

Puerto Rico

understory only

21% (but up to 190%)

3

Coley (1983)

tropical rainforest

7.8%

2

Odum and Ruiz Reyes (1970)

5.5–16.1%

4

Benedict (1976)

2–6%

4

Schowalter (1994)

2–13%

4

Schowalter and Ganio (1999)

Venezuela

understory only

0.1–2.2%

4

Golley (1977)

New Guinea

tropical rainforest

9–12%

1

Wint (1983)

Australia

sclerophyll

15–300%

4

Lowman and Heatwole (1992)

subtropical rainforest

14.6%

4

Lowman (1984)

tropical rainforest

8–12%

4

Lowman et al. (1993)

temperate deciduous

7–10%

1

Bray (1964)

2–10%

4

Reichle et al. (1973)

1–5%

4

Schowalter et al. (1981)

Cameroon (Afrika) Temperate North America

coniferous Australia

Europe

<1%

7

Schowalter (1989)

1–6%

7

Schowalter (1995)

sclerophyll

15–300%

4

Lowman and Heatwole (1992)

dry sclerophyll

11–60%

1

Fox and Morrow (1986)

3–6%

6

Ohmart et al. (1983)

montane or cloud rainforest

26%

4

Lowman (1984)

warm temperate rainforest

22%

4

Lowman (1984)

deciduous

7–10%

1

Nielsen (1978)

1

Updated from Lowman (1995) by permission from Academic Press. Techniques: 1, visual ranking; 2, litter trap; 3, graph paper or template squares; 4, leaf area meter; 5, other calculations; 6, insect frass; 7, estimation of missing or truncated needles.

2

compared herbivory on ﬁve tree species in Australian rain forests, and found highly variable spatial and temporal patterns. For example, Dendrocnide excelsa, typical of early-successional species, had relatively soft, short-lived leaves devoid of tannins or phenolics (Lowman and Box, 1983) and averaged 42% loss of leaf area, due primarily to a host-speciﬁc chrysomelid beetle (Hoplostines viridipennis; Fig. 9.1). In con-

trast, Doryphora sassafras, typical of late-successional species, had relatively tough, long-lived leaves with high concentrations of tannins and phenolics, and averaged only 15% loss of leaf area. Doryphora sassafras had signiﬁcantly less insect damage to its sun leaves in the upper crown (13%) than to its shade leaves in the lower crown (16%; Fig. 9.1), and suffered signiﬁcantly more herbivory in warm-

256

T.D. SCHOWALTER and M.D. LOWMAN

Fig. 9.1. Herbivory averaged over a period of 3 years for neighboring trees of Dendrocnide excelsa and Doryphora sassafras (open circles, sun; solid circles, shade) in Australian rain forest. Dendrocnide excelsa is a colonizer that grew in a disturbed patch, and Doryphora sassafras is a slow-growing, shade-tolerant species in undisturbed conditions. The standard error common to all the observations in the graph is indicated in the lower right corner (site: complex notophyll vine forest in Dorrigo National Park, New South Wales).

temperate rainforest stands (23%) than subtropical rainforest stands (15%; Lowman, 1992). Levels of herbivory are variable between forest stands and among ecosystem types (Table 9.1), although researchers have used a variety of methods, which makes interpretation and comparisons difﬁ-

cult [reviewed by Lowman (1995) and Lowman and Wittman (1996)]. Eucalypt forests have shown annual foliage losses of more than 300% of the foliage standing crop (Lowman and Heatwole, 1992). Although some studies suggest that tropical forests sustain higher levels of herbivory than do temperate forests (Coley

FOREST HERBIVORY: INSECTS

and Aide, 1991), the data in Table 9.1 indicate considerable variability in herbivory among forest types, even when comparable methods are used. Most methods are short-term snapshots of herbivory and do not provide information on deviations in environmental conditions, plant chemistry, or herbivore densities from long-term means. Long-term studies using standardized techniques are necessary to compare rates of herbivory among forest types. Temporal and spatial variability in herbivory has recently become a subject of debate in terms of how best to quantify this process, as biologists are becoming more capable of sampling leaves in the canopy. The methodological problems in sampling herbivory are twofold: 1) logistic problems of access to leaves in tall trees; and 2) selection of a method of measurement that is both accurate and efﬁcient in use of time. The levels of accuracy among methods of measurement must take into account the speciﬁc hypotheses tested. For example, measuring the percentage of leaf area missing at a point in time may be an appropriate measure of the effect of herbivory on photosynthetic capacity or canopy–atmosphere interactions, but is an inaccurate representation of consumption by herbivores because holes expand as leaves expand, and completely consumed or prematurely abscissed foliage cannot be assessed (Lowman, 1984; Risley and Crossley, 1993). Changes in herbivore abundance affect patterns and rates of herbivory. However, few studies have addressed changes in herbivore abundance or herbivory as a result of disturbance or succession. Schowalter et al. (1981) (see also Schowalter, 1994, 1995) used the same methodologies to compare canopy herbivore abundances and folivory in replicate disturbed (by harvest or hurricane) and undisturbed patches of temperate deciduous, temperate coniferous, and tropical evergreen forests. In all three forest types, disturbance resulted in a shift in dominance from folivores to sap-suckers. However, individual species within each of these functional groups showed different responses to disturbance; some folivores became more abundant and some sap-suckers became less abundant after disturbance, even on the same plant species. These different responses indicate that species belonging to a particular functional group are not redundant but maintain distinct functional roles under different environmental conditions. Continued measurements in these forests (Schowalter and Ganio, 1999) indicate that dominance shifts back to folivores within ﬁve years in the tropical forest.

257 INSECT HERBIVORE RESPONSES TO ENVIRONMENTAL CHANGE

Although outbreaks of herbivorous insects may resemble other disturbances, in terms of their effect on ecosystem structure and function, they differ from other disturbances in being a biotic response to changing environmental conditions. Herbivore populations are sensitive to changes in environmental factors, such as abiotic conditions, host plant abundance and condition, and predation, as these inﬂuence herbivore growth, survival, and reproduction. Effects of these abiotic and biotic factors on herbivorous insects are summarized below. Abiotic effects Disturbances affect many herbivore populations directly, by creating lethal conditions. Populations of many species can be greatly reduced by severe disturbances, and rare species may be eliminated (Torres, 1992; Schowalter, 1994, 1995). For example, Willig and Camilo (1991) reported the virtual disappearance of two species of walkingsticks (Agamemnon iphimedeia and Lamponius portoricensis) from forests of tabonuco (Dacryodes excelsa) in Puerto Rico following Hurricane Hugo in 1989. Miller and Wagner (1984) reported that pandora moth (Coloradia pandora) in forests of ponderosa pine (Pinus ponderosa) in western North America pupates preferentially in bare soil where it is more likely to survive ﬁre, a frequent disturbance in these forests. Similarly, sensitive species may be decimated by toxic effects of atmospheric pollutants, such as ozone or acid precipitation (Alstad et al., 1982). In contrast, Torres (1988) reviewed cases of insect herbivores being transported into new areas by hurricane winds, such as African desert locusts (Schistocerca gregaria) deposited on Caribbean islands. Disturbances also have indirect effects on forest herbivores through alteration of light, temperature, and moisture. Altered light penetration through the canopy can affect visual orientation of some insect species toward necessary resources. For example, aphids orient toward green and yellow wavelengths, which are characteristic of young succulent leaves (Matthews and Matthews, 1978), young leaves are more visible and available in canopy zones exposed to light. Light intensity may also affect abundances of predators that regulate herbivore abundances. For example, Oboyski

258

(1995) found that the guild of webspinning spiders (trapping by non-visual means) was signiﬁcantly more abundant in darker canopy zones whereas the guild of jumping spiders (hunting visually) was signiﬁcantly more abundant in sunlit canopy zones. Vertical proﬁles of temperature and moisture typically mirror each other in closed-canopy forests; highest temperatures and lowest relative humidities generally occur in the upper canopy, whereas lower temperatures and higher relative humidities occur in lower canopy layers (Parker, 1995). In contrast, open canopies show more even proﬁles of temperature and relative humidity. Insect respiration, feeding, and developmental rates generally increase with temperature, within tolerance ranges that are species-speciﬁc. Insect species also vary in their tolerances to desiccation. Hence, the distribution of various species within the forest reﬂects gradients in temperature and relative humidity. Temperature and moisture also affect insect herbivores through the increased susceptibility of stressed plants to herbivores, disruption of mate attraction as pheromones are convected out of warm or open-canopied forests, and increased abundance and virulence of pathogenic viruses, bacteria, fungi, and nematodes in cooler, moister forests (Fares et al., 1980; Mattson and Haack, 1987). Biological effects Host condition Healthy plants have a variety of chemical, physical and phenological defenses against herbivorous insects (see Harborne, 1982; Coley et al., 1985; Aide, 1993). Feeny (1976) and Rhoades and Cates (1976) suggested that plant defenses should differ by successional stage: short-lived, rapidly growing early-successional species may be expected to produce highly toxic antiherbivore defenses to prevent damage, whereas longlived later-successional plants should primarily produce feeding deterrents to reduce herbivory. While this hypothesis may explain general patterns in temperate forests, Coley and Aide (1991) compiled data from a number of studies indicating that later-successional tropical species with longer foliar life-spans may have higher concentrations and toxicity of defensive compounds than do species of temperate forests or early-successional species in tropical forests These differences correspond to the higher rates of herbivory generally reported in tropical forests (although they acknowledged a number of biases in the source data).

T.D. SCHOWALTER and M.D. LOWMAN

Disturbances, or other environmental changes, promote herbivore population growth on stressed plants or poorly defended plant species which replace betterdefended species following disturbance. Disturbances injure surviving plants and alter environmental conditions, thus affecting uptake and allocation of water and nutrients by these and early-successional plants (Schowalter, 1985). Wound repair and replacement of lost foliage or root tissues to meet metabolic demands require redirection of carbohydrates and nitrogen from energetically expensive defensive compounds, such as phenols, terpenes, and alkaloids (e.g., Coley et al., 1985; Schowalter, 1985). If herbivorous insects have adapted to particular plants by developing mechanisms to detoxify or avoid their defenses, they feed selectively on these plants. Because these adaptations also place demands on energy resources, herbivorous insects face an evolutionary trade-off between speciﬁc adaptations that tailor (and restrict) them to develop and survive on a speciﬁc host plant and more general adaptations that allow a wider host range but sacriﬁce survival under conditions favorable to some host plants. As stressed plants reduce production of defensive compounds, they become more general biochemical resources suitable for generalist, as well as specialist, herbivores. Consequently, the likelihood of herbivore outbreaks increases following widespread disturbances, or chronic environmental change such as atmospheric pollution (Alstad et al., 1982). For example, outbreaks of bark beetles and rootfeeding beetles typically follow tree-injuring disturbances (Witcosky et al., 1986; Paine and Baker, 1993). Host abundance Disturbances change the abundance of plant species through selective mortality. Severe windthrow may destroy much of the overstory, leaving mainly understory vegetation. Fire, depending on its severity, may eliminate understory vegetation, may destroy both understory and overstory, or may target particularly intolerant species. Heavy rains can saturate certain types of tree trunks, leading to selective tree-fall of speciﬁc species in a forest stand (Lowman, unpubl. data). Plant species abundances subsequently change during recovery (succession), as various species respond to modiﬁed abiotic conditions and increasing competition. These changes affect vegetation diversity, as well as the proximity and apparency (likelihood of being perceived) of particular plant species to hostseeking herbivores (Schowalter et al., 1986). Herbivorous insects are highly sensitive to host

FOREST HERBIVORY: INSECTS

spacing and to confusion by the presence of nonhost plant species (Risch, 1980; Kareiva, 1983; Visser, 1986; Schowalter and Turchin, 1993). Different plant species produce different blends of volatile chemicals, which insect herbivores use as host cues (Visser, 1986). Volatiles from non-host plant species are non-attractive, or even repellent, to a given insect. Spacing affects the ability of insects to perceive host cues, to reach a potential host with limited time and energy reserves, and to avoid the attention of predators while searching for hosts (Kareiva, 1983; Schowalter, 1985). Hence, little search effort is needed to colonize closely-spaced hosts in monocultures, compared to sparse hosts or hosts mixed with non-hosts. Dense hosts also are more likely to be stressed (and susceptible to insects) as a result of competition for limited resources or of injury by falling neighbors. A particular plant species may become more apparent to its associated herbivores if disturbance selectively reduces the abundance of nonhosts (Visser, 1986). Succession following disturbance often produces temporarily dense monocultures of rapidly reproducing ruderal plant species, which generate herbivore outbreaks. For example, Torres (1992) reported outbreaks of several species of Lepidoptera on early-successional grasses, herbs, and vines that became abundant in the Luquillo Mountains of Puerto Rico following Hurricane Hugo. Landscape effects The growth of herbivore populations triggered by changes in host-plant condition or abundance may be restricted to disturbed patches in otherwise unsuitable landscapes, or may propagate into adjacent areas where conditions permit population spread (Schowalter and Turchin, 1993; Schowalter, 1995). Widespread disturbances may create even-aged, low-diversity forests, which promote the growth of herbivore populations over large areas of susceptible hosts. For example, the large-scale ﬁres characteristic of the southern United States and arid western North America historically resulted in the development of low-density pine forests over millions of square kilometers. Populations of bark beetles (Dendroctonus spp., Ips spp., Scolytus spp.) in these forests were restricted to scattered injured or diseased trees, but likely reached levels capable of killing some trees in isolated thickets (Goheen and Hansen, 1993; Schowalter and Turchin, 1993). Anthropogenic suppression of ﬁre and conversion of forests to more rapidly growing, commercially valuable species on a regional scale resulted in dense

259

monocultures of susceptible trees, which now support widespread outbreaks of bark beetles and associated pathogens (Goheen and Hansen, 1993; Hadley and Veblen, 1993). Abundant early-successional species in tropical forests may also increase herbivory, as found in the Cecropia schreberiana thickets that developed in the forests of Puerto Rico after Hurricane Hugo (Schowalter, 1994; and unpubl. data). Forest fragmentation may also affect herbivory in forests. Kruess and Tscharntke (1994) and Schowalter (1994, 1995) reported that diversity and abundance of predator species were greatly reduced in disturbed or fragmented habitats, releasing herbivore populations from control by predation. Roland (1993) reported increased duration of a defoliator in fragmented forests. Herbivore outbreaks across landscapes dominated by susceptible patches could threaten areas within these landscapes that are designated for preservation. Otherwise resistant patches of forest could be overwhelmed by constant inundation by herbivore populations from adjoining patches. For example, Futuyma and Wasserman (1980) reported severe defoliation of scattered white oaks (Quercus alba; an ordinarily resistant host) by fall cankerworm (Alsophila pometaria) in stands dominated by this insect’s primary host, scarlet oak (Quercus coccinea). In contrast, forest fragmentation can interrupt spread of some bark beetles, such as Dendroctonus frontalis (Schowalter, 1985). An obvious issue regarding preservation of particular patches within landscapes is the minimum critical size needed to protect ecosystem processes and prevent disruption by herbivores from neighboring patches (Lovejoy and Bierregaard, 1990). Few natural experiments have addressed this issue. Shure and Phillips (1991) suggested that the signiﬁcant effects of the size of a clear-cut area on arthropod diversity, abundance of functional groups, and biomass might reﬂect the extent of environmental difference between the cleared patch and surrounding forest. In no situation have insect pests been manipulated experimentally between fragments of different size. Such an ambitious project would require overwhelming logistic and ﬁnancial support. Nevertheless, controlled experimental studies have indicated that herbivore damage is more extensive in low-diversity stands than in stands with greater distance between conspeciﬁcs (e.g., Schowalter and Turchin, 1993), just as in agricultural monocultures versus polycultures (Risch, 1980; Kareiva, 1983). In natural forests, herbivory of Doryphora sassafras was

260

greater in stands where the species was more common than where it was relatively rare (Lowman, 1992). In reviews of forest-dieback situations (Old et al., 1981; Huettl and Mueller-Dombois, 1993), insect epidemics commonly were implicated as major causes of mortality, and were usually most severe, in situations where human disturbance was part of land-use history. In contrast, outbreaks are considered less frequent in natural forest ecosystems where human disturbance has not been a major factor (but see Wolda and Foster, 1978; Wong et al., 1990). The relationships between insect populations, forests, and human impacts are the subject of extensive global speculation, as awareness of the importance of forests in regulation of global climate increases (e.g., Salati, 1987).

INSECT HERBIVORES AS AGENTS OF DISTURBANCE

Although debate may continue over whether or not insect herbivores should be regarded as agents of disturbance (Veblen et al., 1994; see also D’Antonio et al., Chapter 17, this volume), rather than simply as ecosystem components that respond to environmental change (Schowalter, 1985; see also Willig and McGinley, Chapter 27, this volume), herbivore outbreaks can dramatically alter ecosystem structure and function in much the same way as other disturbances. Effects of herbivore outbreaks, as for other disturbances, depend on the type, intensity, severity, scale, frequency, and regularity of the outbreak, as described below. Different herbivore species and functional groups (e.g., foliage chewers versus sap-suckers, described above) determine the type of disturbance – that is, the plant species and part(s) affected. Outbreaks of any insect species affect some plant species more than others, depending on host preferences. Even the gypsy moth (Lymantria dispar), an exemplar of polyphagy, shows a distinct preference for tree species producing mainly phenolic defenses, and avoids species with non-phenolic defenses (Miller and Hansen, 1989). Monophagous insects can virtually eliminate a particular plant species from the outbreak area. For example, southern pine beetle (Dendroctonus frontalis) can kill all pine trees (Pinus spp.) over relatively large areas, resulting in replacement by hardwood species (e.g., Schowalter, 1985); hemlock woolly adelgid (Adelges tsugae) has eliminated eastern hemlock (Tsuga canadensis) from large areas of the

T.D. SCHOWALTER and M.D. LOWMAN

northeastern United States (McClure, 1991). Foliage chewers open the canopy and transfer solid canopy material to the forest ﬂoor, in much the same manner as a storm, whereas sap-suckers siphon and transfer liquid materials from the plant. Intensity and severity of herbivory depend on the population level of the herbivore and the susceptibility of hosts. Most herbivore species never reach densities capable of causing serious injury to hosts. However, as noted above, different plants and plant species respond differently to the same intensity of herbivory, resulting in differing severity of injury. Deciduous trees often are capable of repeated refoliation following chronic complete defoliation, as a result in part of adaptations to seasonal senescence and refoliation, whereas many evergreen species are killed by a single complete defoliation (Schowalter et al., 1986). Herbivory can affect forests over a wide range of spatial scales. In some cases, a single tree may be targeted, as in the case of bark-beetle attraction to individual lightning-struck trees (Paine and Baker, 1993). In other cases, dense patches of hosts within an otherwise resistant forest may be affected. In the most extreme cases, herbivory can affect thousands of square kilometers. Furniss and Carolin (1977) have given examples of defoliator outbreaks in western North America that have covered 3000–200 000 km2 . Veblen et al. (1994) reported that, since ~1633, 39% of a montane forest landscape in Colorado (U.S.A.) has been affected by outbreaks of spruce beetle (Dendroctonus ruﬁpennis), compared to 9% by snow avalanches and 59% by ﬁre. Outbreaks of many herbivores occur infrequently and irregularly. Other herbivores show regular cycles of outbreak and decline. For example, outbreaks of the Douglas-ﬁr tussock moth (Orgyia pseudotsugata) in western North America occur at intervals of approximately 9–10 years and generally last 2–3 years (Mason and Luck, 1978). Veblen et al. (1994), in a montane forest landscape in Colorado, reported an average return interval for spruce beetle outbreaks of 117 years, compared to 202 years for ﬁre. Such patterns may correspond to climate cycles or to forest recovery time (re-establishment of dense hosts or host resources) following disturbance. Fuel accumulation resulting from outbreaks of some bark beetles and defoliators at particular successional stages ensures regular ﬁre return intervals, thereby maintaining relatively evenaged forests dominated by host species (Schowalter, 1985). Increasing the reliability of return interval for

FOREST HERBIVORY: INSECTS

herbivory or other disturbance should increase the rate of directional selection for tolerance.

EFFECTS OF INSECT HERBIVORES ON ECOSYSTEM PROCESSES

Herbivory affects ecosystem processes through its inﬂuence on plant reproduction, growth, survival, and turnover. Processes affected include primary productivity, community dynamics, biogeochemical cycling, and canopy–atmosphere–soil interactions, as discussed below. Primary productivity Traditionally, herbivory has been viewed solely as a process that reduces primary productivity. As described above, herbivory can remove several times the standing crop of foliage, alter the growth form of trees, or kill all trees of selected species over large areas during severe outbreaks. However, several recent studies indicate more complex effects of herbivory. Low or moderate levels of herbivory can stimulate plant growth (e.g., Carroll and Hoffman, 1980; Lowman, 1982; Trumble et al., 1993), whereas severe defoliation usually results in mortality or decreased ﬁtness (Marquis, 1984). Trumble et al. (1993) reviewed literature demonstrating that compensatory growth (replacement of consumed tissues) following low to moderate levels of herbivory is a widespread type of response. Wickman (1980) and Alfaro and Shepard (1991) reported that short-term growth losses by defoliated conifers were followed by several years, or even decades, of growth rates that exceeded predefoliation rates. Such compensatory growth likely depends on suitable growing conditions. Lovett and Tobiessen (1993) reported that defoliation resulted in elevated photosynthetic rates of seedlings of red oak (Quercus rubra) grown under conditions both of low and of high nitrogen availability, but that highnitrogen seedlings were able to maintain the high photosynthetic rates for a longer period of time. Hence, resource-limited trees are more likely to succumb to herbivores. Trees stressed by multiple factors, including herbivory, may lack resources or ability to compensate, and often succumb to mortality agents such as bark beetles (Schowalter, 1985). High levels of herbivory can exceed a plant’s ability to compensate and lead to growth reduction and mortality.

261

Although herbivory may reduce productivity of targeted plants, other plant species may compensate as a result of reduced competition for resources. For example, demise of pines during outbreaks of Dendroctonus frontalis in the southern United States typically stimulates growth of understory hardwoods (Schowalter, 1985); Wickman (1980) reported increased growth of ponderosa pine (Pinus ponderosa) following defoliation of white ﬁr (Abies concolor) by the Douglas-ﬁr tussock moth (Orgyia pseudotsugata) in the northwestern United States. Community dynamics Selective herbivory among plant species provides space and other resources to non-targeted plant species, resulting in altered plant community composition (e.g., Schowalter et al., 1986; Davidson, 1993). Herbivores also affect other animals and micro-organisms through changes in food resources and in plant and soil conditions. Depending on various conditions, including the successional stage of the vegetation and the particular herbivores involved, succession may be advanced or reversed (Davidson, 1993). For example, in coniferous forests of interior western North America, pine forest represents an earlier successional stage and ﬁr forest a later successional stage. Where moisture is adequate and ﬁre return intervals are long (riparian corridors and high elevations), bark beetles (Dendroctonus ponderosae) advance succession by accelerating the gradual replacement of host pines by the more shadetolerant, ﬁre-intolerant ﬁrs. However, where moisture is inadequate and ﬁre return intervals are short (lower elevations), defoliators and bark beetles associated with the water-stressed ﬁrs reverse succession, and help to maintain the earlier-successional pine forest. Changes in plant community composition and structure affect habitat and food for other animals and micro-organisms. Changes in abundance of particular foliage, fruits, or seeds affect abundances of animals that use those resources. Populations of animals that require or prefer nesting cavities in dead trees may be increased by tree mortality resulting from herbivore outbreaks. Changes in the rate or quality of litter fall and in the litter microclimate resulting from folivory can affect litter communities. For example, Schowalter and Sabin (1991) reported that three taxa of litter arthropods were signiﬁcantly more abundant under Douglas-ﬁr (Pseudotsuga menziesii) saplings

262

after defoliation. Insect herbivores themselves constitute highly nutritious resources for insectivores. The concentrations of essential nutrients in caterpillars, especially, are several orders of magnitude greater than those in foliage tissues (e.g., Schowalter and Crossley, 1983). Abundance of insectivorous birds and mammals may increase in patches experiencing outbreaks of insect herbivores (Barbosa and Wagner, 1989). The accidental invasion of insect herbivores such as the gypsy moth in the northeastern region of the United States has had widespread biological and economic impacts on forest ecosystems. A native of Europe, the gypsy moth preferentially consumes beech (Fagus spp.), but large populations will consume foliage of other trees. In the absence of predators, gypsy moth populations have had repeated outbreaks, resulting in the absence of oaks (Quercus spp.) from urban regions, forest fragments, and even large hillsides of deciduous woodlands. Other examples of introduced insect pests that have signiﬁcantly altered the forest community include the elm bark beetles Hylurgopinus ruﬁpes and Scolytus multistriatus, carriers of Dutch elm disease caused by the fungus Ceratocystis ulmi (U.S. Department of Agriculture, 1979) and the hemlock woolly adelgid (see above, p. 260). Biogeochemical cycling Insect herbivores affect biogeochemical cycling in a number of ways. Altered vegetation composition, discussed above, changes patterns of acquisition and turnover of various nutrients by the vegetation. For example, insects (such as bark beetles) that affect the relative proportions of pines and hardwoods in the southern United States or of Pseudotsuga menziesii and Thuja plicata in the northwestern United States, indirectly affect calcium dynamics and soil pH, since soil under hardwoods or Thuja has more calcium and a higher pH than under pines or Pseudotsuga (Kiilsgaard et al., 1987). Herbivory can affect biogeochemical cycling directly by changing the seasonal timing and amount of nutrient ﬂows from plants to litter or soil. Kimmins (1972), Seastedt et al. (1983) and Schowalter et al. (1991) found that the leaching of nutrients was greatly increased following chewing by herbivores. Risley and Crossley (1993) reported that herbivory also causes premature abscission of damaged foliage. Whereas litter-fall in the absence of herbivory may be highly seasonal – for instance, concentrated at the onset of

T.D. SCHOWALTER and M.D. LOWMAN

cold or dry conditions – herbivory increases litter-fall and nutrient turnover during more productive seasons when nutrient capture, by soil micro-organisms as well as the trees, may be more efﬁcient. However, Swank et al. (1981), working in the southern Appalachians of the United States, reported that defoliation of hardwood forests resulted in increased nitrate export in streamﬂow. Herbivory also affects the form of nutrients in litter and the rate of litter mineralization (Lovett and Ruesink, 1995). Senescent foliage, foliage fragments lost via herbivory, and foliage passed through herbivore digestive systems differ in the amount and form of nitrogen and carbon compounds, and in the degree of microbial pre-conditioning of this organic material. Where defoliators are dominant among herbivores, as is normally the case in undisturbed forests, material ﬂow is dominated by solid fragments with long turnover times. Disturbance in both temperate and tropical forests, however, causes dominance among herbivores to shift to sap-suckers (Schowalter et al., 1981; Schowalter, 1994, 1995), with the consequence that the material ﬂow is dominated by a rain of soluble materials with short turnover times. Reduced metabolic demands by pruned or defoliated plants can reduce water and nutrient uptake (Webb, 1978) and contribute to plant survival during drought periods (Parks, 1993). Since herbivore outbreaks often are associated with drought (Schowalter et al., 1986), future studies should address the extent to which herbivores contribute to ecosystem stability under these conditions. Canopy–atmosphere–soil interactions Herbivory affects abiotic conditions and the likelihood of future disturbance, especially through canopy opening and fuel accumulation. Canopy opening as a result of herbivore activity has immediate effects on ecosystem conditions similar to those resulting from ﬁre or storm disturbances. Opened canopies are subject to increased penetration of light, precipitation, and wind. These factors affect erosion, soil moisture, and soil fertility. Evapotranspiration and interception of precipitation are reduced (Parker, 1983) and orographic precipitation may be affected in montane or tropical regions (Salati, 1987). Increased leaf fall and accumulation of woody litter increase the likelihood and severity of ﬁre, especially in forests where these materials decompose slowly or where lightning strikes

FOREST HERBIVORY: INSECTS

263

Fig. 9.2. Dieback of Eucalyptus nova-anglica in New South Wales, Australia, illustrating the intensity of insect outbreaks that occur as a result of ecosystem disturbance and result in widespread tree mortality throughout the rural landscape.

are frequent (Schowalter, 1985). In high-diversity forests, intermittent herbivory may result in patches of open and closed canopy. The patchiness created may affect the understory by altering light levels and also variable canopy heights will affect the ﬂow-through patterns of rainfall (see Herwitz et al., 1994). Similarly, Herwitz et al. (1994) found that variation in canopy heights and aspects led to signiﬁcant differences in moisture at a small scale.

CASE STUDY – EUCALYPT DIEBACK IN AUSTRALIA

The defoliation of Australian eucalypts (Fig. 9.2) is considered one of the most severe tree declines in the world (Heatwole and Lowman, 1986; MuellerDombois, 1990/91) and has been linked directly to ecosystem disturbance. This syndrome has been recorded for over a century (Norton, 1886), but only during the past three decades have the episodes of dieback become more frequent and severe. Millions of eucalyptus are estimated to have suffered mortality

from the dieback, although its causes are not well understood. The loss of trees on the Australian landscape is of great signiﬁcance to agriculture, to the water table, to tourism, and to the sustainability of Australian ecosystems. In Australia, dieback has been attributed to different factors according to geographic region (Old et al., 1981; Heatwole and Lowman, 1986). In Western Australia, millions of jarrah trees (Eucalyptus marginata) suffered primarily from a fungal pathogen that invaded the root systems of the trees. Phytophthora cinnamomi was introduced to Australia inadvertently in the 1950s, when its spores were transported on tractor tires from soils in Indonesia. It quickly invaded the soils of Western Australia, and speciﬁcally targeted jarrah trees. Because jarrah was an important economic species, identifying the cause of its dieback was a particular concern. Extensive rehabilitation is under way. In South Australia, the dieback syndrome does not have one major cause as in the case for Western Australia (Boardman, 1981). The increasing salinity of soils, as agricultural practices intensify, has been blamed for dieback in several regions; additions of non-

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native grasses together with exotic grazers (e.g., sheep, cattle) have altered the natural balance of the Australian grassland and dry sclerophyll ecosystems. Fluctuating rainfall and temperature on these agricultural areas serve to exacerbate the conditions that lead to tree decline. In New South Wales and southern Queensland, the dieback is even more complex and controversial (Carter et al., 1981; Lowman and Heatwole, 1993). Diebacks in these states are reputedly the most severe in Australia, and the causes are the most difﬁcult to determine (D. Mueller-Dombois, pers. commun.). All of the factors (salt, weather, agriculture) in the South Australian dieback are implicated in the New South Wales diebacks, and have led to more extreme ﬂuctuations in the populations of organisms, both native and non-native. Distributions of many native birds have been reduced with the decline of suitable trees for nesting (Ford, 1989). Many of these birds were important predators on insects. For some herbivorous insects existing in the soil as larvae conditions for survival have improved as a result of compaction of roots and soil by sheep and cattle; this has led to increased numbers of scarabs (Anoplognathes sp.) and other herbivorous beetles (Roberts et al., 1982). For many decades, farmers have reported intermittent epidemics of scarab beetles, sawﬂy larvae (Perga afﬁnis), and other defoliators attacking eucalypts. The enhanced survival of herbivorous beetles, as well as other herbivores, is a consequence of human alteration or disturbance of the natural landscape. Lowman and Heatwole (1992) evaluated the impact of the herbivore component in the New South Wales dieback by tagging leaves of eucalypts and the related genus Angophora, and measuring herbivory and mortality monthly over ﬁve years (1982–1986). The study involved sampling different heights on trees representing the three different types of stands in rural Australia: healthy trees in woodland stands; trees in pastures; and dying trees in pastures. Cumulative annual levels of herbivory ranged from 8% (Angophora ﬂoribunda in woodlands) to 88% (Eucalyptus blakelyi in isolated stands surrounded by disturbed pastures). Occasional outbreak situations resulted in some species being entirely defoliated repeatedly, and cumulative foliage losses of up to 300% (Eucalyptus novaanglica, also in isolated stands surrounded by disturbed pastures). The highest levels of defoliation occurred in isolated trees or stands of trees that were surrounded by pastures, whereas the canopies within healthy

T.D. SCHOWALTER and M.D. LOWMAN

woodlands suffered only moderate herbivory (Fig. 9.3). Average defoliation for dieback stands in pastures was 60%, as compared to 32% and 14% for healthy trees in pastures and in woodlands, respectively. The differences between healthy and dieback trees in pastures were linked to the history of disturbance of the surrounding pastureland. Changes in the pasture that accompanied intensive stocking of cattle and sheep included trampling of soil, consumption of seedlings by stock, clearing of trees, girdling of trees by cattle, aerial spraying of fertilizers (especially superphosphate), alterations of the water table, planting of non-native grasses for winter feed supplements, and plowing of pastures for crops. The changes in soils following these practices created conditions conducive to epidemics of various folivores. As a result of clearing, fewer trees remained standing as food sources. Subsequently, the beetles, which feed gregariously, defoliated the remaining trees. This scenario was repeated over several years, and in many cases has resulted in complete mortality of forest fragments. As in many complex land-use situations, it is difﬁcult to implicate insects as the major cause of Australian dieback. Which came ﬁrst? – the insect defoliation leading to tree decline, or the environmental stresses on trees leading to increased defoliation? In other regions, the depletion of soils has been a primary progenitor of tree diebacks (e.g., in Hawaii, Mueller-Dombois, 1990/91). The dietary quality of foliage also has been linked to the decline of tree health (Landsberg, 1990). In addition to the stress of defoliation, insects may have other deleterious effects on native vegetation. Scarab beetle larvae (and other soil organisms) feed on tree roots, and create below-ground stresses on eucalypt trees (Lowman et al., 1987). This aspect of insect damage to trees is difﬁcult to quantify because measurements require destructive sampling, and biomass estimates usually include only the plant parts present and cannot account for the portions consumed by insects or other herbivores. Lowman et al. (1987) examined insect damage, both above- and below-ground, to a healthy tree and a dieback tree (New England peppermint, Eucalyptus nova-anglica), growing in close proximity and similar in stature. Borers and termites occupied 19% of the branches of the dieback tree, but only 5% of the healthy tree. These are conservative estimates, because branches and roots entirely consumed by insects could not be detected. The healthy tree had 990 kg of woody biomass, versus 640 kg for the dieback tree.

FOREST HERBIVORY: INSECTS

Fig. 9.3. Schematic illustration of the spatial variability of eucalypt herbivory by beetles, both above- and below-ground components. Numbers represent the proportion (%) of foliage area missing per year as a result of insect consumption, or the proportion (%) of ﬁne-root biomass consumed. All species represent the genus Eucalyptus (except for one Angophora species), of the family Myrtaceae, found on the New England Tablelands, New South Wales, Australia.

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The numbers of beetle larvae in disturbed soils were higher than in undisturbed soils (Roberts et al., 1982), exacerbating stress to the remaining trees. Stem borers, fungal pathogens, and sap-sucking insects have not been measured in terms of their impact on the Australian dry forest, but could emerge as additional stresses on the trees. Forest diebacks are worldwide and have various origins and etiologies (Mueller-Dombois, 1986, 1987). The Australian dieback is particularly severe and directly related to land-use changes and insect outbreaks. Australians have an opportunity to make a positive impact on their environment by aiding in the regeneration of native trees. The Greening of Australia project (Nadolny, 1991) aims to plant 1×109 trees during this decade. The ameliorative aspects of dieback research and land-use management in Australia may set a worldwide precedent, illustrating how scientists and landowners can work together to implement sound land management practices.

CONCLUSIONS

Insect herbivore populations can respond rapidly and dramatically to environmental changes (especially those resulting from natural and anthropogenic disturbances) that stress or increase the abundance of particular plant species in both temperate and tropical forests. Landscapes dominated by susceptible hosts or patches contribute to large-scale population growth of herbivores. Population outbreaks typically are triggered by disturbance or post-disturbance successional processes, but propagate the effects of disturbance into surrounding forest patches, sometimes affecting ecosystem structure and processes over extensive areas. In terms of intensity, severity, scale, frequency, and regularity, herbivore outbreaks are comparable to abiotic disturbances such as ﬁres and storms. Few studies have addressed relationships between herbivory and disturbance, largely because of limitations on measurements of herbivory. Nevertheless, available data suggest that the type of herbivory (e.g., by folivores versus sap-suckers) depends on successional stage in both temperate and tropical forests. Factors triggering elevated levels and herbivore effects on ecosystem processes appear similar between tropical and temperate forests. Furthermore, the scale and severity of herbivory appear to be greater in forests with a history of human intervention than in more

T.D. SCHOWALTER and M.D. LOWMAN

natural forests. Herbivory may be considered as a keystone ecological process in an ecosystem, which can serve as an indicator of disturbance (Lowman, 1999). Low or moderate levels of foliage consumption with infrequent (but occasional) outbreaks typify a healthy forest ecosystem, whereas increased frequency of outbreaks indicates accelerated degradation. Tree condition and density are two major factors affecting herbivory. Both are inﬂuenced by disturbance and forest management practices. Forest diebacks have become a serious problem in all regions of the world and reﬂect a combination of changes in forest conditions, including those that promote insect outbreaks. Forest fragmentation also affects the dispersal of herbivore populations across the landscape and creates patches of plants in which herbivores and their predators may not be in balance. Herbivore effects on plant reproduction, growth and survival, and on the turnover of plants and plant parts, inﬂuence a variety of ecosystem processes, including primary productivity, community dynamics, biogeochemical cycling, and canopy–atmosphere–soil interactions. Herbivore effects on these processes are complex, ranging from the alleviation of plant stresses and facilitation of ecosystem recovery under some conditions to the demise of entire forests. A major challenge is the isolation and experimental testing of the impact of different levels of herbivory over large spatial scales – for instance at the ecosystem level. Given the apparent increase in herbivore outbreaks on vegetation, and the enormous biological and economic implications that such epidemics can have, such research is critical. Several major issues should be addressed in future studies. These include developing standardized methods for measuring longterm herbivory, directly assessing the relationships between disturbance and herbivory, and quantifying the impact of human intervention in ecosystems on herbivory. In particular, future studies must use sufﬁcient replication to compare long-term patterns of herbivory under conditions of experimental disturbance (especially of anthropogenic activities) and recovery.

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Chapter 10

DISTURBANCE IN MEDITERRANEAN-CLIMATE SHRUBLANDS AND WOODLANDS Philip W. RUNDEL

INTRODUCTION

Mediterranean-type ecosystems and their disturbance regimes Mediterranean-climate ecosystems (MTEs), with characteristic and unique climatic regimes of mild wet winters and warm and dry summers, occur in just ﬁve regions of the world (di Castri and Mooney, 1973; di Castri et al., 1981). These regions are California, central Chile, the Mediterranean Basin, the Cape Region of South Africa, and Southwestern and South Australia (Table 10.1). Although the combined area of these ﬁve regions is less than 5% of the land area of the earth, MTEs are home to nearly 50 000 species of vascular plants, 20% of the world’s total (Cowling et al., 1996). Nowhere outside of lowland tropical rainforests are there ecosystems with higher regional diversities of species. Characteristic mediterranean-type climates are characterized by cool, wet winters with typically 90% or more of the annual precipitation falling in the six winter months. Mean annual precipitation is as low as 250 mm in coastal areas of the MTEs and reaches to as high as 900 mm at the upper margins of the shrubland zone. Mediterranean-type climates in the broad sense also include adjacent arid regions which maintain winter rainfall regimes (e.g., the Mojave Desert in California, Atacama Desert in Chile and Succulent Karoo in South Africa) and greater precipitation levels in higher mountain forests. Although frosts may occur throughout much of the MTE regions, these are infrequent and relatively mild. Aschmann (1973a,b) has suggested an upper limit of 3% of the annual hours with temperature below freezing in these regions.

A characteristic feature of MTEs is the widespread dominance of evergreen shrublands dominated by species with sclerophyllous leaves. These shrublands are called chaparral in California (Rundel and Vankat, 1989), matorral in Chile (Rundel, 1981a), garrigue or maquis in the Mediterranean Basin (Polunin and Walters, 1985), fynbos in South Africa (Cowling, 1992), and kwongan or heathlands in southwestern Australia (Pate and Beard, 1984). The dense cover of these shrublands burns readily under dry summer conditions with low humidity, although with differing frequencies characterizing individual MTE regions as described below. Thus, morphological, ecophysiological, and phenological adaptations to post-ﬁre regeneration of these stands through resprouting and ﬁre-stimulated reseeding are characteristic. Despite the general characterization of MTE regions as having dominance by evergreen, sclerophyll shrublands (Rundel, 1988), other vegetation forms are also present. Woodlands are widespread in most MTEs, particularly in areas with deeper or richer soils, or as riparian woodlands or gallery forests in wetter sites. Oak woodlands dominated by species of Quercus are widespread in both California and the Mediterranean Basin, with both evergreen and deciduous species as dominants. These communities can take the form of closed canopy evergreen woodlands grading into shrublands as in live oak woodlands of Southern California and the maquis of Europe, or open savannas of deciduous oaks that are widespread in both regions. Central Chile once had widespread dry sclerophyll and wet sclerophyll woodlands (Rundel, 1981a), while evergreen eucalypt woodlands termed mallees are widespread in western Australia (Noble and Bradstock, 1989). Only the Cape Region of South Africa of all of

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Table 10.1 Plant species diversity and regional disturbance threats in the mediterranean-climate regions of the world 1 Region

Area (106 km2 )

Native plant species

Regional disturbance threat 2

California

0.32

4300

urbanization, agriculture

Central Chile

0.14

2400

deforestation, overgrazing, agriculture

Mediterranean Basin

2.30

25 000

deforestation, overgrazing, agriculture, urbanization

Cape Region, South Africa

0.09

8550

invasive alien plants, agriculture, urbanization

Southwestern Australia

0.31

8000

agriculture, deforestation, introduced pathogens

1 2

Adapted from Cowling et al. (1996). The order of regional disturbance threats roughly follows their signiﬁcance beginning with the most critical factors.

the MTEs is largely lacking in woodlands, with such communities largely restricted to scattered relict stands of low tree diversity along the southern coast. The evergreen, sclerophyll shrublands of California and Chile commonly grade off at their arid interior margin and along drier coastal margins to a vegetation dominated by drought-deciduous shrubs. This community, which may also dominate early successional disturbance sites or arid microsites in evergreen shrublands, is termed coastal sage scrub in California and coastal matorral in Chile (Rundel, 1981a,b; Esler et al., 1998). Structurally similar communities with mixed dominance of low evergreen and deciduous shrubs is called phrygana in Greece and the eastern Mediterranean Basin (Polunin and Walters, 1985). Deciduous shrubs are largely lacking, however, from similar habitats in the Cape Region of South Africa and southwestern Australia. Natural disturbance regimes While it might seem that natural disturbance regimes would be quite similar in the ﬁve MTEs, they differ in a number of distinct ways (Rundel, 1998). Although all ﬁve regions experience characteristic summer drought, the magnitude of this drought is particularly severe in California, Chile and much of the sub-arid portion of the Mediterranean Basin where 6–8 months or more may pass without measurable rainfall. Extreme drought such as this is rare in South Africa and southwestern Australia where summer months frequently have a few light showers (Rundel, 1995). The evergreen sclerophyllous shrublands which form

the major component of vegetation cover in MTEs are highly ﬂammable, and thus ﬁre frequencies and intensities are important components of disturbance regimes (Mooney and Conrad, 1977; Rundel, 1981b, 1983; Trabaud and Prodon, 1993; Moreno and Oechel, 1994). Natural ﬁres are signiﬁcant ecological events in consuming above-ground vegetation in four of the ﬁve MTEs. Natural ﬁre frequencies, however, appear to be quite different between regions. In South Africa, for example, fynbos vegetation in the Cape Region commonly burns at intervals of 10–15 years (Van Wilgen et al., 1992), while in California natural frequencies are thought to be 30–50 years or more (Rundel and Vankat, 1989). Chile is the exception among MTEs, with natural ﬁres as a rare event. The geographic isolation of central Chile, protected from summer storms with lightning by the high Andean Cordillera, has little evidence of ﬁre as an important ecological disturbance regime in the evolution of the native ﬂora. There are other strong environmental differences in addition to potential drought stress which make the Cape Region of South Africa and southwestern Australia distinct from the other three MTEs. These two areas lie in geologically ancient and stable landscapes, resulting in highly leached and nutrientpoor soils. In contrast, earthquakes, volcanic activity, orogenic uplift, and other dynamic processes create natural disturbance regimes in California, Chile, and the Mediterranean Basin that are absent in South Africa and Australia. Unlike South Africa and southwestern Australia, the younger landscapes of these three regions have experienced tremendous changes in climate regime and landscape structure

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Fig. 10.1. The origins of the introduced ﬂoras of California, Chile, the Cape Region of South Africa, and South Australia, expressed as a percentage of the total introduced ﬂora of each region. From Groves (1991).

in Quaternary and even Holocene times and these changes have had profound impacts on community structure and speciation (Cowling et al., 1996). The remarkable patterns of speciation in ﬁre-sensitive shrub lineages in the Cape Region of South Africa and southwestern Australia likely have resulted from a combination of relatively mild and stable Quaternary climatic conditions coupled with high ﬁre frequencies in these nutrient-poor habitats (Cowling et al., 1992).

Human history In addition to their differences in environmental disturbance regimes, the ﬁve MTEs of the world have experienced very different histories of human impact. The Mediterranean Basin was the site of early human cultural advances, and indigenous agriculture and animal husbandry have been practiced there for more than 10 000 years (Naveh and Dan, 1973; Le Hou´erou, 1981; Naveh, 1991). The other four MTEs

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were occupied by relatively small populations of hunter-gatherers until only a few centuries ago. Chile was ﬁrst colonized by the Spanish in 1550, and the Dutch settled in South Africa in 1652. California was not colonized by the Spanish until 1769, and there was only a small population of European ancestry until the discovery of gold in 1848. It was only in 1827 that the British colonized southwestern Australia. It is not surprising, with the cultural differences between colonizing nations, that the different levels of population growth and economic development, and the nature of the natural environment in each region have combined to produce a complex of similar and contrasting disturbance regimes in the ﬁve MTEs. Changes in ﬁre frequencies and intensities, grazing pressure, urbanization, agricultural expansion, deforestation, and the introduction of exotic species and pathogens are all aspects of regional disturbance to natural vegetation in one or more MTEs (Table 10.1). Exotic species The invasion of exotic species into MTEs, which generally involves species native to other MTEs, has had a signiﬁcant impact on the structure and function of many ecosystems within these regions. Disturbance has been a strong correlate of such invasions, but is certainly not the only factor involved. There are two features of MTEs that have been hypothesized to be signiﬁcant in promoting the invasion of alien plant species (Groves and di Castri, 1991). The ﬁrst of these is the nature of open habitats available for colonization with the ﬁrst autumn rains falling after a long summer drought. The majority of herbaceous invaders in MTEs effectively invade such habitats by responding rapidly with germination from soil seed pools once adequate moisture conditions occur and temperatures are suitable. While dormancy mechanisms are often present, these are only sufﬁcient to maintain dormancy from the time of seed maturation in late spring to the onset of rains in autumn. The second important component of the invasion of exotic annual species in MTEs is the evolutionary origin of many of these species under a long history of human disturbance in the Mediterranean Basin (Le Hou´erou, 1981). As human populations in this region slowly developed large population centers and agricultural economies over a period of thousands of years, there was a natural selection process for biological attributes in weedy species which could adapt to such

Philip W. RUNDEL

conditions. Annual cultivation for cropping, grazing, deforestation, and changing frequencies and intensities of ﬁre all introduced selective forces on the life history attributes of native grasses and forbs. While the composition of exotic plant species within each of the ﬁve MTEs shows regional patterns related to the history of colonization, the nature of human settlement patterns, and trade routes for agricultural and livestock products are a prevailing feature in the predominance of European species in these invasive ﬂoras (Fig. 10.1). This is most pronounced in central Chile where approximately 80% of the widespread exotic species are European annuals (Montenegro et al., 1991). This pattern applies even within the Mediterranean Basin itself, where more than half of the introduced ﬂora of the western Mediterranean comes either from northern Europe or from other areas of the Mediterranean Basin (Guillerm et al., 1990). An interesting feature of the exotic ﬂora of MTEs is the surprising dominance of just four families – Poaceae, Asteraceae, Brassicaceae, and Fabaceae. It remains an open question whether or not MTEs are inherently invasible in comparison to other major vegetation associations (Roy et al., 1991). There is clearly a widespread dominance of invasive plant species in many MTE communities such as oak woodlands in California and acacia savannas in central Chile, but other communities have relatively little invasion. While the open community structure of matorral in Chile provides a habitat for many herbaceous exotics, there is relatively little parallel invasion of chaparral communities in California in the absence of human disturbance. These observations would suggest that invasive mediterranean species may be genetically preadapted to disturbance regimes with agricultural clearance, ﬁre, and grazing. Woody invasive species such as acacias and pines in South Africa, however, appear to successfully establish themselves in undisturbed habitats of native vegetation (Richardson et al., 1990; Manders and Richardson, 1992; Musil, 1993).

LANDSCAPE DISTURBANCE REGIMES

The Mediterranean Basin Historical background Natural ecosystems of the Mediterranean Basin were, as described above, among the ﬁrst to be impacted by man as agricultural civilizations evolved

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Fig. 10.2. Changes in land use of Mediterranean landscapes from pre-World War II (above) to today (below). From Perez-Trejo (1994).

and spread in this region beginning 10 000 years or more ago. Thus, anthropogenic disturbances such as deforestation, grazing, agricultural development, and ﬁre management have been factors in inﬂuencing community structure and diversity for a very long period of time (Aschmann, 1973a,b; Naveh and Dan, 1973). Highly developed political systems with kingdoms or city-states and associated agricultural development were extensive in the eastern Mediterranean and Middle East by 2000 B.C. and rapidly spread westward to the Atlantic. Clear examples of extensive deforestation and erosion date back to Greek and Roman times. Many land-use practices established by the Romans,

however, were relatively sustainable, and these forms of agro-pastoral use of Mediterranean woodlands and shrublands continued without signiﬁcant modiﬁcation up until the middle of the 20th century (Caravello and Giacomin, 1993). The changing patterns of land-use and disturbance in the past half century (Fig. 10.2), however, have brought dramatic changes in the structure and diversity of these communities. Landscape disturbance in oak forests of the Mediterranean Basin Oak forests and woodlands cover about 10×106 hectares of the western Mediterranean Basin, with two major plant communities distinguished. Savanna-

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like landscapes of the Iberian Peninsula dominated by Quercus ilex and Q. suber are known as dehesas and montados, and have played a central role in agroforestry systems which combine livestock grazing with harvesting of cork and ﬁrewood. Coppiced woodlands of Quercus ilex and Q. pubescens in southern France and Italy have been harvested for centuries for ﬁrewood and charcoal (Vos and Stortelder, 1992). Dramatic landscape alterations have occurred in both of these ecosystems over the past four decades. Ecological deterioration has been particularly evident in the dehesas and montados due to a change in traditional management practices. A steady decline of agro-pastoral land use has occurred with increasing deforestation and clearing to promote intensive agriculture and the extensive development of plantations of Pinus and Eucalyptus. Pinto-Correia (1993) has described how these ecosystems were adapted to the sustainable utilization of the open evergreen woodlands of oaks, olives (Olea europaea) and chestnut trees (Castanea sativa) for cork, fruit, wood, and livestock production, and cereal cultivation. Managed extensively for many decades with minimum human impact, this landscape maintained a ﬁne-grained and heterogeneous community structure with relatively high biodiversity. Related problems have occurred with oak forests and maquis communities of southern France and Italy. Widespread landscape abandonment associated with increasing urbanization has led to uneven management policies, rapidly expanding frequencies and extents of ﬁre as a disturbance factor, and loss of both landscape and biodiversity (Naveh and Lieberman, 1993; Etienne et al., 1998; Naveh, 1998; Papanastasis and Kazaklis, 1998). Studies of maquis forests and shrublands in northern Israel and southern France have found that high ﬂoristic and structural diversity of Mediterranean Basin communities are associated with moderate grazing pressures which maintain open habitats with regular disturbance (Naveh and Whittaker, 1979; Naveh, 1982; Kaplan, 1992). Where this intermediate disturbance regime has been altered either through reduced disturbance, land abandonment or protection from ﬁre, or increased intensity of disturbance, both ﬂoristic and structural diversity have been reduced considerably. In many areas, species richness has dropped by 75%, from more than 100–120 species ha−1 to less than 30 species ha−1 . Equitability was also much lower in communities free from disturbance, and the dominance and concentration of the most aggressive and shade

Philip W. RUNDEL

tolerant trees and shrubs was high. Sites which had been moderately grazed and coppiced in the past were found to have a woody plant community with good structural diversity in different strata and species, and a rich ﬂoristic diversity contributed by many small annuals, geophytes, and herbaceous perennials. This increase in diversity appears to be due to smallscale heterogeneity created by variable soil depth and micro-relief, availability of moisture and nutrients, rock outcrops, tree cover, litter cover, shade, and the abundance of different plant species, as affected by grazing and trampling (Naveh, 1991). Animal species richness and relative abundance of birds, reptiles, rodents and isopods showed similar trends with moderate disturbance (Warburg et al., 1978; Farina, 1989, 1998). Open oak woodlands of Tabor oak (Quercus ithaburensis) in Israel, which can be considered as the eastern Mediterranean counterpart to the montados and dehesas, have their highest species richness and diversity under moderate grazing pressures and lowest under both heavy and light grazing pressure or complete protection (Naveh and Whittaker, 1979). Desertiﬁcation in semi-arid areas of the Mediterranean Basin The problems of landscape disturbance and associated desertiﬁcation in the semi-arid Mediterranean Basin are major ecological and political issues which can be associated with three sets of linked processes – physical, biological and socio-economic (Fig. 10.3; Fantechi and Margaris, 1986; Perez-Trejo, 1994). Physical processes that promote desertiﬁcation include a degradation in soil structure, reduced inﬁltration capacity, and salinization, all of which work together in concert to promote secondary effects of increased severity of erosion and changes in hydrologic processes related to ground water recharge and surface ﬂow. These hydrologic factors, in turn, may work to promote salinization. Biological disturbance processes are those which relate to decreasing plant cover, and the obvious feedbacks that this has on animal population structure and physical processes as greater amounts of soil are exposed. High disturbance regimes in the semiarid Mediterranean Basin are generally associated with signiﬁcant reductions in the diversity of vegetation structure and plant species. The stability of Mediterranean Basin ecosystems is strongly tied to the nature of both grazing and ﬁre regimes. As plant communities have developed

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Fig. 10.3. Multiple factors involved in desertiﬁcation in the Mediterranean Basin. From Perez-Trejo (1994).

with moderate grazing pressure as a natural selective regime, plant species have evolved reproductive strategies linked to this disturbance. There has also been adaptation over evolutionary time to ﬁre as a natural environmental factor, and reproduction in many species is strongly related to ﬁre frequency and intensity (see Trabaud and Prodon, 1993). The frequency and magnitude of ﬁres in the Mediterranean Basin has doubled in the past two decades, with a consequence of decreased forest vigor and sharp alterations in forest structure and soil stability (Kuzucuoglu, 1989; Naveh, 1990). Socio-economic events which strongly affect processes of desertiﬁcation can be seen in the dramatic intensiﬁcation of land use in the Mediterranean Basin over the past few decades. Mechanization of agriculture, urbanization, and a rapidly developing tourist industry, which have impacted all of the Mediterranean Basin, have caused rapid changes in landscape utilization in more arid regions. These changing patterns of land use, and associated effects on fragmentation of natural plant communities, rates of soil erosion, ﬁre frequencies and intensities, and water resource availability have all promoted desertiﬁcation with dramatic reduction in natural biodiversity (Meeus

et al., 1990; Naveh and Lieberman, 1993; Perez-Trejo, 1994). California Historical background Indigenous populations appeared to have become widespread in California only about 6000–10 000 years ago, and native populations were never large. Nevertheless, these populations are known to have had a signiﬁcant impact on natural vegetation. As hunter– gatherers, indigenous tribes harvested many native species for food, ﬁber, medicine, or building materials, and made wide use of ﬁre to either drive game or clear areas to promote grazing by large animals (Anderson, 1993). Although the Spanish ﬁrst explored California as early as 1542, there were no permanent colonies until the construction of Dominican missions beginning in 1769. Only small numbers of Europeans came to California in these early days, with their estimated population only about 15 000 at the time that gold was discovered in 1848. Large herds of cattle and sheep were present at this time, however, and extensive areas of California grasslands were already being converted

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from dominance by native perennials to exotic annuals from the Mediterranean Basin. Following statehood in 1850, immigration to California occurred at a rapid rate with the subsequent establishment and growth of major cities. By the 1880s, there was already major urbanization of extensive areas of the California coast, and a rapid expansion of agricultural lands in the San Joaquin Valley and Southern California, although grazing pressure began to be reduced as sheepherding became less economically viable. Expanding urbanization and clearance for new agricultural lands has continued to the present day. Disturbance and changes in community structure Natural ecosystems in California have experienced differential levels of disturbance over the past two centuries, with some ecosystems remaining relatively pristine while the structure and diversity of others have been strongly impacted by human activities. The most dramatic example of altered ecosystem structure in California following disturbance is in the grasslands and oak woodlands of the state. These communities have become totally altered from a system largely dominated by native perennial grasses to one in which introduced European annual grasses and forbs provide almost all of the herbaceous plant cover (Heady, 1977). The sources of these invaders are reasonably well documented, with both deliberate and accidental introductions involved (Mooney et al., 1986; Rejm´anek et al., 1991). Flexible life-history characteristics with high reproductive effort, low root to shoot ratios, and large seed banks have pre-adapted these species to habitats with strong grazing pressure and frequent human disturbance (Jackson, 1985). It is interesting to note that edaphically-challenging soil substrates in California such as serpentine-derived or heavy clay soils retain a dominance of native annuals and herbaceous perennials, indicating that the widespread introduced annuals are not competitive in these habitats. Species diversity of such habitats are commonly much higher than that of grasslands dominated by introduced species. A major aspect of changing community structure in California chaparral and woodlands under human inﬂuence has been changes in the frequency and intensity of ﬁres. There is no question that indigenous populations in California have utilized ﬁres in managing plant succession, regrowth, and animal population patterns (Blackburn and Anderson, 1993). The arrival

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and establishment of large American populations in California in the second half of the 19th century led to ﬁre suppression policies that actively fought ﬁres. The result of this policy, which has only begun to be altered in recent decades, is the reduced frequency but increased intensity of chaparral ﬁres (Minnich, 1983; Rundel and Vankat, 1989). Coastal dune and prairie communities in the mediterranean-climate region of California are a second group of habitats where disturbance and the introduction of exotic species have dramatically changed ecosystem structure and diversity. The most signiﬁcant change has resulted from the widespread introduction of Ammophila arenaria from Europe for dune control. This aggressive species changes the natural structure of the dunes through impacts on dune stability and by the shading out of native species. Carpobrotus edulis from South Africa, originally planted for dune cover and erosion control along the coast, has also become a widespread invasive crowding out native species (D’Antonio et al., 1993). Until recent decades, only a small number of invasives from the Mediterranean Basin of Europe had become well established in California desert communities with winter rainfall regimes. Erodium cicutarium, widespread for many years, is most abundant in disturbed desert sites. In recent decades, however, Bromus rubens, has dramatically increased its range and dominance in the Mojave Desert (Rundel and Gibson, 1996). Population densities of this species were small in the 1960s, but began to rise signiﬁcantly in the 1970s. By the late 1980s, B. rubens had become the overwhelming dominant among annual species in the eastern Mojave Desert. Invasions of B. rubens and other mediterranean annual grasses have also become widespread in the Coachella Valley of the western Sonoran Desert in California, particularly in wet years. Extensive dry deposition of nitrogen into this ecosystem from the Los Angeles airshed may be an important factor in this invasion. As annual grass cover expands, dead stalks become available to carry ﬁre between desert shrubs in an ecosystem in which ﬁre was rarely present under natural conditions. The introduction of such ﬁres has the potential to profoundly alter the structure and diversity of this ecosystem. Distinctive assemblages of exotic species and a few weedy natives characterize highly disturbed habitats in California along roads and trails in vacant parcels of

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Central Chile

present in the coastal range and foothills of the Andes, but the structure and diversity of these ecosystems have also been negatively impacted by human actions (Rundel, 1981a). Patterns of anthropogenic changes in community structure and succession in both matorral and acacia woodland communities of central Chile have been described in detail by Balduzzi et al. (1982).

Historical background Indigenous populations of hunter–gatherers lived in central Chile for about 10 000 years before the Spanish colonization of central Chile began with the founding of Santiago in 1542 by Pedro de Valdivia. Agricultural and animal production rapidly became important economic activities following this colonization. Copper and silver deposits discovered early in the 19th century, and later development of nitrate mines, led to an economic development that allowed increasing levels of urbanization, particularly along the coast. With this development came a heavy impact on native plants. The sclerophyll woodlands in central Chile were rapidly felled for construction material, and matorral vegetation was cleared and burned for charcoal production. Agricultural expansion and the cultural use of goats by rural populations combined with this vegetation clearance to remove woody cover over large areas of central Chile (Fuentes and Hajek, 1979; Balduzzi et al., 1982). This impact, which has continued up until recent decades, is particularly apparent in the semi-arid transition region where overgrazing has led to extensive devegetation and desertiﬁcation over broad areas (Bahre, 1979; Etienne et al., 1986). The pre-Columbian vegetation of central Chile (30º– 37ºS) is thought to have been a dense and diverse woodland much like that occurring today in MTEs of California and the Mediterranean Basin with a dominance of evergreen, sclerophyllous trees and shrubs (Balduzzi et al., 1982). Much of this woodland was dominated by Maytenus boaria, Quillaja saponaria, Cryptocarya alba and Peumus boldus. Today, however, after more than four centuries of intensive human impacts, the vegetation of the central valley of Chile is largely covered by an acacia woodland (espinal) dominated by a single species, Acacia caven (Ovalle et al., 1990, 1996; Aronson et al., 1993). These open savannas support an herbaceous cover that is largely exotic annual grasses and forbs from the Mediterranean Basin, much as occur in oak woodlands of California today (Gulmon, 1977). Matorral communities are widely

Landscape disturbance regimes in Central Chile In comparison to California chaparral and Mediterranean garrigue and maquis communities where woody vegetation cover characteristically exhibits a continuous canopy, Chilean matorral is commonly much more open with clumped shrub distributions on all but wet sites. It has been suggested than this open structure is maintained in part by the impact of European rabbits which were introduced to Chile approximately 50 years ago and are now ubiquitous (Fuentes et al., 1983, 1986). The recovery of matorral vegetation following disturbance is very slow. Rabbit grazing, limited seed dispersal, and summer drought stress are all important ecological factors limiting the restoration of natural vegetation cover. A schematic model of matorral succession was presented by Fuentes et al. (1986) to show the interaction of these factors (Fig. 10.4). The rapid growth of the Chilean economy in the past decade and the centralization of this economy in Santiago has led to sharp changes in disturbance regimes over this period. As more and more of the rural population moves to the city, there has been extensive landscape abandonment and reduction in disturbance from overgrazing, fuel wood, and charcoal production. The consequent revegetation of these areas, however, has left them open to anthropogenic ﬁres whose frequency and average size have increased rapidly in recent years as continuous vegetation cover returns over areas formerly too open to carry a ﬁre. Although it has been suggested that vulcanism in central Chile has provided a community exposure to ﬁre through the evolution of species in the matorral (Fuentes and Espinosa, 1986), the absence of specialized modes of post-ﬁre recovery or re-establishment in the ﬂora suggests otherwise. Central Chile lacks any woody species with obligate post-ﬁre reseeding strategies, and there is likewise an absence of annual ﬁre-following species. Nevertheless, matorral shrubs do resprout rapidly after ﬁre, although the dynamics of post-ﬁre regrowth among such species has not been carefully studied.

land that have been cleared. The systematic composition of roadside vegetation in California has been discussed in detail by Frenkel (1970), and Vessel and Wong (1987) have described the natural history of vacant lots in California.

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Fig. 10.4. Patterns of vegetation change with disturbance in matorral communities of central Chile. Ovals in this drawing refer to processes and lines with arrows are transitions. Undisturbed matorral vegetation is shown by the double-squared box at the center. From Fuentes et al. (1986).

A second factor of major change in land-use pattern in central Chile has been the rapid extension of landscape clearance for the establishment of plantations of Pinus radiata and Eucalyptus globulus. More than 65 000 ha of plantations were planted annually during the 1980s, and that rate has nearly doubled in this decade (Aronson et al., 1998). Much of the area planted lies in the foothills of the coastal ranges in what had previously been some of the least disturbed natural ecosystems. Diverse communities of hygrophilous and sclerophyll forest, in particular, are being replaced by such monocultures. The Cape Region of South Africa Human history Human settlement in the Cape Region has a long history, extending back into Pleistocene and early Holocene periods. This region was likely peopled

as much as 150 000 years ago, and there is clear evidence for active hunter–gatherers of the Later Stone Age in this region 21 000 years ago (Deacon, 1992). Herding of sheep and cattle by Bushmen tribes has been present for at least 2000 years, suggesting that these tribes had signiﬁcant impacts on the fynbos ecosystems through changes in ﬁre frequency and accelerated rates of erosion. European colonization of the Cape region began in 1652 when the Dutch established a supply station at the Cape. Although the development and growth of a European population occurred slowly, and much of this region had soils too poor to support typical cereal and vegetable crops, impacts of human disturbance were rapidly evident. These impacts occurred ﬁrst with the reduction in large animal biomass and diversity with hunting, and the cutting of the relict Afromontane forest patches along the south coast (Deacon, 1992).

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Invasive species A striking feature of fynbos landscapes is the presence and frequent dominance of alien trees and shrubs in the landscape. The rapid spread and establishment of dominance by such species in relatively undisturbed natural communities is unprecedented in any other MTE (Richardson et al., 1992). Most signiﬁcant among these invasions has been the establishment of Australian acacias along watercourses, and Hakea sericea and pine species (P. pinaster and P. radiata) into mountain fynbos. The impact of these woody invasives has clearly been made not only in the composition of dominant species, but more signiﬁcantly at the ecosystem process level as well. Changes in fundamental ecosystem processes have occurred with these invasions, including the acceleration of riverbank erosion, reductions in streamﬂow through increased evapotranspiration, changes in nitrogen cycling dynamics, and altered ﬁre frequencies (Richardson et al., 1992; Stock et al., 1995; Le Maitre et al., 1996). While it has been suggested that the fynbos ecosystem is more invasible to alien animals than adjacent biomes in southern Africa, doubts have been raised about this hypothesis (Richardson et al., 1992). Invasions of alien birds have been well studied and show correlations with the degree of disturbance to the habitat (Glyphis et al., 1981; Armstrong and Van Hensbergen, 1994). Concerns with alien animal invasions to South African fynbos also involve invertebrate species whose presence has the potential to profoundly change community structure. Although invasions of Argentine ants (Iridomyrmex humilis) have occurred slowly, this species has the potential to displace native ant species which are important in the seed dispersal of many vascular plants (Bond and Slingsby, 1984; Giliomee, 1986). Southwestern Australia Agricultural expansion The transformation of the MTEs of southwestern Australia from a relatively pristine condition to their present situation of major landscape transformation has taken place in the little more than 150 years since European inﬂuence entered this region. Agricultural development of wheatlands, with associated fertilization and sheep grazing has had the greatest inﬂuence in establishing the fragmentation of kwongan and mallee habitats into a matrix of agricultural lands. Small remaining fragments of kwongan and mallee vegetation

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within the wheatbelt are generally too small to maintain viable populations of plants and animals, leading to rapid declines in native animal diversity and slower reductions in plant species diversity (Fox and Fox, 1986). The wheatbelt region was extensively cleared of native vegetation during the present century, with a rapid rise in the rate of clearing after World War II (Hobbs, 1998). Over the wheatbelt as a whole, only about 7% of original native vegetation remains, although this percentage varies greatly within the region, with the oldest settled areas retaining less vegetation. Before modern settlement, this region consisted of a mosaic of vegetation types, including various woodlands, shrublands, and mallee types. These communities were cleared to differing extents, depending on the suitability of the underlying soils for agriculture. Most impacted among plant communities were woodlands, with some types being all but eliminated from the region (Beard and Sprenger, 1984). Deforestation of native eucalypt forests in the wetter areas of southwestern Australia has led to signiﬁcant changes in forest structure and diversity. Woody understory legumes, an important component of nitrogen inputs to these low nutrient soils, have been reduced in numbers. These cut-over forests are commonly replanted to either native or exotic species, but with mixed success (Hobbs, 1992). Reforestation problems are frequently associated with both rising water tables and introduced pathogens, as described below. Deforestation as well as widespread strip mining of bauxite and other minerals in the heathlands and jarrah (Eucalyptus marginata) forests of southwestern Australia have led to slow ecosystem recovery and re-establishment of native species (Fox and Fox, 1986). One signiﬁcant aspect of the replacement of deep-rooted native trees with secondary vegetation of shallow-rooted herbaceous or semi-woody species has been the enhanced upward movement of saline paleosoil proﬁles, and the resultant salinization of surface waters and agricultural land (Peck, 1978). Nearly 3% of the agricultural lands of Western Australia are affected by salinization today, but this level is predicted to rise as high as 13% of cleared land in the next few decades (Nulsen, 1992). As this progression from landscape clearance to agricultural development to salinization continues to occur, it seems clear that current levels of agricultural production in this important region are not sustainable. Widespread revegetation work has been

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carried out in Western Australia to alleviate many of the problems associated with rising water tables (i.e. salinization, erosion, and waterlogging), but much of this has involved non-native species (McFarlane et al., 1992). Impacts of exotic species One important impact associated with disturbance, from deforestation and mining, has been the establishment of an exotic fungal pathogen, Phytophthora cinnamomi. This pathogen has had dramatic effects on the ﬂoristic structure of woody species in the areas where it has become established (Dell and Malajczuk, 1989). Die-back of native species infected by Phytophthora is particularly signiﬁcant in the Proteaceae, a family which naturally forms a major element of kwongan and woodland cover in southwestern Australia. Areas infected by Phytophthora characteristically have a reduction in proteaceous cover and an increase in herb cover, particularly by Cyperaceae (Hobbs, 1992). Deliberate introductions of mammals, plants, and ﬁsh to Western Australia result mainly from the use of these organisms for agriculture, horticulture, or sport without due consideration for the potential threats of invasiveness. Recent indications suggest that new species of plants continue to be introduced (Keighery, 1995), and there are recent examples of introduced species rapidly extending their range (Hobbs, 1998). Exotic species introductions of animals have also had dramatic effects on the community structure of shrublands and woodlands in southwestern Australia. The most signiﬁcant of these introductions was that of the European rabbit (Oryctolagus cuniculus) which had become established in widespread feral populations by 1859 (Fox and Fox, 1986). Although biological controls sharply reduced these populations in the 1950s, rabbits nevertheless had a dramatic effect on plant community structure through a century of their selective grazing and browsing. Such grazing opened up many areas to invasion by exotic plant species. While rabbits have had a major historical impact on plant populations, exotic carnivores such as foxes, cats, and ferrets have predated heavily on native mammals, ground- and hollow-nesting birds, lizards and insects (Hobbs et al., 1993). While habitat fragmentation has traditionally been considered to be the primary cause of declines in Australian faunal populations, the commencement of rapid fragmentation coincided with the arrival of rabbits and foxes. It is likely that these invading species, and the measures used to control them

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by the human population, were responsible for much of the faunal (especially mammal) decline (Hobbs et al., 1993). CONCLUSIONS

Mediteranean-type ecosystems provide many opportunities for comparative studies of the ecology of disturbance regimes, both natural and anthropogenic. These ﬁve regions share common characteristics of climatic regime, with an independent evolution of plant and animal species adapted to these conditions within each region. Thus, they provide a natural ecosystem experiment with ﬁve independent replications. The value of comparative studies between these regions lies not with only their similarities, but with subtle differences in climatic conditions, topographic diversity, evolutionary history, and human impacts that have led to the patterns of disturbance that we see today. There has been a long history of comparative ecosystem studies between California and Chile, and between South Africa and southwestern Australia, but these have traditionally devoted little attention to comparing and contrasting how natural disturbance regimes have been modiﬁed by human activities. The remarkable biodiversity of MTEs together with the large numbers of rare and endangered species in these regions gives a special signiﬁcance to studies of disturbance regimes in these regions. Serious threats of habitat transformation and degradation today make it critical that there be a better understanding of the conservation biology and sustainable resource management in all ﬁve MTEs. REFERENCES Anderson, K., 1993. Native Californians as ancient and contemporary cultivators. In: T.C. Blackburn and K. Anderson (Editors), Before the Wilderness: Environmental Management by Native Californians. Ballena Press, Menlo Park, pp. 151–174. Armstrong, A.J. and van Hensbergen, H.J., 1994. Comparison of avifaunas in Pinus radiata habitats and indigenous riparian habitat at Jonkershoek, Stellenbosch. S. Afr. J. Wildl. Res., 24: 48–55. Aronson, J., Ovalle, C. and Avenda˜no, J., 1993. Ecological and economic rehabilitation of degraded “espinales” in the subhumid mediterranean-climate region of central Chile. Landscape and Urban Planning, 24: 15–21. Aronson, J., del Pozo, A., Ovalle, C., Avenda˜no, J., Lavin, A. and Etienne, M., 1998. Land use changes and conﬂicts in central Chile. In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in Mediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 155–168. Aschmann, H., 1973a. Distribution and peculiarity of mediterranean ecosystems. In: F. di Castri and H.A. Mooney (Editors),

DISTURBANCE IN MEDITERRANEAN-CLIMATE SHRUBLANDS AND WOODLANDS Mediterranean-Type Ecosystems: Origin and Structure. SpringerVerlag, New York, pp. 11–19. Aschmann, H., 1973b. Man’s impact on the several regions with Mediterranean climates. In: F. di Castri and H.A. Mooney (Editors), Mediterranean-Type Ecosystems: Origin and Structure. SpringerVerlag, New York, pp. 363–371. Bahre, C., 1979. Destruction of the vegetation in northern central Chile. Univ. Calif. Publ. Geogr., 23: 1–117. Balduzzi, A., Tomaselli, R., Serey, I. and Villase˜nor, R., 1982. Degradation of the mediterranean type of vegetation in central Chile. Ecol. Mediterranea, 8: 223–240. Beard, J.S. and Sprenger, B.S., 1984. Geographical Data from the Vegetation Survey of Western Australia. Part 1 – Area Calculations. Vegmap Publications, Applecross, Western Australia, 62 pp. Blackburn, T.C. and Anderson, K. (Editors), 1993. Before the Wilderness: Environmental Management by Native Californians. Ballena Press, Menlo Park, 476 pp. Bond, W.J. and Slingsby, P., 1984. Collapse of an ant plant mutualism – Argentine ant, Iridomyrmex humilis, and myrmecochorous Proteaceae. Ecology, 65: 1031–1037. Caravello, G.U. and Giacomin, F., 1993. Landscape ecology aspects in a territory centuriated in Roman times. Landscape and Urban Planning, 24: 77–85. Cowling, R.M. (Editor), 1992. The Ecology of Fynbos: Nutrients, Fire and Diversity. Oxford University Press, Cape Town, 411 pp. Cowling, R.M., Holmes, P.M. and Rebelo, A.M., 1992. Plant diversity and endemism. In: R.M. Cowling (Editor), The Ecology of Fynbos: Nutrients, Fire and Diversity. Oxford University Press, Cape Town, pp. 62–112. Cowling, R.M., Rundel, P.W., Lamont, B.B., Arroyo, M.K. and Arianoutsou, M., 1996. Plant diversity in Mediterranean-climate regions. Trends Ecol. Evol., 11: 352–360. D’Antonio, C.M., Odion, D.C. and Tyler, C.M., 1993. Invasion of maritime chaparral by the introduced succulent Carpobrotus edulis: the roles of ﬁre and herbivory. Oecologia, 95: 14–21. Deacon, H.J., 1992. Human settlement. In: R.M. Cowling (Editor), 1992. The Ecology of Fynbos: Nutrients, Fire and Diversity. Oxford University Press, Cape Town, pp. 260–270. Dell, B. and Malajczuk, N., 1989. Jarrah dieback – a disease caused by Phytophora cinnamomi. In: B. Dell, J.J. Havel and N. Malajczuk (Editors), The Jarrah Forest. A Complex Mediterranean Ecosystem. Kluwer, Dordrecht, pp. 67–87. di Castri, F. and Mooney, H.A. (Editors), 1973. Mediterranean-Type Ecosystems: Origin and Structure. Springer-Verlag, New York, 405 pp. di Castri, F., Goodall, D.W. and Specht, R.L. (Editors), 1981. Mediterranean-Type Shrublands. Ecosystems of the World 11. Elsevier, Amsterdam, 643 pp. Esler, K.J., Rundel, P.W. and Cowling, R.M., 1998. Biodiversity and conservation biology of coastal transition zones from mediterranean to desert ecosystems: an intercontinental comparison. In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in MediterraneanTypMediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 205–230. Etienne, M., Fa´undez, L. and Vald´es, J., 1986. Evaluaci´on de la tasa de desertiﬁcaci´on en la zona a´ rida de Chile Central. Ambiente y Desarrollo, 2: 105–112. Etienne, M., Aronson, J. and Le Floc’h, E., 1998. Abandoned land

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and land use conﬂicts in southern France. In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in Mediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 127–140. Fantechi, R. and Margaris, N. (Editors), 1986. Desertiﬁcation in Europe. D. Reidel, Dordrecht, 231 pp. Farina, A., 1989. Recent changes of the mosaic patterns in a montane landscape (North Italy) and consequences on vertebrate fauna. In: J. Baudry and R.G.H. Bunce (Editors), Land Abandonment and its Role in Conservation. Proceedings of the Zaragossa/Spain Seminar 10–12 December 1989. Opt. Mediterraneennes Ser. A, 15: 121–134. Farina, A., 1998. Bird diversity in a changing landscape (Tuscany, Italy). In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in Mediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 349–367. Fox, B.J. and Fox, M.D., 1986. Resilience of animal and plant communities to human disturbance. In: B. Dell, A.J.M. Hopkins and B.B. Lamont (Editors), Resilience in Mediterranean-Type Ecosystems. Junk, Dordrecht, pp. 39–64. Frenkel, R.E., 1970. Ruderal vegetation along some California roadsides. Univ. Calif. Publ. Geogr., 20: 1–163. Fuentes, E., Hoffmann, A.J., Poiani, A. and Alliende, M.C., 1986. Vegetation change in large clearings: patterns in the Chilean matorral. Oecologia, 68: 358–366. Fuentes, E.R. and Espinosa, G., 1986. Resilience of central Chile shrublands: a vulcanism-related hypothesis. Interciencia, 11: 164–165. Fuentes, E.R. and Hajek, E.R., 1979. Patterns of modiﬁcation in relation to agricultural practice in central Chile. Environ. Conservation, 6: 265–271. Fuentes, E.R., Jaksic, F. and Simonetti, J., 1983. European rabbits vs. Native mammals: effects on shrub seedlings. Oecologia 58: 411–414. Giliomee, J.H., 1986. Seed dispersal by ants in the Cape Flora threatened by Iridomyrmex humilis (Hymenoptera: Formicidae). Entomol. Gen., 11: 217–219. Glyphis, J.P., Milton, S.J. and Siegfried, W.R., 1981. Dispersal of Acacia cyclops by birds. Oecologia, 48: 138–141. Groves, R.H., 1991. The biogeography of mediterranean plant invasions. In: R.H. Groves and F. di Castri (Editors), Biogeography of Mediterranean Invasions. Cambridge University Press, Cambridge, pp. 427–438. Groves, R.H. and di Castri, F. (Editors), 1991. Biogeography of Mediterranean Invasions. Cambridge University Press, Cambridge, 485 pp. Guillerm, J.L., Le Floc’h, E., Maillet, J. and Boulet, C., 1990. The invading weeds within the western Mediterranean Basin. In: F. di Castri, A.J. Hansen and M. Debussche (Editors), Biological Invasions in Europe and the Mediterranean Basin. Kluwer, Dordrecht, pp. 61–84. Gulmon, S.L., 1977. A comparative study of grasslands of California and Chile. Flora, 166: 261–278. Heady, H.F., 1977. Valley grassland. In: M.G. Barbour and J. Major (Editors), Terrestrial Vegetation of California. John Wiley, New York, pp. 491–514. Hobbs, R.J. (Editor), 1992. Biodiversity of Mediterranean Ecosystems in Australia. Surrey Beatty & Sons, Chipping Norton, Australia, 245 pp.

284 Hobbs, R.J., 1998. Impacts of land use on biodiversity in Southwestern Australia. In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in Mediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 81– 106. Hobbs, R.J., Saunders, D.A. and Arnold, G.W., 1993. Integrated landscape ecology: a Western Australian perspective. Biol. Conservation, 64: 231–238. Jackson, L.E., 1985. Ecological origins of California’s mediterranean grasses. J. Biogeogr., 12: 349–361. Kaplan, D., 1992. Responses of Mediterranean grassland plants to gazelle grazing. In: C.A. Thanos (Editor), Proc. 6th Int. Conf. on Mediterranean Climate Ecosystems, September 1991. University of Athens, Athens, pp. 75–79. Keighery, G.J., 1995. How many weeds? In: G. Burke (Editor), Invasive Weeds and Regenerating Ecosystems in Western Australia. Institute for Science and Technology Policy, Murdoch University, Perth, pp. 3–12. Kuzucuoglu, C., 1989. Fires in the Mediterranean region. Blue Plan. Ecology, 72: 371–412. Le Hou´erou, H.N., 1981. Impact of man and his animals on mediterranean vegetation. In: F. di Castri, D.W. Goodall and R.L. Specht (Editors), Mediterranean-Type Shrublands. Ecosystems of the World 11. Elsevier, Amsterdam, pp. 479–521. Le Maitre, D.C., van Wilgen, B.W., Chapman, R.A. and McKelly, D.H., 1996. Invasive plants and water resources in the Western Cape Province, South Africa: Modelling the consequences of a lack of management. J. Appl. Ecol., 33: 161–172. Manders, P.T. and Richardson, D.M., 1992. Colonization of Cape fynbos communities by forest species. For. Ecol. Manage., 48: 277–293. McFarlane, D.J., George, R.J. and Farrington, P., 1992. Changes in the hydrologic cycle. In: R.J. Hobbs and D.A. Saunders (Editors), Reintegrating Fragmented Landscapes: Towards Sustainable Production and Nature Conservation. Springer-Verlag, New York, pp. 146–186. Meeus, J., Wijermans, M.P. and Vroom, M.J., 1990. Agricultural landscapes in Europe and their transformation. Landscape and Urban Planning, 18: 289–3352. Minnich, R.A., 1983. Fire mosaics in southern California and northern Baja California. Science, 219: 1287–1994. Montenegro, G., Tellier, S., Arce, P. and Poblete, V., 1991. In: R.H. Groves and F. di Castri (Editors), Biogeography of Mediterranean Invasions. Cambridge University Press, Cambridge, pp. 103–113. Mooney, H.A. and Conrad, C.E. (Editors), 1977. Symposium on the Environmental Consequences of Fire and Fuel Management in Mediterranean Ecosystems. USDA Forest Service General Technical Report WO-3, 498 pp. Mooney, H.A., Hamburg, S.P. and Drake, J.A., 1986. The invasions of plants and animals into California. In: H.A. Mooney and J.A. Drake (Editors), Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York, pp. 250–272. Moreno, J.M. and Oechel, W.C. (Editors), 1994. The Role of Fire in Mediterranean-Type Ecosystems. Springer-Verlag, New York, 201 pp. Musil, C.F., 1993. Effect of invasive Australian acacias on the regeneration growth and nutrient chemistry of South African lowland fynbos. J. Appl. Ecol., 30: 361–372. Naveh, Z., 1982. Mediterranean landscape evolution and degradation

Philip W. RUNDEL as multivariate biofunctions: theoretical and practical implications. Landscape and Urban Planning, 9: 125–146. Naveh, Z., 1990. Fire in the Mediterranean: a landscape perspective. In: J.G. Goldhammer and M.J. Jenkins (Editors), Fire in Ecosystem Dynamics. SPB Academic Publishing, The Hague, pp. 401–434. Naveh, Z., 1991. Mediterranean uplands as anthropogenic perturbation dependent systems and their dynamic conservation management. In: O.A. Ravera (Editor), Terrestrial and Aquatic Ecosystems, Perturbation and Recovery. Ellis Horwood, New York, pp. 544–556. Naveh, Z., 1998. From biological diversity to ecodiversity: holistic conservation of biological and cultural diversity of Mediterranean landscapes. In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in Mediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 23–53. Naveh, Z. and Dan, J., 1973. The human degradation of Mediterranean landscapes in Israel. In: F. di Castri and H.A. Mooney (Editors), Mediterranean-Type Ecosystems: Origin and Structure. Springer-Verlag, New York, pp. 373–390. Naveh, Z. and Lieberman, A.S., 1993. Landscape Ecology – Theory and Applications, 2nd edition. Springer-Verlag, New York. Naveh, Z. and Whittaker, R.H., 1979. Structural and ﬂoristic diversity of shrublands and woodlands in northern Israel and other mediterranean areas. Vegetatio, 41: 171–190. Noble, J.C. and Bradstock, R.A., 1989. Mediterranean Landscapes in Australia: Mallee Ecosystems and their Management. CSIRO, East Melbourne, 485 pp. Nulsen, R.A., 1992. Changing soil properties. In: R.J. Hobbs and D.A. Saunders (Editors), Reintegrating Fragmented Landscapes: Towards Sustainable Production and Nature Conservation. Springer-Verlag, New York, pp. 107–145. Ovalle, C., Aronson, J., del Pozo, A. and Avenda˜no, J., 1990. The espinal: agroforestry systems of the mediterranean-type climate region of Chile. Agrofor. Syst., 10: 213–239. Ovalle, C., Avenda˜no, J., del Pozo, A. and Aronson, J., 1996. Land occupation patterns and vegetation structure of the anthropogenic savannas (espinales) of central Chile. For. Ecol. Manage., 86: 129–139. Papanastasis, V. and Kazaklis, A., 1998. Land use changes and conﬂicts in the mediterranean-type ecosystems of western Crete. In: P.W. Rundel, G. Montenegro and F. Jaksic (Editors), Landscape Disturbance and Biodiversity in Mediterranean-Type Ecosystems. Springer-Verlag, Berlin, pp. 141–154. Pate, J.S. and Beard, J.S., 1984. Kwongan: Plant Life of the Sand Plains. University of Western Australia Press, Nedlands, 283 pp. Peck, A.J., 1978. Salinization of non-irrigated soils and associated streams: a review. Aust. J. Soil Res., 16: 157–168. Perez-Trejo, F., 1994. Desertiﬁcation and Land Degradation in the European Mediterranean. Ofﬁce for Ofﬁcial Publications of the European Communities, Luxembourg, 63 pp. Pinto-Correia, T., 1993. Threatened landscapes in Alentejo, Portugal: the montado and other agro–silvo–pastoral systems. Landscape and Urban Planning, 24: 43–48. Polunin, O. and Walters, M., 1985. A Guide to the Vegetation of Britain and Europe. Oxford University Press, Oxford, 228 pp. Rejm´anek, M., Thomsen, C.D. and Peters, I.D., 1991. Invasive vascular plants in California. In: R.H. Groves and F. di Castri (Editors), Biogeography of Mediterranean Invasions. Cambridge University Press, Cambridge, pp. 81–101.

DISTURBANCE IN MEDITERRANEAN-CLIMATE SHRUBLANDS AND WOODLANDS Richardson, D.M., Cowling, R.M. and Le Maitre, D.C., 1990. Assessing the risk of invasive success in Pinus and Banksia in South African mountain fynbos. J. Vegetation Sci., 1: 629–642. Richardson, D.M., MacDonald, I.A.W., Holmes, P.M. and Cowling, R.M., 1992. Plant and animal invasions. In: R.M. Cowling (Editor), The Ecology of Fynbos: Nutrients, Fire and Diversity. Oxford University Press, Cape Town, pp. 271–308. Roy, J., Navas, M.L. and Soni´e, L., 1991. Invasion by annual brome grasses: a case study challenging the homocline approach to invasions. In: R.H. Groves and F. di Castri (Editors), Biogeography of Mediterranean Invasions. Cambridge University Press, Cambridge, pp. 207–224. Rundel, P.W., 1981a. The matorral zone of central Chile. In: F. di Castri, D.W. Goodall and R.L. Specht (Editors), Mediterranean-Type Shrublands. Ecosystems of the World 11. Elsevier, Amsterdam, pp. 175–201. Rundel, P.W., 1981b. Fire as an ecological factor. In: O.L. Lange, P.S. Nobel, C.B. Osmond and H. Ziegler (Editors), Physiological Plant Ecology I: Response to the Physical Environment. SpringerVerlag, Berlin, pp. 501–538. Rundel, P.W., 1983. Impact of ﬁre on nutrient cycles in mediterraneantype ecosystems with reference to chaparral. In: F.J. Kruger, D.T. Mitchell and J.U.M. Jarvis (Editors), Mediterranean-Type Ecosystems: The Role of Nutrients. Springer-Verlag, Berlin, pp. 192–207. Rundel, P.W., 1988. Leaf structure and nutrition in Mediterraneanclimate sclerophylls. In: R.L. Specht (Editor), Mediterraneantype Ecosystems: A Data Source Book. Kluwer Academic Publ., Dordrecht, pp. 157–167. Rundel, P.W., 1995. Adaptive signiﬁcance of some morphological and physiological characteristics in Mediterranean plants: facts and fallacies. In: J. Roy, J. Aronson and F. di Castri (Editors), Time Scales of Biological Responses to Water Constraints: The Case of Mediterranean Biota. SPB Academic Publishing, Amsterdam, pp. 119–139.

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Chapter 11

GRAZING, FIRE, AND CLIMATE EFFECTS ON PRIMARY PRODUCTIVITY OF GRASSLANDS AND SAVANNAS M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

INTRODUCTION

Grasslands and savannas share a number of biotic and abiotic features that differentiate them from other vegetation types. Their major characteristic is the dominance of a herbaceous layer largely composed of grasses and sedges (Walter, 1977). In grasslands, the herbaceous component forms a unique vegetation layer, whereas in savannas there is an additional woody layer which can range from sporadic, isolated shrubs or trees to relatively dense woodlands. Climatically, grasslands and savannas extend over a broad range of mean annual precipitation (200–1300 mm) and temperature (0–30ºC) (Lauenroth, 1979). However, all of them have a negative water balance and undergo marked dry seasons (Lauenroth, 1979; McNaughton et al., 1982). Grasslands and savannas are fuzzy segments of a climatic gradient from deserts through closed forests. In many regions of the world, deserts turn gradually into grasslands, as grasses become more important in response to an increase in mean precipitation (Paruelo and Lauenroth, 1996; Paruelo et al., 1998). A southeast–northwest gradient in eastern Europe from Artemisia desert to Stipa steppe (Walter, 1977), an east–west gradient in southern South America, from Nassauvia glomerulosa semi-desert to the western steppe with Festuca pallescens (Le´on and Facelli, 1981), and a west–east gradient in North America from the Chihuahuan Desert with Larrea tridentata to shortgrass steppe with Bouteloua gracilis (Gosz, 1993) are all examples of this transition. At the other end of the gradient, dense grasslands and savannas are replaced by forests. In some regions, this transition is sharp enough for no intermediate regions with savannas to be distinguishable at the scale of world or continental maps. For example, in Patagonia, Festuca pallescens

grassland turns into a Nothofagus forest (Soriano, 1983; Schulze et al., 1996). In other regions, particularly in the tropics, there is a gradual increase in the tree/grass ratio as precipitation increases: grasslands are progressively replaced by savannas whose woody component becomes more important as precipitation increases (Sinclair, 1979; Walker and Noy-Meir, 1982; McNaughton, 1983a; Belsky, 1990). The broad range of mean annual precipitation of the grassland/savanna biome is one of the most important causes of its structural and functional diversity. Plant cover, plant biomass, leaf area, and canopy height of the herbaceous layer predictably increase along gradients of increasing precipitation. For example, the heterogeneity of the grassland region of North America has been repeatedly described on the basis of a gradient from the relatively dry, shortgrass steppe in the West through the more humid, tallgrass prairie in the East (Coupland, 1992). Similar gradients are observed in southern South America (Oesterheld et al., 1992) and East Africa (McNaughton, 1983a). Parallel to these changes in structural features, ecosystem function also changes dramatically and predictably along the precipitation gradient. In particular, more than 75% of the variation in above-ground net primary production of the herbaceous layer of grasslands and savannas can be accounted for by their mean annual precipitation (McNaughton, 1985; Sala et al., 1988a; McNaughton et al., 1993). Precipitation also explains a substantial proportion of the seasonal variability of carbon gains (Paruelo and Lauenroth, 1996, 1998). Taken together, these observations suggest that mean annual precipitation causes a great deal of variation among grasslands and savannas, but, for this same reason, it becomes a highly integrative variable to be

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used when generalizations about the whole biome are needed. This conceptual gradient of precipitation will be an important framework for this chapter. In addition to this variation associated with mean annual precipitation and primary productivity, there is a ﬁner scale of variation associated with soil types. The gently rolling landscape of many grasslands and savannas determines wide differences in hydrologic regime, nutrient cycling, vegetation structure, and ecosystem function without difference in mean annual precipitation (McNaughton, 1983a; Schimel et al., 1985; Soriano, 1992; Corona et al., 1995). At a broader scale, grasslands and tropical savannas differ signiﬁcantly as a result of soil texture and mineralogy, with important consequences not only for vegetation structure and composition, but also for herbivore populations (Huntley, 1982; East, 1984; Epstein et al., 1997). In addition to the important role of mean annual precipitation and soil type as determinants of grassland and savanna structure and function, there is the role of disturbance regimes. The world-wide expansion of grasslands and savannas since the Miocene has been paralleled by the evolution of large mammalian grazers and browsers (McNaughton et al., 1993), the occurrence of ﬁre (Vogl, 1974; Collins, 1990), and the marked climatic ﬂuctuations that characterize these subhumid and semiarid environments (Anderson, 1982). Thus, three types of disturbance are particularly important in shaping the structure and function of grasslands and savannas. First, the type and magnitude of herbivore load determines a pattern of biomass removal, trampling, defecation, and urination, with profound consequences for the entire system (McNaughton, 1983a; Detling, 1988; McNaughton et al., 1988). For example, grazing can turn a tall grassland into a short one (McNaughton, 1984; Facelli, 1988), and its exclusion may have opposite effects (McNaughton, 1984; Sala et al., 1986). Second, the frequency and intensity of ﬁres determine a pattern of biomass removal, nutrient volatilization, and ash deposition which affect the whole ecosystem (Daubenmire, 1968; Vogl, 1974; Hulbert, 1988; Hobbs et al., 1991). For example, in Australian savannas, ﬁre can remove up to 94% of nitrogen, 53% of phosphorus, and 82% of potassium present in plant biomass (Cook, 1994); and, in the tallgrass prairie of North America, a single ﬁre event can remove approximately twice the annual

input of nitrogen (Seastedt, 1995). Finally, year-toyear changes, and even larger-scale trends in climatic variables, can affect water availability, a major driving force of the structure and function of these systems (Risser, 1985; Le Hou´erou et al., 1988; Briggs et al., 1989; Lauenroth and Sala, 1992). These disturbance factors become particularly important at the boundaries of the biome because they may be responsible for changing the biome status of the system: at the drier boundary, grazing and climatic variations have been repeatedly identiﬁed as likely causes of desertiﬁcation, the transformation of a grassland into a system with desert features (Dodd, 1994), whereas at the humid end of the gradient interactions among grazing, ﬁre, and climate have been identiﬁed as the cause for woody-plant encroachment, and the transformation of a grassland or savanna into a closed woodland or forest (Archer et al., 1988; Archer, 1989, 1995). The objective of this chapter is to show the major effects of these three disturbances on grasslands and savannas of the world. This subject has been reviewed many times in the past (e.g., Daubenmire, 1968; Vogl, 1974; Anderson, 1982; Risser, 1985; Detling, 1987, 1988; Collins, 1990). Our contribution will focus on those aspects that we believe have received less attention and that we are in the position of analyzing now because of the existence of new data and the development of new ideas. Most reviews have concentrated on structural aspects of vegetation at the community level while neglecting the more functional, ecosystem-level attributes. Thus, we have chosen a single functional response variable as the center of our chapter: aboveground net primary productivity (ANPP). In addition, most researchers, while agreeing on the importance of grazing, ﬁre, and climate ﬂuctuations as agents of disturbance, have tended to treat them separately, so that a measure of their relative importance is lacking. Thus, we have attempted to study their effects on productivity in comparable ways, so that they can be ranked. Finally, we believe that there have been good reviews focusing on particular grassland/savanna regions, but less effort has been devoted to integrating patterns for all kinds of systems. Thus, we have framed all our analysis in the context of a gradient of annual precipitation that encompasses a wide range of grassland and savanna ecosystems. Naturally, this approach has tradeoffs. We will lose spatial detail and diversity of response variables. The rest of the chapter is organized in the following way: ﬁrst, we will describe the disturbance regime of grasslands and savannas;

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

second, we will show how disturbances affect aboveground net primary productivity.

DISTURBANCE REGIMES

The frequency and intensity of grazing, ﬁre, and climatic ﬂuctuations change with mean annual precipitation. Because precipitation is linearly related to above-ground net primary productivity, mean annual precipitation is associated with both the availability of forage, which partially determines grazing regime, and the production of ﬂammable fuel, which partially determines ﬁre regime. Grazing and ﬁre are alternative consumers of productivity: grazing may preclude the accumulation of fuel and completely suppress ﬁre, and ﬁre, in turn, may either consume productivity that could be used by herbivores or change its quality as forage (Kucera, 1981). Finally, mean annual precipitation is associated with year-to-year variations in precipitation. Thus, there is a gradient of mean annual precipitation along which the likelihood of a system being grazed, burned, or struck by drought or exceptionally wet conditions, vary.

289

argued that it could stem from the fact that the most productive end of the data set was heavily inﬂuenced by data points from tropical grasslands and savannas of East Africa, with their high density of megafauna, whereas the low end of the productivity gradient was inﬂuenced by temperate grasslands that currently lack an important population of large herbivores because of hunting and other habitat alterations by humans. The key question concerns the extent to which the exponential increase of consumption is an intrinsic property of the productivity gradient or simply a result of human-induced extinctions. We extracted from the data set of McNaughton et al. (1989, 1991) only the data points classiﬁed as grasslands and savannas (Fig. 11.1). The relationship between consumption and productivity, which should be regarded here as the pattern of consumption along the gradient of mean annual precipitation, was of the form: log C = −5.16 + 2.2 log ANPP 2 (11.2) r = 0.75, P < 0.0001, d.f . = 42 This relationship is not signiﬁcantly different from the

Grazing regime It may not be surprising that populations of herbivores and the amount of energy they consume increase with productivity across a wide range of ecosystem types, from deserts through tropical forests (McNaughton et al., 1989, 1991). However, it is more intriguing that different ecosystem types form a single cohesive function, and that the shape of that function is exponential: as ecosystems become more productive, they have a growing herbivore load per unit of primary production and, as a result, a growing proportion of their productivity is consumed by herbivores (McNaughton et al., 1989, 1991). Considering only foliage production (i.e., excluding wood), consumption by herbivores across all these ecosystem types dominated by native herbivores was related to production by: log C = −4.8 + 2.04 log NFP 2 r = 0.59, P < 0.0001, d.f . = 73

(11.1)

where C is consumption, NFP is net foliage production, and units are kJ m−2 yr−1 (McNaughton et al., 1989). Being signiﬁcantly greater than 1, the slope of this log–log relationship indicates an exponential shape. Discussing this result, McNaughton et al. (1991)

Fig. 11.1. Relationship between herbivore consumption and aboveground net primary production (ANPP) of grasslands and savannas dominated by native herbivores. Solid squares are tropical sites (n = 31) dominated by vertebrates and open circles are temperate sites (n = 14) dominated by invertebrates (from McNaughton et al., 1989, 1991). The solid line is the best-ﬁt line for both types of system taken together. Open squares are sites (n = 9) from Yellowstone National Park, a temperate, vertebrate-dominated system (from Frank and McNaughton, 1992). The two parallel dashed lines are best-ﬁt lines for vertebrate-dominated systems (top) and invertebrate-dominated systems (bottom).

290

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

general relationship obtained for all ecosystem types, but it is stronger, as shown by its higher coefﬁcient of determination (r 2 = 0.75, as against 0.59 for all ecosystem types). However, Fig. 11.1 shows that this function actually connects two parallel point clouds: a highconsumption cloud formed by the tropical, vertebratedominated grasslands, and a low-consumption cloud formed by the temperate, invertebrate-dominated grasslands. Frank and McNaughton (1992) have provided a crucial handful of data points by measuring consumption along a productivity gradient in a temperate ecosystem dominated by large herbivores: Yellowstone National Park with its dominant population of elk (Cervus canadensis). They showed that consumption data for that system perfectly ﬁt the line of the tropical, vertebrate-dominated systems, suggesting that the highly exponential nature of the general relationship was indeed inﬂuenced by the lack of large herbivores in the temperate grasslands included in the original data set. When considered separately, the functions relating consumption and productivity of vertebrate-dominated and invertebrate-dominated systems are much less exponential than the general relationship: log C = −1.54 + 1.32 log ANPP 2 r = 0.83, P < 0.0001, d.f . = 38

(11.3a)

for vertebrate systems, and log C = −2.08 + 1.26 log ANPP 2 r = 0.39, P < 0.02, d.f . = 12

(11.3b)

for invertebrate systems. These relationships indicate that consumption, as a percentage of productivity, grows from 30 to 75% across the productivity gradient of vertebrate-dominated systems, but only from 5 to 10% across the productivity gradient of invertebratedominated systems (Fig. 11.2). The relationships shown in Figs. 11.1 and 11.2, however, seem to be inﬂuenced by spatial scale. Independent data sets on herbivore densities and biomass suggest that the consumption data discussed above represent areas within regions with particularly high levels of herbivory. Consumption rates derived from large-scale surveys of herbivore density are much lower. The consumption data discussed above were extracted from studies aimed at exploring the effects of herbivory on various aspects of community (and ecosystem) structure and function, and it is reasonable to expect that they were conducted in areas where herbivory was particularly important. For example,

Fig. 11.2. Percentage of ANPP (above-ground net primary production) consumed by herbivores along a gradient of ANPP as predicted by the two parallel models of Fig. 11.1. Top, thick line: systems dominated by vertebrates. Bottom, thin line: systems dominated by invertebrates.

McNaughton (1985) explicitly stated that his regional study sites were located in areas of the Serengeti with high animal utilization. When animal biomass is measured for large regions, such as an entire national park or reserve, however, the data suggest that consumption rates are lower. Coe et al. (1976) and East (1984) reported such data sets, included also in McNaughton et al. (1989, 1991), and showed that herbivore biomass of East African grasslands and savannas varies roughly between 10 and 50 kg ha−1 and only rarely reaches values near 100 kg ha−1 . Considering that a 300 kg ruminant consumes approximately 11 kg kg−1 yr−1 of dry matter, consumption in grasslands with herbivore biomass between 10 and 50 kg ha−1 should be around 110 to 550 kg ha−1 yr−1 . The range of productivity for those grasslands is 2400–12 000 kg ha−1 , which indicates a 5 to 10% consumption rate. A similar pattern arises from Yellowstone. Herbivore biomass in the grassland/shrubland areas of the park is estimated as 36.9 kg ha−1 and is largely dominated by elk (D. Frank, pers. commun.). Thus, consumption should be around 400 kg ha−1 yr−1 , nearly an order of magnitude lower than the maximum reported by Frank and McNaughton (1992) for particular study sites. Densities of predators change with scale in the same way as those of mammalian herbivores; density of mammalian carnivores has a close negative relationship to the size of the study area (Smallwood and Schonewald, 1996). Most of the area covered by grasslands and savannas of the world, however, is populated by livestock

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

rather than by native herbivores, and the patterns of consumption in that situation differ in some respects from those just described. Livestock biomass per unit area at a county level across a wide range of mean annual precipitation in rangelands of southern South America (Argentina and Uruguay) increased exponentially with precipitation, and was an order of magnitude greater than herbivore biomass of native systems around the world (Oesterheld et al., 1992). Rough calculations of consumption based on these data for livestock biomass indicate that percent consumption along the productivity gradient ranged from 20% to 75%. Thus, animal husbandry on rangelands has made an entire region look very much like the relatively small portions of a landscape such as the Serengeti National Park or Yellowstone that are heavily grazed by native herbivores. These results suggest that (1) herbivore biomass and consumption increase along the precipitation gradient in an exponential fashion; (2) the relationship is very different among vertebrate-dominated and invertebratedominated systems; (3) consumption rates measured directly in particular communities or study sites likely represent the upper boundary of a range with great spatial variability; (4) consumption rates inferred from large-scale herbivore densities are much lower and likely reﬂect the actual average values that result from that spatial variability; and (5) livestock biomass and plant consumption also increase exponentially with precipitation but animal husbandry makes them both drastically larger and spatially more uniform within extended regions of grasslands and savannas.

291

which will be presented later, lacks data for grasslands below 400 mm yr−1 . At the other extreme of the gradient, annual burning is a common practice both in research and management. It is less clear, however, what mechanisms underlie the pattern of increased ﬁre with increased precipitation. One mechanism, which has been repeatedly invoked, is that as precipitation increases so does the difference between production and losses of biomass, resulting in an increasing net accumulation of ﬂammable fuel (standing dead material plus litter). This would be so if biomass losses by grazing and decomposition increased less with precipitation than primary productivity. Regarding grazing, we have shown in the previous section (Figs. 11.1 and 11.2) that consumption by herbivores accounts for an increasing proportion of productivity as precipitation increases. We calculated the amount of above-ground primary productivity which would remain unconsumed on an annual basis according to those patterns (Fig. 11.3). Unconsumed productivity indeed increases along a gradient of mean annual productivity, as this mechanism for ﬁre-

Fire regime There is a high degree of consensus about the way in which ﬁre frequency varies along the gradient of annual precipitation that characterizes grasslands and savannas. In order to occur, ﬁre requires a minimum amount of ﬂammable fuel and the proper conditions for ignition and spread (Vogl, 1974). Since the rate of production of biomass (productivity) increases linearly with precipitation, ﬁre frequency is thought to increase monotonically along the mean annual precipitation gradient that we have been analyzing (Kucera, 1981; Frost and Robertson, 1987; Risser, 1990). For example, researchers have experienced serious difﬁculties studying ﬁres in areas with precipitation below 450 mm yr−1 because of repeated failures with experimental burning (Trollope, 1984). Our own set of data on ﬁre effects,

Fig. 11.3. Annual accumulation of above-ground biomass along a gradient of mean annual ANPP (above-ground net primary production), which is linearly related to precipitation. The dashed line represents the hypothetical accumulation if biomass were neither grazed nor decomposed (equality line). The other three lines with increasing thickness respectively represent the accumulation of biomass left unconsumed by herbivores (derived from the equation for vertebrate systems in Fig. 11.1), the accumulation of biomass left undecomposed without grazers (derived from decomposition rates by Schimel et al., 1990), and the accumulation of unconsumed biomass left undecomposed (derived from a combination of the two previous curves).

292

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

frequency gradients predicts, but the increase is much less steep than would be expected if grazers either were absent or consumed a constant proportion of productivity. Regarding litter decomposition, there is a great deal of evidence for forest ecosystems that the proportion of litter annually lost by microbial decomposition is a positive function of actual evapotranspiration (AET), and a negative function of lignin content (Meentemeyer, 1978; Couteaux et al., 1995). Actual evapotranspiration is a close linear positive function of precipitation in grasslands (Ripley, 1992), and the few available data indicate that lignin content, particularly above-ground, decreases with precipitation (Schimel et al., 1990). Thus a higher, not lower, proportion of litter should be lost annually to decomposers as precipitation increases. Actually, Risser et al. (1981) have shown that, in the North American tallgrass prairie, litter losses by decomposition nearly equal the inputs on an annual basis. The CENTURY model, which includes a number of biogeochemical processes in grasslands (Parton et al., 1987), predicts that annual decomposition rates of the Great Plains in North America should range from 20% at the low end of the precipitation gradient to 85% at the high end (Schimel et al., 1990). We used these data to estimate the amount of above-ground net primary productivity that would remain undecomposed annually in the absence of herbivores across a gradient from the north-west of the region through the south-east (Fig. 11.3). The increase in this accumulation rate was very low: a 270% increase in above-ground net primary productivity, from 150 to 400 g m−2 yr−1 , resulted in an increase of fuel of 60% (deviations from this curve are obtained if the gradient of productivity is considered at low or high latitudes, instead of the NW–SE diagonal we utilized, because of the effect of temperature on decomposition rates). If above-ground net primary productivity exceeds 400 g m−2 yr−1 , decomposition rates increase more than productivity, resulting in a decrease in accumulation of undecomposed aboveground material. If these decomposition rates are applied to the unconsumed productivity, simulating a more real system with herbivores and decomposers, fuel accumulation is obviously lower (Fig. 11.3). Figure 11.3 also reveals that the CENTURY model predicts a decrease in the above-ground inputs to the soil pool of carbon from subhumid to humid systems. However, belowground inputs continuously increase along the gradient (Schimel et al., 1990),

and seem to be responsible for the associated increase in soil organic matter with increasing precipitation (Burke et al., 1989). Thus, high ﬁre frequency in humid grasslands and savannas does not seem to be a consequence of litter accumulation through years due to larger differences between productivity and losses by grazing and decomposition. Standing-dead accumulation during the current year, instead, seems to account for regional patterns of ﬁre frequency. Another set of mechanisms consistent with predictions of an increase in ﬁre frequency with mean annual precipitation or productivity has been provided by Wedin (1995). He proposed the existence of strong feedbacks involving nitrogen, soils, plants, herbivores, and ﬁre, and related them to disturbance regimes. His conceptual model is integrative and thought-provoking, and is highly relevant here because it covers the same gradient from low precipitation and low productivity to high precipitation and high productivity that we have been considering. The model assumes that, along this gradient of increasing precipitation and productivity, animal biomass and forage quality decline. This unexplained assumption leads to predictions about the major consumers of productivity (herbivores, decomposers, and ﬁre) and, as a consequence, about disturbance regimes. The model stresses the importance of carbon and nitrogen stoichiometry of plants in regulating the consumer pathway that primary production will follow in a particular ecosystem. Unlike ﬁre, herbivores and decomposers have a minimum nitrogen requirement to consume biomass. Thus, grazing and ﬁre regimes operating along the gradient should depend on the carbon/nitrogen ratio of the dominant grasses. Consequently, one would expect herbivory and decomposition to be the major consumers in low-production systems with high quality of forage and litter, and ﬁre to be the major “consumer” in high-production ecosystems with lower quality of forage and litter. We have shown in the previous section that the pattern of consumption along the gradient is opposite to that assumed by this model. We will concentrate now on the second assumption: that forage quality decreases along the gradient. [This assumption was based on a set of data published by Breman and de Wit (1983) who showed, using an adjusted curve, that protein concentration and biomass production were inversely correlated in the Sahel region of Africa.] We compiled information from the literature about nitrogen content in biomass of grasslands and savannas

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

located along a broad range of mean annual precipitation to test this assumption (Table 11.1). Approximately 200 potentially relevant articles were surveyed, and 28, representing 61 data sets, were used in the analysis. The general criteria for data selection and analysis were as follows: (1) we selected data from climatically-determined native grasslands or savannas (alpine communities, anthropogenic pastures, and wetlands were excluded); (2) these data had to be taken from the ﬁeld; (3) data from the same site but obtained on different dates within a year were pooled and averaged; (4) data from studies of individual species were included only when they were a dominant component of the community; (5) the variable used for analysis was nitrogen concentration in above-ground green biomass. In the cases reported by Turner et al. (1993) and Jackson et al. (1990), this variable was estimated from nitrogen concentration of the above-ground dead biomass according to a linear regression obtained from cases where both variables were reported: y = 0.535 + 1.06x; 2 r = 0.7; P = 0.0003;

d.f . = 12

(11.4)

where y is nitrogen concentration in green biomass and x is nitrogen concentration in dead biomass; (6) data used by Breman and de Wit (1983) were not included in the analysis because they were not available (their paper only presented an adjusted curve). Figure 11.4 shows that the two variables were not signiﬁcantly correlated: [N] = 1.59 − 0.00024 × ppt, 2 r = 0.026; P = 0.20; d.f . = 60

(11.5)

where [N] is nitrogen concentration in above-ground biomass and ppt is mean annual precipitation. This result does not support Wedin’s assumption. Multiple regression models including grazing regime (ungrazed or grazed) or the proportion of C3 vs. C4 plants did not add to the proportion of the variance being explained. Figure 11.4 suggests that greater productivity of more humid systems does not necessarily involve a trade-off with nitrogen concentration, which is related to forage quality and presumably to litter quality (Wedin, 1995). High ﬁre frequency in more productive grasslands is neither a consequence of lower nitrogen content of the forage, which would reduce herbivore consumption,

293

Fig. 11.4. Nitrogen content (%) in above-ground green biomass along a gradient of mean annual precipitation of grasslands and savannas.

nor of low litter quality, which would reduce decomposition. Wedin’s model has been extremely helpful in leading the way to an integrated explanation for disturbance regimes of grasslands and savannas, but a fully satisfactory picture is still lacking. Perhaps too much importance is being given to the annual balance between production and decomposition to explain ﬁre frequency, and, in any case, this exaggerated importance is being applied to the wrong side of the gradient. In order to burn, very productive grasslands do not need to accumulate litter for many years. On the contrary, a highly productive season followed by a dormant season is enough. For example, in a North American tallgrass prairie, grasslands burnt in March were able to carry a second ﬁre in October of the same year (Bragg, 1982). Grasslands with such a high productivity simply need a dormant season that kills most green tissue, an ignition event, and atmospheric conditions that favor the spread of ﬁre (Vogl, 1974). Differences between production and litter decomposition may indeed be very important for ﬁre occurrence on the drier (<600 mm), but not the wetter, side of the gradient. Low productivity will not generate enough ﬂammable biomass in a single season, and only by year-to-year accumulation will these grasslands be able to carry ﬁres. Climatic ﬂuctuations The gradient of grasslands and savannas we have been considering is one of mean annual precipitation,

294

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

Table 11.1 Nitrogen concentration values based on green above-ground biomass 1 (characteristics of information shown in Fig. 11.4) Site

MAP 2

Ungrazed 3

Grazed 4

Mixed grass prairie (North Dakota, USA)

400

0.98

1.19

400

–

1.98

400

1.43

1.13

400

–

1.25

400

1.18

1.57

400

–

1.21

Oak savanna (California, USA)

600

1.92

2.00

Center et al. (1989)

Subhumid grassland (Buenos Aires, Argentina)

960

1.30

1.57

Chaneton et al. (1996)

Shrub steppe (Wyoming, USA)

290

1.13

1.46

Coughenour (1991)

Reference Biondini and Manske (1996)

Shortgrass steppe (South Dakota, USA)

450

–

1.06

Day and Detling (1990)

Shrub steppe (Wyoming, USA)

398

–

2.00

Frank and McNaughton (1992)

Shrub steppe (Wyoming, USA)

398

–

2.18

Frank et al. (1994)

398

–

2.17

398

–

1.81

398

–

1.54

398

–

1.53

Oak savanna (California, USA)

600

–

1.20

Hart et al. (1993)

Tallgrass prairie (Kansas, USA)

835

0.91

1.20

Hobbs et al. (1991)

Shortgrass steppe (Colorado, USA)

310

–

1.20

Hunt et al. (1988)

Oak savanna (California, USA)

600

1.59

–

Jackson et al. (1990)

600

1.70

–

Mixed grass prairie (South Dakota, USA)

338

–

1.20

Jaramillo and Detling (1992)

Mediterranean grassland (Andaluc´ıa, Spain)

648

1.56

–

Joffre (1990)

Semiarid grassland (Thessaloniki, Greece)

215

–

1.53

Mamolos et al. (1995)

215

–

1.19

321

1.76

–

321

–

1.62

321

–

1.65

321

1.70

–

Tallgrass prairie (Kansas, USA)

840

1.20

1.20

Owensby et al. (1993)

Dehesa savanna (Salamanca, Spain)

500

–

1.46

P´erez Corona et al. (1995)

300

–

1.65

Tallgrass prairie (Oklahoma, USA)

840

0.89

0.89

Risser and Parton (1982)

Semiarid steppe (Chubut, Argentina)

150

1.02

1.02

Sala et al. (1991)

Shortgrass steppe (Colorado, USA)

310

–

1.10

Schimel et al. (1985)

310

–

1.08

310

–

1.41

Tallgrass prairie (Kansas, USA)

835

1.19

–

Subhumid grassland (Serengeti, Tanzania)

862

–

1.80

Seagle and McNaughton (1992)

Tallgrass prairie (Oklahoma, USA)

756

–

0.98

Seastedt et al. (1988)

Shortgrass steppe (Colorado, USA)

Milchunas et al. (1995)

Schimel et al. (1991)

continued on next page

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

295

Table 11.1, continued Site

MAP 2

Subhumid grassland (Buenos Aires, Argentina)

960

1.59

1.59

Semmartin and Oesterheld (1996)

1035

1.31

–

Singh (1993)

1.53

–

835

1.74

1.20

835

1.26

1.56

450

–

1.30

–

1.60

–

0.97

Savanna (Sonbhadra, India) Tallgrass prairie (Kansas, USA) Mixed grass prairie (South Dakota, USA) Semiarid steppe (Beer Sheva, Israel) 1 2 3 4

200

Ungrazed 3

Grazed 4

Reference

Turner et al. (1993) Whicker and Detling (1988) Zaady et al. (1996)

Different values within a study represent different years or different topographic positions. Mean Annual Precipitation (mm). [N] of ungrazed plots (%). [N] of grazed plots (%).

which integrates in a single ﬁgure a variable that ﬂuctuates from year to year. In this section we examine whether that variability changes along the gradient of mean annual precipitation of grasslands and savannas. Evidence from such dispersed sources as the North American Great Plains, the Serengeti National Park (Tanzania), and Patagonia (Argentina) clearly shows that absolute interannual variability in precipitation is positively correlated with its mean (Sinclair, 1979; Jobb´agy et al., 1995; Lauenroth and Burke, 1995). The standard deviation of annual precipitation, which provides a measure of absolute variability, increases with mean precipitation. However, the coefﬁcient of variation (CV), which provides a measure of variation relative to the mean, decreases with precipitation. For example, in the North American Great Plains, a grassland with annual precipitation around 200 mm has a standard deviation of 50 mm and a coefﬁcient of variation of 25%, whereas a grassland with annual precipitation around 800 mm has a standard deviation of 150 mm and a coefﬁcient of variation of 19% (Lauenroth and Burke, 1995). In Patagonia, a mean annual precipitation gradient from 150 to 500 mm is associated with a decrease in the coefﬁcient of variation from 40% to 20% (Jobb´agy et al., 1995). Thus, drier grasslands have a greater chance of experiencing more important relative changes in precipitation from year to year than more humid grasslands. The variability of water demand by the atmosphere is much lower than the variability of precipitation. The standard deviation of the mean annual potential

evapotranspiration in the Great Plains ranges from 15 to 30 mm, whereas the standard deviation of precipitation ranges from 50 to 250 mm (Lauenroth and Burke, 1995). The coefﬁcient of variation of the mean annual potential evapotranspiration ranges between 1 and 2.5%, whereas that of the annual precipitation ranges between 15 and 35%. Thus, drier systems have a greater chance of experiencing drought or unusually high water status because of variation in the input rather than in the output of water.

EFFECTS OF GRAZING, FIRE, AND CLIMATE FLUCTUATIONS ON PRODUCTIVITY

Researchers have repeatedly emphasized the importance of grazing, ﬁre, and climatic ﬂuctuations, largely drought, in shaping the structure and function of grasslands and savannas (Anderson, 1982; McNaughton, 1983a; Medina and Silva, 1990; Risser, 1990). In this section, we address the effects of these disturbances on above-ground net primary productivity, a single but important functional variable that not only indicates the amount of energy that enters an ecosystem but also integrates many other attributes (Odum, 1969; McNaughton et al., 1989). We focus our attention on two major sets of questions. First, what is the relative magnitude of these three disturbance effects? We accept they are all important, but are they equally important, or do they differ in this respect? Secondly, are these effects and responses different along a

296

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

Table 11.2 Data on ANPP (above-ground net primary production) of grazed and ungrazed plots of grassland and savanna sites from publications that appeared after Milchunas and Lauenroth’s (1993) analysis 1 Site

MAP 2

Yellowstone National Park (Wyoming, USA)

379

55

85

379

400

590

379

190

340

379

175

295

379

205

235

379

110

130

275

26

26

275

26

27

Shortgrass steppe (Colorado, USA)

Tropical savanna (India)

1 2 3 4

Ungrazed 3

Grazed 4

275

26

23

926

548

794

926

432

626

1145

693

819

1145

703

732

1145

741

876

1145

590

614

Reference Frank and McNaughton (1993)

Hobbs et al. (1996)

Pandey and Singh (1992)

Different values within a study represent different years or different topographic positions. Mean Annual Precipitation (mm). ANPP of ungrazed plots (g m−2 yr−1 ). ANPP of grazed plots (g m−2 yr−1 ).

precipitation gradient that encompasses a wide range of grasslands and savannas? For example, are drier systems more or less sensitive to each of these disturbances than more humid systems? And does the ranking of importance of these disturbances change along the gradient? In order to address these questions, we used two different approaches, one for the effects of grazing and ﬁre, and the other for the effects of climatic ﬂuctuations. To study the effects of grazing and ﬁre, we searched the literature and compiled two data sets from studies in which above-ground net primary productivity had been simultaneously measured in either grazed and ungrazed or burned and unburned conditions. The data set encompassed a wide variety of grasslands (see criteria below), and methodologies, both for treatments and data collection. To study the effects of climate, we used a surrogate for aboveground net primary productivity, the integral of the normalized difference vegetation index (NDVI-I). This index corresponds to the relative difference between red and infra-red reﬂectance of the Earth’s surface as

detected by satellite imagery. It has been shown for many areas that the normalized difference vegetation index is strongly correlated with the above-ground net primary productivity (Paruelo et al., 1997). We looked at how the normalized difference vegetation index ﬂuctuated from year to year in a wide variety of grassland types, and translated that to above-ground net primary productivity based on the relationship mentioned. Grazing and ﬁre We based our analysis on published data (Tables 11.2 and 11.3) complying with the following criteria. We only considered sites that received a mean annual precipitation within the range 200–1200 mm, and were occupied by climatically-determined grasslands or savannas grazed by ungulates (i.e., we excluded cultivated pastures, shrublands, and azonal communities such as wetlands and salt marshes). We only included data for above-ground net primary production measured on an annual basis. Peak biomass data

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

297

Table 11.3 Annual above-ground primary production (ANPP) of burned and unburned plots of different grassland sites along a wide precipitation gradient 1 Site

MAP 2

Unburned 3

Tallgrass prairie (Stillwater, Oklahoma, USA)

810 810

475

470

Mixed grass prairie (South Dakota, USA)

440

118

162

440

249

218

475

Burned 4

Reference

500

Bidwell et al. (1990)

440

249

246

Bunchgrass steppe (Alberta, Canada)

439

118

137

439

118

115

Tallgrass prairie (Illinios, USA)

843

389

648

Tallgrass prairie (Iowa, USA)

782

420

544

Tallgrass prairie (Kansas, USA)

846

170

340

Tallgrass prairie (Nebraska, USA)

681

129

315

681

129

174

681

129

258

472

269

208

472

242

240

Sandhills (Nebraska, USA)

472

111

80

Mixed grass prairie (Nebraska, USA)

472

525

500

Mixed grass prairie (Kansas, USA)

582

238

120

582

238

157

Mixed grass prairie (Oklahoma, USA)

804

132

216

Shortgrass steppe (Kansas, USA)

582

303

71

582

303

215

582

442

167

Annual grassland (California, USA)

529

551

354

Tallgrass prairie (Konza, Kansas, USA)

880

358

367

880

404

506

810

500

530

810

500

580

Bragg (1995)

Briggs and Knapp (1995)

Tallgrass prairie (Konza, Kansas, USA)

880

510

903

Hulbert (1988)

Tallgrass prairie (Illinois, USA)

843

302

1321

Kucera (1981)

843

361

591

Tallgrass prairie (Missouri, USA)

897

509

933

897

482

522

Tallgrass prairie (Iowa, USA)

782

349

750

Tallgrass prairie (Eastern Kansas, USA)

880

186

340

Mixed grass prairie (Western Kansas, USA) Serengeti National Park (Tanzania) Tallgrass prairie (Manhattan, Kansas, USA)

582

380

171

1129

374

491

1129

374

337

880

205

516

McNaughton (1985) Owensby et al. (1970) continued on next page

298

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

Table 11.3, continued Site

MAP 2

Mediterranean grassland (Thessaloniki, Greece)

500

Tallgrass prairie (Kansas, USA) Tallgrass prairie (Illinois, USA)

Unburned 3

Burned 4

Reference

203

176

Papanastasis (1980)

500

203

119

500

203

58

880

473

380

880

180

340

843

395

1397

843

634

756

Tallgrass prairie (Iowa, USA)

782

364

455

Mixed grass prairie (South Dakota, USA)

440

243

200

440

243

234

440

243

237

440

106

116

440

106

88

Texas high plains (Amarillo, Texas, USA)

Juniper community (Callahan County, Texas, USA)

1 2 3 4

440

106

95

542

123

84

542

123

96

542

123

91

542

123

84

542

123

82

542

87

57

542

87

51

542

87

51

542

87

48

542

87

68

600

131

185

600

103

90

782

369

447

Risser et al. (1981)

Steuter (1987)

Trlica Jr and Schuster (1969)

Wink and Wright (1973)

Different values within a study represent different years or different topographic positions. Mean Annual Precipitation (mm). ANPP of unburned plots (g m−2 yr−1 ). ANPP of burned plots (g m−2 yr−1 ).

were considered equivalent to above-ground net primary productivity only for sites with a well-deﬁned growing season (Sala et al., 1988b). Most data on grazing effects were obtained from Milchunas and Lauenroth’s (1993) meta-analysis, whereas the data on ﬁre effects were obtained from reviews by Kucera (1981), Risser et al. (1981) and Bragg (1995), and a literature search. Data for above-ground net primary

productivity were transformed into logarithms for normality. It is widely recognized that grazing alters primary production of grasslands and the herbaceous layer of savannas. However, the direction of the change it provokes at the community level is currently a matter of debate. While there is evidence that grazing may promote productivity (e.g., McNaughton, 1979, 1993),

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

299

in other circumstances grazing has been shown to have detrimental effects on productivity (e.g., Painter and Belsky, 1993; see also Bradbury, Chapter 24, this volume). Figure 11.5 shows that, for a wide range of environments, productivity of grazed plots is linearly related with the productivity of their ungrazed counterparts by:

Fig. 11.5. Relationship between ANPP (above-ground net primary production) of grazed and ungrazed plots of grassland and savanna sites comprising a wide range of primary productivity. The solid line corresponds to the best-ﬁt line, and the dashed line represents the equality line, where ANPP of grazed plots is equal to the ANPP of ungrazed plots. The solid circles belong to the data set of Milchunas and Lauenroth (1993), and the solid squares to papers published after 1990, which were not included in their review.

ln ANPPgrazed = 0.1017 + 0.9475 × ln ANPPungrazed 2 r = 0.84, P < 0.00001, d.f . = 103 , (11.6) where ln ANPPgrazed and ln ANPPungrazed are the natural logarithms of the above-ground net primary production (g m−2 yr−1 ) of grazed and ungrazed plots, respectively. The intercept, not different from 0, the slope, marginally lower than 1 (P < 0.10), and the high r 2 value, all indicate that the effects of grazing on productivity are small, can be either positive or negative, but are more frequently negative. The effects of grazing on productivity were in general no larger than a 50% increase or decrease and showed no relationship with mean annual precipitation (Fig. 11.6). Grazing increased productivity in 28% of the cases and decreased it in 72% of the cases. Productivity of burned plots was linearly related with the productivity of their unburned counterparts by:

ln ANPPburned = −0.63 + 1.118 × ln ANPPunburned 2 r = 0.62, P < 0.00001, d.f . = 64 , (11.7) where ln ANPPburned and ln ANPPunburned are the natural logarithms of the ANPP (g m−2 yr−1 ) of burned and unburned plots respectively (Fig. 11.7). A slope not signiﬁcantly different from one (P > 0.14) and an intercept not signiﬁcantly different from zero (P > 0.28) indicate that ﬁre may have both positive and negative effects on productivity. The effects of ﬁre on productivity were relatively larger than the effects of grazing, and showed a signiﬁcant pattern along the precipitation gradient. The dispersion of the points around the equality line of Fig. 11.7 is larger than the dispersion of the points around the grazed vs. ungrazed line of Fig. 11.5. Thus, both positive and negative effects of ﬁre on productivity appear to be more intense than the effects of grazing. This is clearly shown by Fig. 11.8: ﬁre might increase Fig. 11.6. Effects of grazing on ANPP (above-ground net primary productivity by 300%, or reduce it to less than 20% production) (calculated as (ANPPgrazed − ANPPungrazed )/ANPPungrazed ) of control treatments. The proportional effect of ﬁre of grassland and savanna sites along a gradient of mean annual precipitation. The dashed line indicates the null effects of grazing on productivity was positively associated with mean annual precipitation (r 2 = 0.30, P < 0.00001, d.f. = 62): on ANPP. the positive cases were on the more humid side of the

300

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

gradient and the negative cases on the drier side of the gradient, with a transition between 600–700 mm. Climatic ﬂuctuations

Fig. 11.7. Relationship between ANPP (above-ground net primary production) of burned and unburned plots of grassland and savanna sites comprising a wide range of primary productivity. The full line corresponds to the best-ﬁt line, and the dashed line represents the equality line, where ANPP of burned plots is equal to the ANPP of unburned plots. Solid circles and triangles correspond to sites receiving less and more than 600 mm of mean annual precipitation respectively.

We compiled a data set of the normalized difference vegetation index (a surrogate for above-ground net primary productivity) for 13 grassland sites in North and South America. Site selection was based on the following criteria: a broad gradient of precipitation had to be encompassed, annual precipitation data had to be available, and natural grassland had to be the dominant vegetation type within the scanning unit or pixel. We used the Pathﬁnder Land Program data set (National Oceanic and Atmospheric Administration/National Aeronautic and Space Administration; James and Kalluri, 1994), which covers the 1981–1992 period and is based on maximum composites for tenday periods. Spatial resolution is 8 km. Data for normalized difference vegetation index were transformed into productivity by means of the equation provided by Paruelo et al. (1997). To make this analysis comparable to our previous analyses, we plotted maximum and minimum productivity for the 12-year period as a function of average productivity (Fig. 11.9). The ﬁgure shows that variations in productivity were, in relative terms, much greater on the drier side of the gradient. Both regression lines signiﬁcantly differed from the equality line. The relationship between maximum and average productivity was ln ANPPMAX = 0.84 + 0.90 × ln ANPPAVG 2 r = 0.98, P < 0.00001, d.f . = 12 .

(11.8)

The intercept was signiﬁcantly larger than 0 and the slope signiﬁcantly lower than 1. The relationship between minimum and average productivity was ln ANPPMIN = −2.11 + 1.31 × ln ANPPAVG 2 r = 0.96, P < 0.00001, d.f . = 12 .

(11.9)

It had the opposite pattern: the intercept was signiﬁcantly lower than 0 and the slope signiﬁcantly greater than 1. Thus, relative ﬂuctuations in productivity from year to year tend to be smaller as mean annual rainfall Fig. 11.8. Effects of burning on ANPP (above-ground net primary increases (Fig. 11.10). Extreme productivity values production) (calculated as (ANPPburned − ANPPunburned )/ANPPunburned ) were 80–90% greater or smaller than the mean in dry of grassland and savanna sites along a gradient of mean annual sites, and only 20% in humid sites. precipitation. The dashed line indicates the null effects of burning We investigated to what extent this interannual on ANPP. variability in productivity was related to precipitation

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

301

ﬂuctuations. We analyzed the relationship between the relative variations in productivity shown in Fig. 11.10 and the relative variation of the precipitation of the year in which the maximum or the minimum productivity was recorded [(annual precipitation for maximum or minimum ANPP – mean precipitation)/mean precipitation]. This analysis showed that 50% of the interannual ﬂuctuations in productivity were accounted for by the year’s relative deviation in precipitation with respect to the mean (r 2 = 0.50, P < 0.0001). Discussion of disturbance effects The relatively mild effects of grazing and their uniformity throughout the gradient of rainfall reﬂect the importance of compensatory mechanisms in all sorts of systems and conditions. Compensatory growth, the increase in production per unit of remaining biomass after grazing (McNaughton, 1983b), is responsible not only for the positive effects of grazing on productivity but also for the relatively low magnitude of the negative effects. Since percent consumption increases along the gradient, our results indicate that compensatory growth also increases as precipitation increases. Without this increasing compensatory growth, the effects of grazing should have been all negative and directly related to consumption: more negative from the dry to the wet end of the gradient. Thus, the lack of strong effects on productivity and the common pattern along the gradient are in fact the result of strong feedback processes, which, through drastic changes in species composition, canopy structure, canopy photosynthesis, within-plant resource allocation, nutrient cycling, and water economy, among others, buffer potential changes in the functional, ecosystem-level variable, productivity (McNaughton, 1979, 1983b; Detling, 1987). It is particularly interesting that grazing effects on species composition are strongly inﬂuenced by the position of a system along the precipitation gradient (Milchunas and Lauenroth, 1993); grazing has minor effects on species composition of dry grasslands and large effects in more humid grasslands and savannas. Thus, the more or less similar relative effects of grazing on productivity along the gradient are maintained despite strong structural changes. A challenging, and promising, aspect of our analysis is the difference between ﬁre and grazing effects along the gradient. It has been stated many times that ﬁre effects on productivity vary from predominantly negative

Fig. 11.9. Fluctuation of ANPP in grasslands along a gradient of productivity. Solid squares correspond to maximum ANPP (aboveground net primary production) and open circles correspond to minimum ANPP in a series of 12 years between 1981 and 1992.

Fig. 11.10. Relative variation of ANPP (above-ground net primary production) of grasslands along a gradient of mean annual precipitation. The relative variation was calculated as (ANPPMAX − ANPPAVG)/ANPPAVG for the maximum data and as (ANPPMIN − ANPPAVG)/ANPPAVG for the minimum data. The dashed line is the equality line that indicates the null effects of climate ﬂuctuations on ANPP.

in arid sites to positive in humid sites (Daubenmire, 1968; Vogl, 1974; Anderson, 1982; Bragg, 1995). However, we do not know of a quantitative test of

302

M. OESTERHELD, J. LORETI, M. SEMMARTIN AND J.M. PARUELO

that statement such as the one we have presented. The observation that ﬁre generally does not have negative effects on productivity in humid sites, as grazing does, and does not have positive effects in drier sites, as grazing does, poses interesting questions regarding the different mechanisms through which these two agents of disturbance affect productivity. Interannual variability in productivity has been an important aspect of ecosystem studies because it affects one’s ability to predict productivity. Lauenroth and Sala (1992) have shown that 40% of the variation in productivity of a shortgrass site was explained by annual precipitation. Thus, productivity would ﬂuctuate following the wide relative ﬂuctuations in precipitation that characterize those semiarid environments (Le Hou´erou et al., 1988; Lauenroth and Burke, 1995). Our analysis based on patterns of the normalized difference vegetation index suggests that this pattern of variation closely matches the pattern of the coefﬁcient of variation of rainfall along the gradient. For North American grasslands, the relative variability of the integral of normalized difference vegetation index (our estimator of above-ground net primary productivity) decreased exponentially with an increase in mean annual precipitation (Paruelo and Lauenroth, 1998). Our analysis fails to consider the interaction among disturbances, a potentially important aspect of disturbance phenomena, which has received particular attention recently (Collins, 1987; Hobbs et al., 1991; Briggs and Knapp, 1995; Noy-Meir, 1995). We do not have enough data to study these interactive effects at the large scale we have selected for our chapter. In any case, the patterns we have shown are strong enough, despite any potential interactive effect that we have not accounted for.

A CONCEPTUAL MODEL

Our analyses allowed us to build a conceptual model of the relative effects of three types of disturbance in grasslands and savannas (Fig. 11.11). The central element of this model is that both the regimes and the ranking of importance of the effects of these three types of disturbance depend on the position of a particular system on the gradient of mean annual precipitation. The effect of a disturbance on the productivity of a system located at any point along the gradient will be a function of the disturbance regime at that point, and the

Fig. 11.11. A conceptual model of the relative effects of grazing, ﬁre, and climate on ANPP (above-ground net primary production) along a gradient of precipitation. The lines enclosing the response areas for each disturbance represent approximate boundaries of variation of ANPP according to Figs. 11.6, 11.8 and 11.10. Grazing effects are bounded by thick lines, ﬁre effects are bounded by dashed lines and enclose a dotted area, climate effects are bounded by thin lines and enclose an area with stripes.

way that system responds to disturbances with changes in productivity. Between 200 and 450 mm, herbivory rates as a proportion of consumption and ﬁre frequency are very low. In contrast, the systems are exposed to wide relative climatic ﬂuctuations from year to year that can drastically change water economy, largely on the input side. Grazing in these systems may either increase or decrease productivity, but its effects are relatively mild compared to the interannual variation in productivity. One does not know what the effect of an eventual ﬁre will be, but an extrapolation of our curve suggests that it may be severe. Productivity of these systems will ﬂuctuate greatly around the mean, mainly as a consequence of interannual climatic ﬂuctuations, reaching values ranging from less than one third of the mean to twice the mean. Between 450–700 mm of precipitation, grazing intensity increases, but still is a low proportion of productivity. This, together with slow decomposition rates limited by low soil water and high lignin content of the litter, sets the stage for the occurrence of ﬁre with increasing frequency. Fire depends on fuel accumulation over more than one year. Grazing, as

GRAZING, FIRE, AND CLIMATE EFFECTS ON GRASSLANDS AND SAVANNAS

in any segment of the gradient, may have both positive and negative effects on productivity, negative effects being more frequent than positive. Fire in these systems, however, usually decreases productivity. The relative effects of ﬁre and grazing in this segment of the gradient are of the same magnitude. Year-to-year variations in precipitation are much lower than in the driest end of the gradient, and so is the variation in productivity. It is therefore in this intermediate portion of the gradient that the three agents of disturbance have effects of similar relative magnitude. The presence or absence of grazing, the occurrence of ﬁre, or the occurrence of an unusual year may be equally important in changing productivity. Above 700 mm of rainfall, grazing intensity in systems dominated by large ungulates may be very high, with a more patchy distribution in native systems than in livestock production systems. Fire frequency is much higher, favored by high annual production and the occurrence of a dormant season. Interannual ﬂuctuations in weather are minimal in relative terms. Grazing has minor relative effects on productivity due to compensatory growth, but ﬁre increases productivity up to ﬁve times the mean. Climatic ﬂuctuations, in contrast, can only change productivity by less than 25%. Thus, productivity of these systems will fundamentally depend on ﬁre occurrence in the ﬁrst place, and grazing in the second. Grazing may have a larger effect in these systems through the regulation of the ﬁre regime than by herbage removal per se. Ecologists have known for decades that mean annual productivity of grasslands and savannas is linearly related to mean annual precipitation (Walter, 1939 cited by Rutherford, 1980; McNaughton, 1985; Sala et al., 1988a; McNaughton et al., 1993). Our results provide a quantitative measure of the variation that may be observed around that mean as a consequence of disturbance agents. These results have several implications. First, livestock and wildlife managers can use them when making decisions about setting longterm levels of herbivore populations: the food base will ﬂuctuate in different ways for different systems, and that variation may cascade to affect animal populations and the human societies based upon their production of economic goods. Second, these results can be used to rank the potential importance of each of these agents in driving productivity ﬂuctuations for particular ecosystems along the precipitation gradient. Instead of the qualitative suggestions from ﬁre ecologists, grazing ecologists, climatologists, and grassland ecologists

303

in general stressing the importance of one factor or considering them all equally important, a more balanced, broader view of their relative effects is now available, together with a more precise reference to the heterogeneity of responses across the biome. Finally, the results suggest that grasslands and savannas of intermediate mean precipitation are the most stable in terms of ﬂuctuations in productivity, whereas the two extreme ends of the gradient are more prone to change. However, by controlling ﬁre regime, grazing may reduce ﬂuctuations in productivity at the humid end of the gradient, providing more stability to those systems. This may actually be a byproduct of human utilization of grasslands and savannas with livestock at much higher densities than wildlife (Oesterheld et al., 1992). ACKNOWLEDGEMENTS

We thank D. Frank for providing published and unpublished data, A. Breltr´an and E. Jobb´agy for providing climatic data, and R. Fern´andez Alduncin for sending us the literature we could not locate in Argentina. L. Walker, K. Bradbury and an anonymous reviewer signiﬁcantly improved an earlier version of this manuscript. Our work is supported by CONICET, University of Buenos Aires, and Fundaci´on Antorchas. REFERENCES Anderson, R.C., 1982. An evolutionary model summarizing the roles of ﬁre, climate and grazing animals in the origin and maintenance of grasslands: An end paper. In: J.R. Estes, R.J. Tyrl and J.N. Brunken (Editors), Grasses and Grasslands: Systematics and Ecology. University of Oklahoma Press, Norman, Oklahoma, pp. 297–308. Archer, S., 1989. Have southern Texas savannas been converted to woodlands in recent history? Am. Nat., 134: 545–61. Archer, S., 1995. Tree-grass dynamics in a Prosopis-thornscrub savanna parkland: Reconstructing the past and predicting the future. Ecoscience, 2: 83–99. Archer, S., Scifres, C., Bassham, C.R. and Maggio, R., 1988. Autogenic succession in a subtropical savanna: conversion of grassland to thorn woodland. Ecol. Monogr., 58: 111–127. Belsky, A.J., 1990. Tree grass ratios in East African savannas. A comparison of existing models. J. Biogeogr., 17: 483–490. Bidwell, T.G., Engle, D.M. and Claypool, P.L., 1990. Effects of spring headﬁres and backﬁres on tallgrass prairie. J. Range Manage., 43: 209–212. Biondini, M.E. and Manske, L., 1996. Grazing frequency and ecosystem processes in a northern mixed prairie, USA. Ecol. Appl., 6: 239–256.

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Chapter 12

DISTURBANCE IN DESERTS James A. MacMAHON

INTRODUCTION

In this chapter I discuss disturbances in deserts of the world. On the surface this seems a straightforward task which could be completed by listing natural disturbances and documenting their effects on organisms and then presenting a self-conscious treatment of anthropogenic disturbances. A more complete approach might include consideration of recovery from disturbance, and would give special attention to the effects of disturbance on animals as well as their role as agents of disturbance, rather than discussing only the responses of plants. Indeed, this is the general approach that I take. However, a further understanding of disturbance in deserts requires elaboration of some characteristics of deserts that make the term “disturbance” somewhat difﬁcult to deﬁne rigidly, and that suggest additional topics for this chapter. The ambiguity of the word disturbance has frequently been mentioned in the literature (e.g., Grime, 1979). A seminal deﬁnition was proposed by White and Pickett (1985) when they suggested that “A disturbance is any relatively discrete event in time that disrupts ecosystem, community, or population structure and changes resources, substrate availability, or physical environment.” White and Pickett observe that disturbance is relative to dimensions in space (organism size) and time (organism life-span), and that a common consequence of disturbance is an increase in the patchiness of an area. More recently Huston (1994) elaborated on this conceptual approach when he deﬁned disturbance as “. . . any process or condition external to the natural physiology of living organisms that results in the sudden mortality of biomass in a community on a time scale signiﬁcantly shorter (e.g., several orders of magnitude faster) than that of the accumulation of

the biomass.” His deﬁnition emphasizes the time-scale consideration but not that of size. Superﬁcially, these deﬁnitions suggest that discussion of disturbance in deserts should be limited to consideration of the sudden death of individuals, or of parts of individuals such as leaves of plants, and the subsequent development of increased patchiness. Generally, consideration of the effects of contemporary climate change or other long-interval events would not be included as a disturbance using the perspective developed above. However, some desert plants, at least as clones, may live for 10 000 years (Vasek, 1980); thus, climate change occurring over several decades, the rate that seems to be occurring now, may be a “disturbance” because the environment is changing several orders of magnitude faster than the accumulation of biomass, as indicated by life-span, of some desert plants. Because of this I consider recent, rapid climate change and buildup of carbon dioxide as disturbances and treat them separately. Other workers have adopted a similar perspective, at least for plants (Bazzaz, 1996; Smith et al., 1997). A second problem in deﬁning disturbance is that associated with separating disturbance factors that are inherent to the system and upon which the system depends, as opposed to those that are “foreign” to it (Vogl, 1980). Vogl (1980) discussed a variety of “inherent” disturbance factors such as rain and ﬂoods, wind and storms, ﬁre, snow and frost, erosion, and perturbations caused by animals or man. I will discuss most of these categories for deserts, but generally treat these factors as disturbances of a foreign nature because they occur with unpredictable frequency and amplitude. At the outset, one must constrain one’s view of deserts. The literature contains many deﬁnitions of deserts. Meigs (1953) developed a widely used

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Table 12.1 Arid areas of the world 1,2 Area (total) Africa (30 312) North America (21 322) South America (17 818) Asia (43 770) Australia (7618) Europe (500) World (130 737) 1 2 3

Arid 3

Semi-arid

Total

Arid + semiarid 2

Arid 2

12 819

2951

15 770

52

42.3

1125

1935

3060

14

5.3

1358

1268

2626

14

7.6

10 235

4817

15 052

34

23.4

3250

1375

4625

61

42.7

80

20

32

22.1

0.100 28 889

0.300 12 651

0.400 41 540

Data are reworked from Le Hou´erou (1992a). Data are millions of km2 except the last two columns which are arid or arid + semiarid as a percent of the total continental or world area. Here, “arid” includes extremely arid and hyper-arid lands.

approach for UNESCO that includes rainfall and temperature. This system recognizes hyper-arid, arid, semiarid, and several more mesic subdivisions. Arid and hyper-arid areas receive less than 200 mm (~8 inches) of annual precipitation and semi-arid areas receive between 200 and 600 mm (up to 24 inches). Le Hou´erou (1992a) deﬁned arid lands by the ratio between precipitation and potential evapotranspiration. Using his system, Africa has the greatest absolute area of arid land and is closely tied with Australia as to the proportion of total land area that is arid (Table 12.1). The continent with the smallest proportion of arid land is Europe. The largest single area of desert, the Sahara of North Africa, covers nearly 9×106 km2 , an area approximately the size of the United States. In this chapter, I will discuss disturbances in hyper-arid and arid areas and drier semi-arid areas (transitions), calling them collectively deserts. I will not include semi-arid grasslands [but see Oesterheld et al. (Chapter 11, this volume) and Ghersa and Le´on (Chapter 20, this volume) for coverage of grasslands]. Deserts range from areas devoid of any conspicuous vegetation to areas moderately well vegetated with shrubs and sub-trees, a scattering of grasses, and a variety of annuals and succulents. Deserts represent a highly variable group of ecosystems that occur in areas ranging from temperate to tropical zones around the world. I speciﬁcally will not refer to those cold, highlatitude areas termed polar deserts. Generally, deserts are caused by one of four phenomena. First is the rain-shadow effect whereby moisture is lost from air as it moves inland over mountains. Moisture-laden air condenses to form rain

as it moves up over mountains. As the air crosses the mountains and descends it becomes drier and creates arid conditions. The Mojave Desert of North America is predominantly a rain-shadow desert, as are some deserts of Central Asia. Another source of condensation of water out of the air, and thus an increased drying power of the air, is that caused by air crossing cold ocean currents. Examples of this occur in Africa forming the Namib Desert, along the coast of Peru forming the Atacama Desert, and the peninsula of Baja California forming a portion of the Sonoran Desert. The available moisture in these coastal deserts is often in the form of fog rather than rain. The third possibility is that some land lies in the interior of a continent where moisture-laden air is not common because of the distance from sources of moisture. Some deserts of central Australia and of China are caused by the continental nature of their climates. Finally, there is the effect of high-pressure zones that occur at about 30º North and 30º South latitude that are caused by Hadley Cells. These convection entities are driven by solar energy and the spinning of the earth on its axis. Air lifted at the equator produces rain. As the air moves both north and south of the equator it descends, dries, and creates high-pressure areas that preclude inward movement of moist air. The Sahara and Kalahari, in northern and southern Africa respectively, are deserts of this type. Disruption of any of these climate patterns can be a major disturbance to deserts, altering both their inhabitants and their geomorphology. For introductions to the speciﬁc deserts of the world (Fig. 12.1), their causes, and biological and geological characteristics, the reader is referred to the following

DISTURBANCE IN DESERTS

309

Fig. 12.1. Hot deserts and their transitions. Seasons of precipitation are indicated. This ﬁgure includes most of the arid areas referred to in this chapter except some “cold” deserts of Asia and the United States. Reprinted from Evenari (1985) with permission.

publications: Petrov (1976); West (1983); Evenari et al. (1985, 1986); Walter and Breckle (1986); Allan and Warren (1993); Arritt (1993); Lovegrove (1993); and MacMahon (1999).

NATURAL DISTURBANCE IN DESERTS

Deserts are ecosystems occurring in extreme environments. Because desert organisms are constantly exposed to extreme and often unpredictable values of environmental factors, especially temperature, annual precipitation, and insolation, they are frequently at the limits of their tolerance ranges, and slight changes in the values of these factors can move the organism into zones of their tolerance curves where they cannot survive – that is, mortality occurs or reproduction is not possible. Additionally, because of the open, sparse vegetation, geomorphic processes are changing the landscape at an appreciable rate, often too fast for the vegetation to adjust. Vegetation itself controls some geomorphic processes such as deposition and deﬂation of soils, and some geomorphic instability

may foster diversity (McAuliffe, 1994). Some weather variables act as agents of disturbance, and at the same time they foster other types of disturbance. Examples include: ﬁres, which are particularly likely following periods of above-average rainfall; ﬂoods that occur frequently in response to high-intensity, short-duration rainfall and cause erosion; drought, a phenomenon that occurs periodically in deserts and may destabilize soils; outbreaks of animal pests; extreme swings of temperature due in part to the relative lack of insulating cloud cover; mass movement caused by a variety of geomorphic processes; and, ﬁnally, global climate change, the longer-term climatic changes that can change a particular site from desert status to that of some other community type or to that of a more extreme desert. Given this wide range of types and frequency of disturbances and the long time scale required for desert communities to recover, perhaps 1000 years for some species (McAuliffe, 1988, 1994), it is reasonably argued that desert systems are seldom, if ever, in equilibrium (Sullivan, 1996). This perspective is the basis for a series of models of deserts as “pulse–

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reserve” systems (Noy-Meir, 1973). In a sense, desert systems respond to pulses of favorable rainfall and nutrient availability, and then persist using the reserves developed during the last pulse. The argument that deserts are not equilibrium systems has changed the way that we view their potential management (Westoby et al., 1989; Laycock, 1991). If all of the short-term possibilities for disturbances in deserts are combined with the longer-term prospects of global changes that can increase the variability of climatic factors, then one has to ask whether or not the concept of disturbance, other than those extreme disturbances caused by human beings, is of the same importance in deserts as in more mesic systems. Deserts are so often “disturbed” that disturbance may be one of the deﬁning characteristics of the system rather than a periodic anomaly as in some other systems reviewed in this book. Additionally, the slow recovery of deserts following disturbance, because of the low and variable rainfall, may make them more vulnerable to further disturbance. An example will demonstrate this natural variation. Goldberg and Turner (1986) reported on vegetation changes over a period of 72 years (1906–1978) in a Sonoran Desert site on Tumamoc Hill, near Tucson, Arizona. While virtually no directional change in vegetation was noted, signiﬁcant ﬂuctuations in cover and density coincided with sequences of very wet or very dry years. Establishment of new plants was episodic rather than an annual event. After listing all of the potential changes that can occur in a desert, I have to indicate that there can also be signiﬁcant persistence of system components despite disturbance. Some plant species and even some individual plants can persist for long periods of time. For example, clones of creosotebush (Larrea tridentata) survive for nearly 10 000 years in the same spot (Vasek, 1980), and at least 15 woody plant species in the Sonoran Desert are known to survive over 100 years (Bowers et al., 1995). These longevity data highlight one of the apparent contradictions of desert communities. Despite the rigors of the desert environment, individual plants persist for long periods, and communities may persist for thousands of years (Axelrod, 1979); yet if one looks at the year-to-year variation, for example, the ephemeral nature of the species mix and abundance of annuals, there may be dramatic changes in the composition of individual small plots over very short time scales. All of these factors make it necessary to

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approach the topic of disturbance more carefully than one might in ecosystems where weather and climatic factors may neither vary as much nor reach the same extremes as in deserts. Temperature Desert plants and animals are exposed to extremely high temperatures, and to wide variation in temperatures in a 24-hour period. Prolonged exposure to freezing temperatures is uncommon. As might be expected, desert organisms seem well adapted to extremely high temperatures either by avoidance, a common method used by both plants and animals, or endurance by shutting down photosynthesis, having microphyllous leaves and shedding these during periods of prolonged intense heat, or any one of dozens of other adaptive syndromes. High temperature, per se, probably kills many seedlings but many fewer established plants (Smith et al., 1997). Some succulent species may heat up to 65ºC, for short periods, with no ill effects (Nobel, 1988). The secondary effects of high temperature, as a synergist with low water availability to increase drying power of air or decrease soil moisture content, are often more likely to kill plants and animals than the direct effects. In contrast, low temperatures are known to cause death in a variety of established desert species. One of the best-studied is the saguaro cactus (Carnegiea gigantea; Steenbergh and Lowe, 1977), which succumbs to freezing temperatures. Death from freezing requires both very low temperatures and lengthy periods of exposure. Bowers (1981) analyzed temperature data for Tucson, Arizona. She observed that catastrophic freezes, those that kill desert plants, occurred four times between 1946 and 1979, and that they were more common in the past 100 years than in the previous century. Catastrophic freezes are characterized by minimum temperatures of at least −8.4 to −5.6ºC and durations of at least 15 to 20 hours. Freezing damage to plants during catastrophic freezes has extended hundreds of kilometers south of Arizona to the southern border of the Mexican state of Sonora, where in 1937 temperatures reached −8.9 to −6.7ºC, damaging genera such as Ficus, Pithecellobium, and Randia. Since 1900, some of the genera experiencing damage in Arizona include: Ambrosia, Celtis, Encelia, Jatropha, Olneya, Sapium, a variety of cacti, but especially saguaro, and a variety of subtrees and shrubs (Bowers, 1981). These

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genera include many of the dominants on Sonoran Desert sites. The nurse-plant phenomenon, where young plants cluster under the canopies of other species, may relate to freezing avoidance. Many plants that cluster beneath other plants in their seedling stages may gain the advantage of a warmer microenvironment during cold periods in the cover of the nurse plant than they would if exposed in the open (Nobel, 1988). It is difﬁcult to prove that this is the major advantage to the plant, since the environment under a nurse plant also provides shade, organic matter, the availability of mycorrhizal fungi, protection from herbivory (e.g., McAuliffe 1984a,b), and more favorable water holding capacity because of higher soil organic matter, as well as a variety of nutrients. The nurse-plant phenomenon has been studied extensively (e.g., Nobel, 1988; McAuliffe, 1984a; Yeaton and Manzanares, 1986; Georgiadis, 1989; Cody, 1993). Finally, temperature as an agent of disturbance has been implicated as a controlling factor in the distribution of a life form. Von Willert et al. (1992) extensively reviewed the biology of succulent plants, especially species in the Namib Desert. They suggested that a major limiting factor for this life-history strategy is disturbance in the form of freezing temperatures – that is, those below −4º or −5ºC during the growing season. They suggested that this relationship explains the paucity of succulents in the deserts of Central Asia, the Great Basin of North America, and the Patagonian desert of Argentina. Werger (1983) made similar suggestions earlier. These assertions remain to be rigorously tested. Water Water is probably the environmental factor that most often “drives” a variety of ecosystem processes in deserts (Noy-Meir, 1973, 1974; Crawford and Gosz, 1986). Water and wind share several characteristics. Either can act as agents of erosion or deposition; thus, excesses of either can be damaging, while moderate amounts of both are necessary for the functioning of desert ecosystems. Perhaps the most important effect of water is that it causes the patterning of desert vegetation in time and space (Allen, 1991). Water erosion may be more important in semi-arid regions while wind may dominate as the erosive force in arid areas (West, 1988). In the Patagonian deserts and semi-

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deserts, 28 590 km2 of land have been altered by water erosion (Soriano, 1983). Several aspects of water in deserts, such as availability, predictability, and ﬂooding, will be considered independently. Lack of water can be a signiﬁcant detriment to desert plants although, even after long periods of drought, many systems seem to have the capacity to “spring back” to relatively normal community composition (Goldberg and Turner, 1986). Certainly a few species may be lost during periods of drought, but the majority of characteristic dominants seem either able to survive or to have sufﬁcient seed reserves in the soil to re-establish after drought. This is not unexpected, given the periodic nature of drought in almost all deserts. Plant clumps (resource islands) may increase survivorship of individual plants during drought (Reynolds et al., 1999). Interestingly, clumpage does not matter. There is a great degree of unpredictability in the availability of water in desert ecosystems. Worldwide, there is a correlation between the year-to-year variation in rainfall and the amount of rainfall. As rainfall decreases, its coefﬁcient of variation (and thus unpredictability) increases dramatically (Fig. 12.2). Other calculations for predictability have been suggested (Weis and Schwartz, 1988). This unpredictability, to a large extent, drives desert systems, even inﬂuencing primary production (Le Hou´erou et al., 1988). Nonetheless, some animals (Noy-Meir, 1974) and plants have adapted to the periodic paucity of water. For many plants, an unusual series of back-toback, above-average rainfall years are needed before establishment of a new generation can occur. Single good years of precipitation may cause germination, but these seedlings subsequently die if there is not a second year of above-average rainfall to allow establishment of the root systems. Such episodes, on average, occur with a frequency of about once in forty years (MacMahon and Wagner, 1985) in some areas. This differs markedly from the situation in grasslands and forests, where establishment is nearly an annual phenomenon. Observations of the age structure of populations of barrel cacti (Ferocactus) in the northern Mojave Desert conﬁrm this periodicity (Ehleringer and House, 1984). This episodic establishment has also been noted for a variety of other North American plants, for example ironwood, Olneya tesota (B´urquez and Quintana, 1994) and the agave, Agave deserti (Jordan and Nobel, 1979). During the ﬁrst half of

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Fig. 12.2. Relationship between coefﬁcient of variation (standard deviation/mean annual value) and the mean annual precipitation (mm) for some North American hot-desert sites. Localities are: 1, Yuma, Arizona; 2, Inyo, California; 3, Blythe, California; 4, Mojave, California; 5, Bakersﬁeld, California; 6, Las Vegas, Nevada; 7, Wellton, Arizona; 8, Barstow, California; 9, Lucerne Valley, California; 10, Gila Bend, Arizona; 11, Guaymas, Sonora; 12, Saltillo, Coahuila; 13, La Paz, Baja California Sur; 14, Chihuahua, Chihuahua; 15, Organ Pipe Cactus National Monument, Arizona; 16, Ajo, Arizona; 17, Tucson, Arizona. Reprinted from Evenari (1985) with permission.

this century creosotebush (Larrea) and paloverde (Cercidium) declined with virtually no recruitment of new individuals. Mortality of some species during this period coincided with a prolonged drought from 1936 to 1964. New plants established in an episodic manner following unusually heavy precipitation during certain seasons (Turner, 1990). The re-greening of the Sahel following a long period of drought is another example, in part, of this process (Tucker et al., 1991). Much of the effect of periodic droughts on vegetation and periods of plant establishment can now be observed by the use of satellite imagery (e.g., Nicholson et al., 1990; Peters et al., 1993). This should allow us to detail and quantify this phenomenon at a landscape scale in the future. The speciﬁc season of rainfall often determines the characteristics of rains and even the response of perennials to rainfall (Ehleringer et al., 1991). In the North American deserts, winter rains are of long duration, great areal extent and low intensity.

In contrast, summer rains are characterized by short duration, low areal extent, and high intensity. Many other areas of the world including Asia, Australia, and Africa (Walter and Breckle, 1986), experience these same differences in seasonal rainfall characteristics. Summer rains rapidly saturate the ground surface, causing it to become water-repellent. This leads to surface ﬂow of water, which moves particulate matter, including litter and soil, to stream channels, alters the surface characteristics of the soil in ways that inﬂuence plants and animals, and ultimately causes a feedback to additional runoff events (Fig. 12.3). One interesting feedback is that digging (bioturbation) by animals – for instance, isopods (Hemilepistus reaumuri) and porcupines (Hystrix indica) – in the Negev signiﬁcantly increases erosion (Yair, 1995). Whitford and Kay (1999) review bioturbation in desert soils. A result of surface ﬂow is that ephemeral stream channels often have a richer vegetation, (1) because the water seeps to deeper levels under such storm conditions, and (2) there

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Fig. 12.3. A ﬂow diagram of a conceptual model of the interactions among ecosystem components and rainfall as they create a feedback to desert soil properties in the Negev Desert. From Yair and Shachak (1982) with permission.

is an accumulation of organic matter, which further aids plant growth. Some detailed experiments in the Mojave Desert, in which surface ﬂow was diverted, raise questions about the magnitude of the effects on the soil surface (Schlesinger and Jones, 1984; Schlesinger et al., 1989). These studies suggest that there is little effect on soil properties as a result of artiﬁcially diverting surface ﬂow away from alluvial piedmonts. Although plants grow better where the water has not been diverted, this is interpreted as a result of a differential distribution of biomass of shrubs in natural areas, as compared to the experimental plots. The lack of an effect may be caused by the low energy of winter rainfall in the Mojave Desert. Runoff and sediment yield are often the same on vegetated and denuded plots if the precipitation is of low energy (Bolin and Ward, 1987). Studies of runoff from grasslands and shrublands in the Chihuahuan desert show more N loss in bare areas, less in shrublands, and least in deserts (Schlesinger et al., 1999). Phosphorus loss was small in all habitats. The overland movement of water from high-intensity storms often leads to the formation of ephemeral ponds or lakes in low-lying areas. Such water bodies (playas) are a common feature of desert landscapes where

soils are not porous enough for percolation to prevent accumulation of water on the surface. Playas are generally less than 100 km2 in area, and, although they may look similar, the 50 000 or so that occur worldwide are often geomorphically quite distinct (Neal, 1969, 1975). Playas have been studied in a number of places, and many authors have found signiﬁcant adaptations of animals and plants to playa environments. Changes in the seasonality of rainfall, such as might occur under various scenarios of climate change, could prevent the formation of playas and thus contribute to the loss of a unique ﬂora and fauna. In some areas, high rainfall amounts may simply cause ﬂoods, to which the organisms are not adapted. In such cases, some plants and animals may be disturbed, and in some cases killed or uprooted, but this would generally be an unusual, transient situation. Flooding on most desert surfaces is usually not a phenomenon of any appreciable geomorphic consequence, since the vast bulk of the water is diverted into stream channels, which periodically cut down into the substrate as they accumulate water. Thus, in normal desert systems, if the water can percolate, it adds to the soil water reserves. If rainfall is at high intensity and water ﬂows across the surface, it may remove

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organic matter from the immediate vicinity of plants; the organic matter is deposited elsewhere, and plants on scoured surfaces are already somewhat tolerant of low-nutrient systems. Deﬂation by surface water and wind movement through interplant spaces often causes the development of phytogenic mounds around the bases of plants, especially around clumps of prostrate species (Goudie and Wilkinson, 1977). In contrast, clumps of plants may trap ﬁne materials moved by water or wind, aiding in the formation of lenses of relatively rich soils referred to as “islands of fertility” (Schlesinger et al., 1996). Interestingly, creosotebushes with conical crowns trap less material than those with hemispherical crowns (De Soyza et al., 1997). Such capture processes may actually direct succession in some habitats (Vasek and Lund, 1980). Wind and water erosion in deserts is generally greater on non-vegetated surfaces than on those that are vegetated (Fig. 12.4) (Schimpf and MacMahon, 1981). Vegetation decreases both wind and soil erosion and, to a point, affects the severity of drought (Fig. 12.5). Additionally, the seasonal pattern of rainfall and its energy inﬂuences the shapes of the curves in Fig. 12.4.

Fig. 12.4. A conceptual model of the dependence of erosion by wind (dashed lines) and water (solid lines) on rainfall and vegetation cover. Curve A is relative sediment yield with vegetation cover and B is without. Curve C is relative wind erosion with vegetation cover and D is without. Note that the response axis has no scale, but shows the relative magnitude of the effect based on existing data. From World Meteorological Organization, Geneva, Switzerland (1983) with permission.

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Grime (1979) rightly pointed out that the most difﬁcult environments for plants to adapt to are those with extreme and unpredictable values of environmental variables. Deserts clearly are environments with these characteristics, especially for water availability. Wind Wind can have a variety of effects on deserts that are similar to those of water, while others need to be considered separately. Wind effects are not independent of rainfall or temperature, and they are altered by the degree of vegetation cover (Fig. 12.6). A minor effect of wind is direct physical damage in the uprooting of plants. This is an unusual phenomenon, normally occurring when speeds approach 100 km hr−1 , but it does occur at lower velocities in tall plants such as saguaros, especially if they have been girdled at their bases by the feeding of jackrabbits (Lepus californicus). Additionally, wind may increase the drying power of the air, desiccating plants and animals to critical levels. The aridity of Patagonian deserts and semideserts is due to wind increasing the drying of surfaces rather than just to the low rainfall directly (Soriano, 1983). Finally, perhaps the most important role of wind is as a carrier for particulate matter, both organic and inorganic. In some cases, the movement of materials is a positive inﬂuence and in others it can be quite negative, especially when the material carried is deposited in ways or in quantities that disrupt the functioning of plants or animals. The importance of wind is quite variable from place to place. In North American deserts the effects of wind are generally unimportant. In North Africa, winds are strong and violent, and may blow on at least 50 days in the spring with average speeds of 20–28 km hr−1 ; they can commonly attain speeds of up to 60 km hr−1 , with maxima of over 100 km hr−1 (Grenot, 1974). Dust is regularly removed from deserts and transported elsewhere. Saharan dust is often deposited in southern Europe (Mattsson and Nihlen, 1996). The red rain that occurs in the British Isles is also an example of Saharan dust being deposited in Europe. Similarly, hazes that appear over the Arctic during the summer probably have a source in the Central Asian deserts (Allan and Warren, 1993). It has even been postulated that Saharan dust is a source of nutrients for Amazonian rainforests, where the airborne input of materials may

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Fig. 12.5. Interrelationships among drought severity, plant cover (%) and soil erosion. On barren ground the effects of water and wind erosion and consequently drought severity are increased. From World Meteorological Organization, Geneva, Switzerland (1983) with permission.

Fig. 12.6. A conceptual model of the general relations between annual rainfall and soil erosion on vegetated surfaces for various amounts of rainfall; and the increased erosion caused by wind when vegetation is removed. From World Meteorological Organization, Geneva, Switzerland (1983) with permission.

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reach 26.9 kg−1 yr−1 for phosphate phosphorus, and 12.6 kg−1 yr−1 for potassium (Reichholf, 1986). The effect of wind is especially noticeable in areas of sanddune development – ﬁrst, because the sand dunes are formed as a result of winds, and second, because their subsequent movement is wind-driven. In the Takla Makan Desert of northwestern China, an area covering 320 000 km2 , dunes are moving and expanding the extent of that desert. Some of these dunes reach over 300 m in height. There are two interesting relationships in respect to the effect of wind on vegetation. In areas with virtually no vegetation and extremely low rainfall, movement by wind is not an obvious factor. Presumably, surfaces at these sites have become stabilized, perhaps by the formation of desert pavements, after being exposed to winds for very long times. In contrast, areas with some rainfall seem to generate the greatest amount of particulate matter for movement. Part of this is organic matter. In the Namib Desert where little vegetation occurs on sand dunes, the transport of organic matter by wind and its deposition on dunes provides the organic matter that is the basis for a food chain including detritivorous beetles, especially tenebrionids such as Onomacris spp. (Lovegrove, 1993). Were the wind patterns to change and the source of organic matter to be cut off, it is likely that there would be signiﬁcant negative effects on animals in this dune system, including not only the beetles but also the lizards that feed on them. Clearly, sand dunes represent a specialized desert landform with properties very different from those of less mobile desert substrates. Detailed discussion of sand dunes is beyond the scope of this particular chapter. The reader is referred to Bowers (1982) and Danin (1996) for biological effects of dune environments. For the present purpose, unusual movements of dunes mediated by human activity, for instance by destabilizing existing dunes, are a disturbance. The normal movement of dunes has to be considered as an integral part of the disturbance regime of the dune system, but a foreign disturbance when dune sands cover adjacent non-dune vegetation (Goudie, 1978, 1983). Studies in the Mojave Desert of California determined that wind and the consequent deposition of dust on plants increased leaf temperature and lowered maximum rates of photosynthesis for three species (Atriplex canescens, Hymenoclea salsola, and Larrea tridentata; Shariﬁ et al., 1997). Interestingly,

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at Rock Valley, Nevada, another Mojave Desert site, dust did not appear to alter vegetation composition (Rundel and Gibson, 1996). In some areas of the Mojave Desert the use of off-road vehicles seems to be responsible, in part, for dust storms (Webb and Wilshire, 1983). Plumes from such events may cover more than 1700 km2 and can be seen from space, as was the case in the western Mojave Desert in California on 1 January 1973 (Nakata et al., 1976). Dust storms may cause death to plants. In the Nile Delta, military operations between 1940 and 1943, in which tanks and other military vehicles broke up the undisturbed hard desert surfaces, increased the frequency of dust storms (Smith, 1984). A potential consequence of increases in winds and their related geomorphic effects is that they may inﬂuence a variety of organisms; for instance, many animals with waxy protective layers which prevent desiccation, may have these abraded, causing death through the destruction of the water barrier. Wind-transported dust can cover roads and pipelines, powerlines can be destroyed, and dwellings can be buried. These disturbances may have major economic consequences (Allan and Warren, 1993). Like many other “disturbances” discussed in this chapter, wind is a natural phenomenon in deserts. The geomorphic results of wind are often either erosive or depositional events, occurring over long periods of time. Numerous organisms are well adapted to loose substrates deposited by wind and their subsequent movement. Wind causes geomorphic disturbance only when it occurs in an area not generally exposed to wind, or when an area that had become stabilized by the establishment of natural ecosystems has been destabilized through human inﬂuences or by natural forces. Fire Fire is an unusual phenomenon in undisturbed deserts. If one looks at the aridity gradient from transitional desert grasslands to extreme deserts essentially without vegetation, the frequency of ﬁre decreases dramatically to zero. This is no surprise; it is due to an increasingly meager and dispersed fuel load along the gradient. In altered habitats there are signiﬁcant changes caused by different disturbances. Deserts, in response to unusually high annual rainfall at least two years in succession and/or invasion by alien plants, become more susceptible to ﬁre. In contrast, when transition grasslands are exposed to drought or high grazing

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intensity, ﬁre frequency tends to decrease. These effects are signiﬁcant, because many of the grassland components in the transitions to deserts are highly adapted to ﬁre, and the compositional integrity of these ecosystems may depend on ﬁre, whereas many desert species cannot survive ﬁres. Thus, the grasslands are invaded by desert species and the desert either rebounds (O’Leary and Minnich, 1981) or is dramatically altered (McLaughlin and Bowers, 1982). In the presence of alien annuals that perpetuate a more frequent ﬁre cycle (Rundel and Gibson, 1996), some desert shrub species will simply be displaced. The recovery of desert systems from ﬁre may depend, in part, on the effects of the vegetation before the ﬁre on the soils. For example, in some areas that had supported creosotebush, annuals did not colonize after a ﬁre, whereas interspaces were colonized (Adams et al., 1970). The reason for this phenomenon is not clear. For particular species, ﬁre may be especially damaging. A single ﬁre in the Sonoran Desert may destroy up to 68% of all of the individuals of saguaro (Carnegiea gigantea; Rogers, 1985). Since saguaros require 30 years to reach reproductive maturity (Steenbergh and Lowe, 1977), ﬁres with a return frequency of less than 30 years could essentially remove saguaro from the communities. There may be numerous species that respond similarly but that have not been as well studied. In contrast, creosotebush (Larrea tridentata) may burn to the ground but sprout back in a short time (O’Leary and Minnich, 1981). This sprouting ability is undoubtedly an adaptation to being covered by loose moving soil, but is also a preadaptation to ﬁre. A second example, with additional ramiﬁcations for the effect of ﬁre, involves ironwood (Olneya tesota), generally the tallest and oldest (up to 1200 yr) subtree in the Sonoran Desert. More species of perennials occur under ironwood canopies than in the open. Of 65 species in an area, 52 occur under ironwood, 31 of them only in this situation (B´urquez and Quintana, 1994). If an alien species, buffel grass (Cenchrus ciliaris = Pennisetum ciliare) is established there, it builds up fuel and increases ﬁre probability, and the consequent ﬁres kill ironwood and its associated perennials. The post-ﬁre environment is dominated by buffel grass, and perennials cannot reestablish, the composition of the communities thus being altered (B´urquez and Quintana, 1994). Some of the interesting general characteristics of desert ﬁres have been highlighted for the Sonoran Desert. In a period of 29 years, 1611 ﬁres consumed

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41 000 ha in an area of 391 000 ha of upland Sonoran Desert (Schmid and Rogers, 1988). Fires caused by lightning were fewer in number, but burned twice as great an area as those set by humans, and covered a greater area in desert than in non-desert vegetation (Rogers, 1986). When two consecutive years of aboveaverage winter precipitation occurred, the density of native annuals increased signiﬁcantly. As annuals died and created fuel, the desert, as expected, became more susceptible to burning, and ﬁre frequency increased more than after single wet years (Rogers and Vint, 1987). It has been estimated that, in natural systems in the Sonoran Desert, the return time between ﬁres is about 295 years (Rogers, 1986). This return time would allow many species to re-establish, since average-sized ﬁres would require 276 years to burn every hectare of land, all other things being equal. If the ﬁre return time is shortened by climate change or the invasion by alien annuals, it could be devastating to native communities. Detailed data such as these are less available for natural systems elsewhere. In all areas of the world, however, an increase of alien annual species in deserts generally leads to an increase in ﬁre frequency (Brown and Gubb, 1986). This has been especially well documented in the Great Basin of North America, where the invader Bromus tectorum has increased ﬁre frequency in the Great Basin Desert dominated by sagebrush, and in many areas has virtually eliminated big sagebrush (Artemisia tridentata) (Smith and Nowak, 1990). Invasion of Bromus rubens has had a similar effect in the Mojave Desert (Rundel and Gibson, 1996). It is interesting to note that alien species are not invaders only on sites that have been disturbed by human forces (Knapp, 1996). Many sites that are completely undisturbed have been invaded by these species, leading to the consequent increase in the susceptibility of the system to ﬁre and perhaps its longterm alteration (Brandt and Rickard, 1994). In Australia, ﬁre has demonstrably inﬂuenced aridland vegetation in the pre- and post-Quaternary (Kemp, 1981). Fires have been used by aboriginals for the past 40 000 years, and may have substantially inﬂuenced the ﬂora because of differential effects on ﬁre-susceptible and ﬁre-tolerant species (Williams and Calaby, 1985; Walter and Breckle, 1986). Burned areas are also general features of the contemporary Australian desert environment. In the 1974–75 ﬁre season, 12% of the continent, including much desert, was burned (Cunningham, 1981). In contrast, some places in South Africa, for example the

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Karoo, are generally thought to be little affected by ﬁre as a signiﬁcant factor, despite the fact that limited local ﬁres are caused by lightning (Huntley, 1984). As mentioned above, the inﬂuence of ﬁre in areas transitional to grasslands is generally to reduce the shrub component, since many desert shrub species are susceptible to ﬁre. Occupation of transitional sites by humans has decreased ﬁre frequency, favoring shrub invasion. The relative importance of ﬁre in increasing shrub-dominated lands at the expense of grasslands has been debated for decades. The actual cause of these invasions is likely quite complex. In addition to ﬁre, climate change and grazing have been proposed as the agents of change. Whatever the cause, shrub incursions into grasslands have been demonstrated photographically (Hastings and Turner, 1965; Humphrey, 1974, 1987). Desert animals are often relatively little affected by ﬁre for two reasons. First, aboveground species may be large enough and swift enough to outrun the ﬁre, while other species seek refuge below ground (Polis and Yamashita, 1991), and are insulated by a few centimeters of soil from the effects of the ﬁre. This even includes relatively slow-moving species like desert tortoises (Gopherus agassizii) (pers. observ.). Animals Most of this chapter emphasizes the responses of plants to disturbance regimes. Obviously, as the plant community is altered, so are the associated animal communities. While this is not the place for a detailed analysis of animal communities, I want to mention three kinds of interactions between plants and animals in the context of disturbance. First, animals can positively inﬂuence plants in ways that may help the plant avoid the effects of disturbance; second, animals may act as disturbing agents; and, third, animals themselves may respond to disturbances directly or to alterations in plants caused by disturbances. I will brieﬂy discuss their positive and negative effects on plants and will not discuss direct effects of disturbance on animals here, but include them under the various agents of disturbance where appropriate. Aside from the obvious inﬂuence of animals on processes such as pollination, movement of mycorrhizal spores, and dispersal of seeds (Chambers and MacMahon, 1994), their digging activities may increase the rates of water inﬁltration around the bases of plants by as much as 21% (Laundre, 1993). Digging

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also aerates soils that would otherwise be compacted, and may move substantial amounts of organic matter from above the ground to below it, in essence enriching below-ground soils (Lightfoot and Whitford, 1990). Decreases of some animal populations may act as a disturbance factor for some co-evolved plants, or even for animal species that merely occur in the same community. There are a variety of ways in which animals negatively inﬂuence plants. Obviously, they consume plants or plant parts. However, to qualify as a disturbance, this would have to be in signiﬁcant proportions that are not usually seen. Some cases of consumption that reach the point of being a disturbance include outbreaks of grasshoppers (Locustana and Schistocerca species) and lepidopterous larvae in several African deserts (Werger, 1986) where many plant species are decimated. Fortunately for humans, grasshopper control measures have reduced the effects of these periodic outbreaks. In the past, outbreaks have been very serious and have been recorded for thousands of years. According to Popov et al. (1984, p. 150), Pliny recorded an outbreak in 125 BC that caused death to 800 000 people in Cyrenaica and 300 000 in Tunisia. There are also cases where single species of plants have essentially been removed through animal feeding. Crawford (1991) observed larvae of chrysomelid beetles completely defoliating the evening primrose (Oenothera) plants on New Mexican duneﬁelds. Tarbush (Flourensia cernua) endures up to 30% defoliation by larval Lepidoptera (especially Bucculatrix ﬂourensiae) as well as by a chrysomelid beetle (Zygogramma tortuosa) (Schowalter, 1996). The longterm effects of these defoliations are not known, but simulation of browsing in other systems suggests that there may be signiﬁcant effects (Bilbrough and Richards, 1993). Exclosure studies suggest the possible magnitude of both positive and negative consumer effects on plants. In a 50-year study of exclusion of lagomorphs (Lepus californicus and Sylvilagus audubonii) from creosotebush-dominated communities in New Mexico, there was a 30-fold increase in the basal area of spike dropseed (Sporobolus contractus), and signiﬁcant increases of honey mesquite (Prosopis glandulosa), tarbush, and mariola (Parthenium incanum), in the exclosure plots (Gibbens et al., 1993). Clearly, under conditions of high animal population densities, longterm inﬂuences on shrub and grass populations may be signiﬁcant.

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The numbers of seeds that animals collect may inﬂuence certain highly-favored plant species (Crist and MacMahon, 1992). In a transition zone between the Chihuahuan Desert and desert grassland in southeastern Arizona, exclusion of three species of granivorous kangaroo rats (Dipodomys spp.) caused increases in density of tall perennial and annual grasses, and in populations of rodents characteristic of arid grasslands (Brown and Heske, 1990). The continuation of these studies has suggested that exclusion of birds and rodents increased density of winter and summer annuals, but that the winter annuals were more sensitive (Guo et al., 1995). Finally, these general animal effects may be more complicated than they appear. Individual plants respond differently to animal damage, and the animals respond differently to individual plants of the same species. Individual creosotebushes varied from being seldom and lightly browsed to having 90% of their branches clipped in a single month. In an experiment, jackrabbits browsed more heavily on plants with a history of being browsed than on those which had been only lightly browsed, and less often on artiﬁcially “browsed” (clipped) shrubs than on controls, the latter result suggesting induced resistance to browsing and the former suggesting low constitutive resistance (Ernest, 1994). Most of the examples of the negative effects of plant consumption by animals presented above have related to warm deserts; however, there are similar effects, especially through girdling caused by rodent populations, on cold-desert shrubs such as Artemisia tridentata (Parmenter et al., 1987). There are some indirect effects of animal feeding activities. For some plants, such as saguaro, signiﬁcant amounts of tissue may be removed from individuals by browsing rodents (Neotoma), lagomorphs, and bighorn sheep (Ovis canadensis). In most cases the amount of material removed does not directly kill the cactus, even though it may appear dramatic. However, the individuals become more susceptible to death from freezing, from being blown over, and from fungal infections (Steenbergh and Lowe, 1977). Natural repair Natural ecosystem repair is generally referred to as “succession” (Clements, 1916). There have been varying viewpoints about whether or not succession occurs in deserts. The controversy is often stimulated by the observation that, following a disturbance,

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there does not seem to be a set of plant species replacements leading to a climax, but rather the species forming the “climax vegetation” seem also to form the pioneering vegetation; there may also be several simultaneously co-occurring groups of species early in succession, which include colonizers as well as species characteristic of mature stands. In an experimental study of disturbance in the Negev Desert, Evenari and Gutterman (1976) found that 16 of 20 species typical of mature communities germinated during the ﬁrst year after the disturbance. No “pioneer” ﬂora was obvious. Recently, workers have warned against using the term “succession” because it obfuscates the actual dynamic processes involved in vegetation regeneration (Webb et al., 1987; McAuliffe, 1988). It should be noted that, although McAuliffe (1988) suggested not using the word “succession”, he used the term in a paper in 1991 indicating that the process may occur in deserts, but is not typical, and requires more careful interpretation in arid areas than in more mesic areas where the concept was developed (Goldberg and Turner, 1986; McAuliffe, 1991). Some factors causing deserts to appear to undergo atypical succession include timing of establishment, the biology of individual species, and the particular agent of disturbance, among others. Following a disturbance in an arid area, there may be a long period before establishment occurs, mostly because of the episodic nature of the conditions necessary for establishment (MacMahon, 1981). (See the section on Water, pp. 311–314.) When plants do establish, some species require the presence of others to provide cover for their seedlings, the nurseplant phenomenon discussed above. Under certain circumstances, the very same species may be able to establish in the open. This compounds the problem of deﬁning a typical successional sequence. One observation (McAuliffe, 1988) suggests that, if only presence/absence data are used there are few differences between early and later successional stands in desert areas, paralleling the Evenari and Gutterman (1976) results. For some sites, especially transitional sites, there may be some species that can be predicted to occur early in the process of recovery – pioneerlike species (McLendon and Redente, 1990). For any disturbed site, the actual species list for the community depends on the type and age of disturbance, and whether it was natural or anthropogenic in origin (Webb et al., 1987). Additionally, the type and degree of disturbance affects the trajectory of succession,

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Fig. 12.7. Digraph model showing principal transitions among states in a species rich Sonoran Desert community. This model depicts a realistic set of relations among desert species, in which every state can reach every other state, eventually. This potential outcome, in part, helps to explain the mosaic-like patches of species in deserts and suggests that a single-trajectory for a successional pathway is unlikely. See text for discussion. Reprinted from McAuliffe (1988) with permission.

but not the long-term outcome – that is, a variety of trajectories seem to converge on the same ﬁnal vegetation (McLendon and Redente, 1990). A further complication is that alien species, now abundant in deserts around the world, can slow the establishment of native species by pre-empting a site (Allen and Knight, 1984). In the Canyonlands National Park, Utah, Kleiner (1983) found that arid grasslands subject to grazing may return to a state mimicking their pre-disturbance condition in what appears to be a directional way when grazing is prevented. This return (recovery) included the development of cryptogamic crusts. In other areas where grazing has been studied, the season of grazing had more effect on the trajectory of recovery than the grazing intensity (Whisenant and Wagstaff, 1991). As mentioned previously, the use of nurse-plants by some species, conditioning the early successional environment, is typical in many deserts. For example, in the Karoo, mound-building members of the Mesembryanthemaceae establish early, and the mounds are later invaded by woody species. Such sequences of dependent species look like successional seres. This process is also affected by burrowing of animals (Yeaton and Esler, 1990). One problem with applying the normal concept of succession to deserts is that many of the plants are

extremely long-lived, even persisting for thousands of years. Given this span of time, some plants may persist while the substrates on which they occur are dramatically altered, as are other species in the community. McAuliffe (1991) observed that alluvial terraces of different age supported different plant communities, illustrating the importance of substrates in determining the trajectory of succession. Later, he pointed out that, in contrast to a common assumption, desert soils may indeed be rather well developed and very old, even though they may not appear so if one only examines their surface properties (McAuliffe, 1994). He also asserted that invasion by undesirable species such as the non-native annuals mentioned above is likely, in part, to be a function of the geomorphic surfaces on which the native community occurs. I believe that the process of succession or natural repair does occur in deserts, but (1) the establishment phase may be episodic, (2) the pioneer species may (MacMahon, 1981; Zedler, 1981) or may not (Evenari and Gutterman, 1976; McAuliffe, 1988) be the same as the later successional species, and (3) the geomorphic surfaces may direct the successional outcome as can the type and intensity of the disturbance. Deserts tend to return to their former states, ceteris paribus, even though it may take a long time. For example, there is very slow recovery of vegetation following

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construction of pipelines and roads (Vasek et al., 1975a,b); McAuliffe (1988) estimated that 1000 years may be required for regeneration of the creosotebush community. Carpenter et al. (1986) found that recovery of “old-ﬁelds” in the Mojave Desert took 65–130 years, the time required varying with elevation. McAuliffe (1988) presented a system of the states of spaces on a desert landscape and the transitions among these states, which suggests that these approximate a Markov chain. Markov chain models are based on transitions among the states of a system and the characteristic probabilities of these transitions, which depend only on the current state of the system. McAuliffe started with a simple model of transitions among three states: areas occupied by two plant species (Ambrosia and Larrea) and open space. Later he created a digraph model (Fig. 12.7) of a more complex Arizona Upland desert community. In this instance, as was the case for the simple three-state model, any state can be reached from any other state. The time to reach a particular state depends on the probabilities of the state transitions along the whole state-transition path. This model accurately mimicked data collected by McAuliffe in the Sonoran Desert, and was consistent with the relation between a species and its nurse plant, and other biological relationships. The observation that any state can, with some probability, change to any other state suggests at least one reason that a classic single-trajectory view of succession would not be appropriate in deserts. McAuliffe’s approach provides a possible explanation of the apparent variation in the patterns of recovery of disturbed sites in deserts, which has caused so much confusion over the applicability of the concept of succession to desert communities.

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not likely to directly inﬂuence desert plants, which are already adapted to extremes of temperature, nor the animals associated with them. However, associated with average temperature changes are changes in the extremes of temperature and in world-wide circulation patterns and consequently the patterns of rainfall (Woodward, 1987). Schneider and Root (1996) have warned that there may be a great variety of such “surprises” that are far more important than average temperature changes. One might anticipate that, if both the high and low temperatures become more extreme, they may exceed the tolerance limits of some species, particularly if the duration of freezing temperatures increases by as little as 10–20 hours. One may also anticipate that changes in the patterns of world-wide circulation will change the relative proportions of summer and winter rainfall over deserts. This could cause changes in community composition that are related to a complex set of interactions (Cook and Irwin, 1992), which differ according to whether drought is present or absent (Fig. 12.8) (Post,

Global change Global change involves a variety of world-wide dramatic alterations in the earth’s atmosphere and in ecosystems (Solomon and Shugart, 1993). The four causes of change most often mentioned are: an increase in human population; deforestation; an increase in ambient carbon dioxide levels; and an increase in the earth’s mean annual temperature. I will discuss human population inﬂuences later (pp. 323–325). Here, I conﬁne my comments to climate change and direct effects of an increase in carbon dioxide on deserts, ignoring deforestation because deserts are usually devoid of typical trees. A small increase in average annual temperature is

Fig. 12.8. Feedbacks in terrestrial ecosystem responses to climate change induced by carbon dioxide. Arrows with plus signs (+) indicate processes that have positive effects or increase the rates of other processes. Arrows with minus signs (−) indicate processes that have the opposite effects. GPP represents gross primary production. From Post (1993) with permission.

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Fig. 12.9. Hypothesized response of ecosystems to enhanced carbon dioxide in relation to prevailing nutrient and water availability. Those ecosystems that have been studied in the ﬁeld are marked with heavy outlines; those where aspects of the system have been studied under controlled environmental conditions are shown using broken outlines. The remaining ecosystems are largely unstudied. Note that deserts are generally predicted to be responsive to changes in carbon dioxide, and that different deserts differ in their responses. From Mooney et al. (1991) with permission.

1993). It is predicted for both Australia and the Great Basin of North America that winter precipitation will decrease and summer precipitation increase, conditions to which the vegetation is not adapted. Also, summer precipitation is likely to increase the proportion of grasses and, in turn, the frequency and intensity of ﬁres. Often overlooked are the direct effects of increase in carbon dioxide concentration. These include effects on photosynthetic rates, stomatal conductance, decomposition rates, herbivore consumption, and competition (Bazzaz, 1996). It is postulated that such direct effects occurred in Chihuahuan Desert ecosystems during the last glaciation (Cole and Monger, 1994). These studies indicate a shift on alluvial fans in New Mexico from

domination by C4 grasses to domination by C3 shrubs between 7000 and 9000 years ago, correlated with a rapid increase in carbon dioxide concentration. There is no reason to believe that contemporary systems might not respond in a similar manner, changing the composition of vegetation and a variety of autogenic ecosystem processes including ﬁre frequency (Betancourt, 1996). The assertion about the role of carbon dioxide in shrubland expansion is, however, not without controversy (Archer et al., 1995). Likely direct responses of vegetation to carbon dioxide concentration and to climate change will vary from biome to biome and even within biomes, as indicated for deserts on alluvial or other surfaces (Fig. 12.9). It has been argued by Schlesinger et al.

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(1990) that the boundaries of desert vegetation might be the best places to look for vegetation shifts produced by climate change because of the sensitivity of arid systems to carbon dioxide concentration. To date, these boundary changes are not obvious, although this does not mean that they are not occurring (Schlesinger and Gramenopoulos, 1996). The need for careful, long-term monitoring of changes of vegetation composition in arid transitional areas is clearly indicated.

HUMANS AS AGENTS OF DISTURBANCE

Deserts are home to about 13% (Allan and Warren, 1993) of the world’s human population. Thus, despite their rigorous environments, they are extensively used and inhabited by human beings. Natural disturbances to desert systems can obviously affect many humans and, conversely, human populations can act as agents of disturbance for deserts. In many cases it is difﬁcult to separate natural disturbances from those induced by humans. Historical records of the use of deserts over thousands of years are often reasonably good – for instance, the occupation of deserts by early Egyptian cultures (Zahran and Willis, 1992); but the consequences of those uses are generally not well documented (Werger, 1983). When scientists began studying the nature of deserts, they often viewed landscapes that were already altered. For example, parts of the well-studied Sonoran Desert of Mexico have been occupied by the hunter–gatherer Seri Indians for at least 2000 years with effects that cannot be quantiﬁed (Felger and Moser, 1985). The human use of deserts and of transitional areas, especially those with grasslands, has caused the processes of desertiﬁcation and desertization (Mainguet, 1994). World-wide, more “desert-like” areas are being created at an ever increasing rate. While estimates vary, it is thought that nearly 6×106 ha are subjected to desertiﬁcation each year (Kassas, 1995). The implication is that human mismanagement, in combination with drought conditions in some areas, is turning grasslands into shrublands that have the physiognomy of a desert. There are even cases where the actions of humans, through the removal of plants by domestic livestock, are thought to have changed the albedo of the earth’s surface, leading to an altered rainfall pattern and, consequently, to desertiﬁcation (Charney, 1975, 1977). It is suggested that the removal of vegetation increases surface reﬂectivity, in turn changing the patterns of

Fig. 12.10. Daily precipitation differences that are calculated to occur at various latitudes in the Sahel if vegetation is removed and there are albedo changes of either 14 or 35%. Other percentage changes in albedo would lead to slightly different values; however, these two changes bracket the most likely scenarios. From Charney (1975) with permission.

vertical air movement and consequently amounts of rainfall (Fig. 12.10). I will generally not consider areas that have clearly become deserts through human activities, although in some areas of the world what has caused desert conditions may not be obvious. It is important to add a caveat to my description of the desertiﬁcation/mismanagement cycle. Natural forces can cause “desertiﬁcation”. Changes in climatic patterns can certainly tip the balance between shrublands and grasslands, and have done so in the past in many areas including the Sahara Desert (Le Hou´erou, 1992b), Australia (Smith, 1982), the Negev Desert (Evenari et al., 1971), the Chihuahuan Desert (Dick-Peddie, 1993), and the Karakum (Petrov, 1976). While there does not seem to be any world-wide pattern of change in the total annual rainfall in arid lands, many areas have experienced prolonged declines in precipitation – for example, northern Chile (Burgos et al., 1991),

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North Africa (Le Hou´erou, 1993), and the southwestern United States and adjacent Mexico (Humphrey, 1987). The apparent magnitude of desertiﬁcation can be appreciated by the example of the expansion of the Sahara Desert into the semi-arid Sahel. In the 1960s and 1970s, the Sahara expanded southward by a distance of about 350 km. It is hypothesized that this expansion was caused, in part, by human activities (livestock production, fuel-gathering, population growth), but was signiﬁcantly driven by a drought that was severe from 1968 to 1973, and continued at some level to the late 1980s and early 1990s. Rains returned in the 1980s and some vegetation was restored (Tucker et al., 1991). The return of vegetation caused many to suggest that anthropogenic inﬂuences were, in fact, minor and that this waxing and waning of the Sahara was a regular desert phenomenon. Regardless of the ﬁnal resolution of this debate (Hutchinson, 1996; Rietkerk et al., 1996), a large area extending beyond the Sahel (Ellis and Galvin, 1994), with its dependent inhabitants, was changed from productive grasslands to a lowerproduction desert-like landscape, presumably by a combination of natural and anthropogenic disturbances. Introductions of animals that are managed for human beneﬁt, such as cattle, goats, and sheep, have dramatic direct effects on vegetation, soils, run-off, inﬁltration rates, etc. For example, most of Afghanistan and Iran are so altered by domestic animals that the original vegetation is hard to discern (Breckle, 1983). Introduced species that become feral may also have inﬂuence. For example, non-native honeybees (Apis mellifera) remove at least 90% of the pollen of saguaro cacti (Schmidt and Buchmann, 1986). This dramatic use of pollen by an introduced species may have signiﬁcant effects on normal pollinators of saguaros such as bats, native bees, doves, and moths. The presence of feral burros (Equus asinus) in Death Valley National Park (U.S.A.) has decreased concentrations of native species such as Ambrosia dumosa, Oryzopsis hymenoides, and Sphaeralcea ambigua and some grasses (Loope et al., 1988). These reductions are caused by browsing, grazing, and trampling the native plants. The effect of introduction of rabbits into Australian arid lands is well documented. In addition to altering soil by their burrowing activities, rabbits have displaced native animals, and consumption by rabbits has decreased establishment of some plant species, for instance some acacias (Williams and Calaby, 1985). The introduction of plants such as tamarisk (Tamarix spp.) can have signiﬁcant negative inﬂuences on

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animals and plants. In Death Valley, the presence of tamarisk, a desert riparian species, has lowered the water table, threatening native marsh plants as well as the desert pupﬁsh Cyprinodon nevadensis (Loope et al., 1988). In the Namib Desert of southwestern Africa, introduced species such as Nicotiana glauca (desert tobacco) may be crowding out some native riparian species (Loope et al., 1988). Introduced annuals can alter ﬁre frequency; however, their direct competitive effects are also of concern in some areas. For example, the introduced grass Pennisetum ciliare seems to be displacing the brittle bush (Encelia farinosa) in the Sonoran Desert, at least on some longterm observation plots (Burgess et al., 1991). Such competitive displacement of species is a common result of the introduction of aliens (Williamson, 1996). Tourism effects are highly variable in desert areas; but visitors, anxious to experience deserts and their stark beauty, often trample native vegetation and alter the course of waterways because of rilling effects produced by “path-making”; their footprints and vehicle tracks can disturb microphytic soil crusts, decreasing nitrogenase activity by 30–100% (Belnap, 1996), and have other damaging effects on these somewhat fragile systems (Cunningham, 1981). Offroad vehicles negatively affect ﬂora and fauna of desert areas throughout the world (Seely and Hamilton, 1978); they cause instability of some geological substrates, while compacting others (Webb and Wilshire, 1983). Compacted soils prevent natural system repair, and may take 80–140 years to regain their pre-disturbance physical properties. Interestingly, this process is fastest in colder areas, probably because of the soil-loosening effects of freezing and thawing (Webb et al., 1986). Surprisingly, some deserts recover from human perturbations as extreme as atomic bomb testing in relatively short periods (Yool, 1998). Human repair of disturbed systems It is not possible to cover fully the attempts of humans to repair the results of disturbances in desert ecosystems. Sufﬁce it to say that the unusual characteristics of the biotic and abiotic environment of deserts cause humans to adopt a set of restoration, reclamation, or rehabilitation techniques different from those they would use in more mesic areas. Throughout the arid world, workers have been attempting to divorce themselves from the traditional agricultural approaches to restoration, and to make new sets of

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plans tailored speciﬁcally to desert conditions. These include: dealing with the dispersion pattern of plants during the restoration process to mimic the clumps that provide the nurse-plant effect and accumulate organic matter; attention to the establishment phase of recovery so as to obviate the effects of low and unpredictable rainfall; and other actions tailored to these environments (Majer, 1989; Aronson et al., 1993; MacMahon, 1997). The need to develop these speciﬁc techniques has been well summarized for plants (Allen, 1994), soils and their microbial populations (Kieft, 1991), and animals (Majer, 1989). As a ﬁnal note, because desert plants are so physiologically resilient, once they are established they may actually be used in semi-arid areas to reclaim sites after severe disturbance. Examples of such rehabilitation efforts include the use of cacti (Opuntia spp.), Atriplex species, and other plants in the Mediterranean Basin to control the erosional effects of both water and wind, thus halting the negative effects of opening the vegetation to erosion (Le Hou´erou, 1996a,b). While not all of these past efforts have been successful, there are many desert species that will likely be found to be useful in the future.

CONCLUSION

I have discussed natural disturbance processes in desert systems and, in a much more superﬁcial way, alluded to anthropogenic disturbances. One might ask the question, “Why is disturbance in deserts important?” In natural systems, disturbance is a characteristic of the desert that organisms seem to withstand through survival or regeneration. Anthropogenic disturbances often take deserts beyond their capacity to maintain themselves by introducing new agents of stress or creating values of environmental factors that exceed the limits of organisms. The loss of desert species and the degradation of desert communities is a loss for humankind that is important in the same ways that the more publicized loss of rainforests is important. Many plants and animals in deserts represent untapped genetic resources, act as valuable sources of food, ﬁber and forage, are known to contain many medicinal products (Goodin and Northington, 1985), and provide sites upon which agriculture can be developed if scientiﬁc knowledge is carefully implemented to maintain sustainability of those land surfaces (Hoekstra and Shachak, 1999).

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With a burgeoning human population, all species occurring in deserts, as well as the desert ecosystems themselves, are necessary to provide for the welfare of humans. At the same time, some deserts must be set aside in their natural state, with their natural disturbance regimes, as benchmarks against which overall changes in the global environment and man’s inﬂuence on natural systems can be measured. Economic signiﬁcance of deserts does not encompass the aesthetic attributes of natural desert systems. Many people, including myself, ﬁnd a solace and spirituality in deserts which is unmatched elsewhere and cries out for protection from anthropogenic disturbances.

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329 With Special Reference to the Namib Desert. Cambridge University Press, Cambridge, 340 pp. Walter, H. and Breckle, S.W., 1986. Tropical and Subtropical Zonobiomes. Springer-Verlag, Berlin, 465 pp. Webb, R.H. and Wilshire, H.G. (Editors), 1983. Environmental Effects of Off-Road Vehicles. Springer-Verlag, New York, 534 pp. Webb, R.H., Steiger, J.W. and Wilshire, H.G., 1986. Recovery of compacted soils in Mojave Desert ghost towns. Soil Sci. Soc. Am. J., 50: 1341–1344. Webb, R.H., Steiger, J.W. and Turner, R.M., 1987. Dynamics of Mojave Desert shrub assemblages in the Panamint Mountains, California. Ecology, 68: 478–490. Weis, I.M. and Schwartz, S.S., 1988. The calculation and interpretation of climatic predictabilities. J. Biogeogr., 15: 419–429. Werger, M.J.A., 1983. Vegetation geographical patterns as a key to the past, with emphasis on the dry vegetation types of South Africa. Bothalia, 14: 405–410. Werger, M.J.A., 1986. The Karoo and southern Kalahari. In: M. Evenari, I. Noy-Meir and D.W. Goodall (Editors), Hot Deserts and Arid Shrublands. Ecosystems of the World 12B. Elsevier, Amsterdam, pp. 283–359. West, N.E. (Editor), 1983. Temperate Deserts and Semi-Deserts. Ecosystems of the World 5. Elsevier, Amsterdam, 522 pp. West, N.E., 1988. Spatial pattern–functional interactions in shrubdominated plant communities. In: C.M. McKell (Editor), The Biology and Utilization of Shrubs, 5. Academic Press, Inc., New York, pp. 283–305. Westoby, M., Walker, B. and Noy-Meir, I., 1989. Opportunistic management for rangelands not at equilibrium. J. Range Manage., 42: 266–274. Whisenant, S.G. and Wagstaff, F.J., 1991. Successional trajectories of a grazed salt desert shrubland. Vegetatio, 94: 133–140. White, P.S. and Pickett, S.T.A., 1985. Natural disturbance and patch dynamics: An introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Orlando, Florida, pp. 3–13. Whitford, W.G. and Kay, F.R., 1999. Bioperturbation by mammals in deserts: a review. J. Arid Environ., 41: 203–230. Williams, O.B. and Calaby, J.H., 1985. The hot deserts of Australia. In: M. Evenari, I. Noy-Meir and D.W. Goodall (Editors), Hot Deserts and Arid Shrublands. Ecosystems of the World 12A. Elsevier, Amsterdam, pp. 269–312. Williamson, M., 1996. Biological Invasions. Chapman and Hall, London, 244 pp. Woodward, F.I., 1987. Climate and Plant Distribution. Cambridge University Press, Cambridge, 174 pp. World Meteorological Organization, 1983. Meteorological Aspects of Certain Processes Affecting Soil Degradation – Especially Erosion. Tech Note Number 178, World Meteorological Organization Number 591:149, Geneva, Switzerland. Yair, A., 1995. Short and long term effects of bioturbation on soil erosion, water resources and soil development in an arid environment. Geomorphology, 13: 87–99. Yair, A. and Shachak, M., 1982. A case-study of energy, water and soil ﬂow-chains in an arid ecosystem. Oecologia, 54: 389–397. Yeaton, R.I. and Esler, K.J., 1990. The dynamics of a succulent karoo vegetation. Vegetatio, 88: 103–113. Yeaton, R.I. and Manzanares, A.R., 1986. Organization of

330 vegetation mosiacs in the Acacia schaffneri–Opuntia streptacantha association, southern Chihuahuan Desert, Mexico. J. Ecol., 74: 211–217. Yool, S.R., 1998. Multi-scale analysis of disturbance regimes in the northern Chihuahuan Desert. J. Arid Environ., 40: 467–483. Zahran, M.A. and Willis, A.J., 1992. The Vegetation of Egypt. Chapman and Hall, London, 424 pp.

James A. MacMAHON Zedler, P.H., 1981. Vegetation change in chaparral and desert communities in San Diego County, California. In: D.C. West, H.H. Shugart and D.B. Botkin (Editors), Forest Succession. Springer-Verlag, New York, pp. 406–430.

Chapter 13

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS Karen L. McKEE and Andrew H. BALDWIN

INTRODUCTION

The term “wetland” refers to land that is either periodically saturated with or permanently covered by water, exhibits unique soil conditions, and contains hydrophytes (plants adapted to periodic ﬂooding). Ecosystems or communities that can be considered to be associated with wetlands occur in all climatic regions, can exist under widely variable hydrogeologic conditions, and contain very different species assemblages (Whigham et al., 1993). Their diversity and transitional position in the landscape have led to difﬁculties in classiﬁcation of wetland types, confusion regarding which areas are in fact wetlands, and problems in determining the areal extent of wetlands. Wetland classiﬁcation systems are usually based on some combination of vegetation, hydrology, and waterquality features [see, for instance, Cowardin et al. (1979) for wetlands in the United States], but no one system adequately classiﬁes all wetland types worldwide. Wetlands may be tidal or non-tidal, coastal or inland, freshwater or saline. They may be dominated by trees, shrubs, herbaceous macrophytes, or mosses. Wetlands may be further categorized according to soil type (e.g., mineral or peat) or water quality (e.g., acidic or basic, nutrient-rich or nutrient-deﬁcient). The variability among wetlands of different regions and their importance to humans throughout history have resulted in many different terms describing wetlands, including the familiar marshes, swamps, mires, fens, and bogs. Other terms (e.g., mangroves, pocosins, salinas, savannas, and playas) may be applied to speciﬁc wetland types in certain regions (Chapman, 1977; Larsen, 1982; Gore, 1983; Whigham et al., 1993). For the purposes of this chapter, we have provided a simpliﬁed classiﬁcation scheme that will deﬁne some common types for the reader unfamiliar

with wetlands and to guide the more knowledgeable reader (Table 13.1). Because wetlands develop in the zone between dry land and open water, they possess qualities transitional between terrestrial and aquatic systems. Wetlands may play a biogeochemical role as source, sink, or transformer of chemicals due to their intermediate position between terrestrial (source) and aquatic (sink) systems. Wetlands also generally exhibit a high primary production compared to upland or aquatic systems. The hydrology of wetlands, which varies from intermittently to permanently ﬂooded, is a major determinant of ecosystem structure and function (Mitsch and Gosselink, 1993). Hydrology determines their unique physico-chemical attributes; controls the movement of water, sediment, nutrients, and toxins through them, and inﬂuences the occurrence and distribution of plants and animals within them. Hydrology controls abiotic conditions such as soil moisture, oxygen, and nutrient availability, and inﬂuences biotic processes such as dispersal of diaspores. Coastal wetlands are additionally affected by ocean tides, which generate ﬂuctuations in water level and/or salinity. These factors in turn affect the distribution and relative abundance of plant and animal species, and ecosystem functions such as productivity, energy ﬂow, and nutrient cycling. Wetland organisms are varied in terms of size, form, complexity, and mobility. They may be planktonic (ﬂoating), natant (swimming), benthic (living in or on the soil substrate), rooted, or epiphytic (attached to macrophytes). Plants and animals range in size from microscopic to massive. The microbiotic community includes bacteria, protozoans, algae, and fungi. Creeks and channels and the water column overlying the wetland surface contain phytoplankton (dinoﬂagellates, diatoms, and blue-green algae), which are important primary producers, and zooplankton, which include

331

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Karen L. McKEE and Andrew H. BALDWIN

Table 13.1 Common wetland types covered in this chapter and some of their distinguishing characteristics Wetland type

Synonymous terms

Marsh

Dominant vegetation

Hydrology

herbaceous (grasses, rushes, sedges, forbs)

Water chemistry 1 /soil type

generally non-acidic, mineral or organic soil

Salt marsh

tidal or non-tidal

>18‰ salinity 1

Brackish marsh

same

0.5–18‰ salinity

Freshwater marsh

same

<0.5‰ salinity

trees, shrubs

tidal or non-tidal

<0.5‰ salinity, generally non-acidic, mineral or organic soil

mangal, tidal forest, swamp

trees, shrubs, palms

tidal

>0.5‰ salinity, mineral or organic soil

mire, peatland, moor, muskeg, fen, pocosin

mosses, grasses, sedges, rushes

non-tidal

often acidic, low nutrient, peat-forming, rain-fed, <0.5‰ salinity

Forested wetland

woody

Swamp 2

Mangrove Bog

1 ‰, parts per thousand by weight of dissolved salts in solution. Salinity regimes in brackish marshes are sometimes further categorized as: oligohaline, 0.5–5‰; mesohaline, 5–18‰. 2 Some herbaceous-dominated wetlands are occasionally referred to as reedswamps (e.g., those dominated by the genera Phragmites or Typha), but to avoid confusion, we include these in the marsh category.

protozoans and many juvenile invertebrates and ﬁsh. The meiofauna include organisms such as nematodes, copepods, amphipods, polychaetes, oligochaetes, and ostracods. Macroalgae may occur as ﬂoating mats, or attached to the soil or to macrophytes. Rooted macrophytes include grasses, sedges, rushes, forbs, vines, shrubs, and trees. Macrofauna include molluscs, arthropods, ﬁsh, amphibians, reptiles, birds, and mammals. Forested wetlands may contain huge trees (e.g., mangroves and bald cypress, Taxodium distichum) more than 30 m tall. Some wetlands (e.g., mangroves) are inhabited by large herbivores such as manatees and large predators such as crocodiles or alligators. Humans must also be included in this list, since they often inhabit, modify, or exploit wetlands. Wetlands are dynamic ecosystems characterized by an ever-shifting mosaic of physico-chemical and biotic factors. Across any wetland landscape there is variation in physical, chemical, and biological features, even though these conditions may superﬁcially appear to be relatively uniform. Even small differences in factors such as topography, plant canopy structure, or nutrient concentrations generate a variety of opportunities for plant and animal recruitment, growth, and reproduction. This spatial heterogeneity can be found at any scale of resolution, from microns to kilometers, and

thereby may affect a wide variety of wetland life forms. Also, because environmental conditions change over time, mosaic patterns in wetlands are dynamic. Relative abundances of species and ecosystem functioning consequently vary over time as a wetland develops and matures. The boundary between wetland and nonwetland is also constantly shifting, although such changes may only be obvious on time scales of centuries to millennia. Boundaries of coastal wetlands such as salt marshes are often linked to duration, frequency, and amplitude of tides (McKee and Patrick, 1988), but may slowly shift in response to sea-level changes (Reed and Cahoon, 1992). The literature on wetlands consists of thousands of citations, many of which contain information about disturbance regimes as well as responses of the biota to various types of disturbance. However, few studies have explicitly examined disturbance, particularly by natural agents, as a force shaping the structure and function of wetlands. General descriptions of patterns and processes in wetlands typically lack a discussion of disturbance, except in the context of plant succession (e.g., van der Valk and Davis, 1976, 1978; Larsen, 1982; Kantrud et al., 1989) or human-related impacts leading to wetland loss (e.g., Johnson, 1985; Mitsch and Gosselink, 1993; Viles and Spencer, 1995). This

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

situation is in contrast to the extensive literature on disturbance in other natural communities (see Sousa, 1984; Pickett and White, 1985a) and its role in spatial patterning and succession of communities in marine (Connell, 1978; Lubchenco, 1978; Sousa, 1979, 1985; Paine and Levin, 1981; Connell and Keough, 1985) and terrestrial (Connell, 1978; Brokaw, 1985; Denslow, 1985) habitats. Since disturbance is a major source of temporal and spatial heterogeneity, as well as an agent of natural selection in the evolution of life-history characteristics (Sousa, 1984), its potential role in wetland ecosystems is obvious. The interrelationship between disturbance regimes and the differential expression of inherent attributes of species can clearly generate patterns in wetland communities that differ from those in the absence of disturbance. Some examples include maintenance of pine–wiregrass (Pinus–Aristida stricta) savannas by ﬁre (Walker and Peet, 1983), cyclic succession in glacial prairie wetlands caused by drought–ﬂooding cycles (van der Valk and Davis, 1976, 1978), hurricane impacts on mangrove forests (Steinke and Ward, 1989; Roth, 1992; Smith et al., 1994; Baldwin et al., 1995; Doyle et al., 1995), patch creation by wrack deposition in salt-marsh communities (Bertness and Ellison, 1987; Ellison, 1987; Bertness, 1991a,b; Bertness et al., 1992) and changes in community structure of coastal marshes by grazing animals (Jensen, 1985; Kerbes et al., 1990; Taylor and Grace, 1995; Baldwin, 1996; Ford, 1996; Grace and Ford, 1996) or of peatlands by ﬁre (Maltby et al., 1990; Legg et al., 1992; Motzkin et al., 1993; Kirkman and Sharitz, 1994). Disturbance not only impacts the relative abundances of populations, but also inﬂuences community and ecosystem processes such as succession, primary and secondary production, energy ﬂow, and nutrient cycling. Fire, for example, can release a pulse of nutrients from combustion of organic matter and create openings in the vegetation canopy, allowing invasion of species requiring high light and nutrients. Wind-storms and grazing by geese or mammals also open up space in vegetation and cause redistribution of organic matter and nutrients. Many ecosystems have been altered by anthropogenic activities, which can cause direct destruction of biomass, alter hydrology, or introduce pollutants and exotic species. Disturbance is thus widely recognized by ecologists as a fundamental process shaping many types of ecosystems, including wetlands. Our speciﬁc aim in this chapter is to provide an overview of disturbance in wetlands, with an emphasis

333

on North America, since complete coverage of disturbance regimes and responses by the biota in all wetlands is impossible here. Instead, our goal is to examine some general concepts about disturbance regimes using a few speciﬁc examples and to consider their role in wetland structure and function. Because this book deals with ecosystems of disturbed ground, we are limiting our discussion to wetlands with emergent herbaceous or woody vegetation, omitting permanently-inundated aquatic areas containing submergent vegetation or lacking vegetation entirely. Although we review studies reporting disturbance effects or responses in all the major wetland types listed in Table 13.1, about 40% are salt-marsh studies and the remaining 60% are divided fairly evenly among other marsh types, swamps, and bogs. A major focus of our chapter is on natural sources of disturbance in wetlands, but we also include information about anthropogenic types. We also emphasize effects of disturbance on the emergent plant communities in wetlands, but animal responses to disturbance, as well as their role as disturbance agents, are discussed. Although many of the concepts covered in our discussion are applicable to all wetlands, speciﬁc responses may vary greatly from region to region on account of differences in species assemblages and disturbance regimes. In this section, we have presented some basic characteristics of wetlands to provide a foundation for an understanding of the role of disturbance in wetlands. The following sections describe the history of disturbance in wetlands, sources of variation in disturbance, and responses to disturbance at different levels of organization.

HISTORY OF DISTURBANCE IN WETLANDS

Natural disturbance is important in the creation as well as destruction of wetlands. For example, northern peatlands such as the prairie pothole region in North America were created by glaciation or the melting of glacial ice (Larsen, 1982; Kantrud et al., 1989). During the Pleistocene, great ice sheets transported huge amounts of rock and soil, scoured depressions in the landscape, and deposited large ice blocks which melted after the glacier retreated. The resultant glacial topography, combined with variations in surface and ground-water hydrology, has generated the development of numerous basin wetlands. The origin of the Carolina bays, a type of shrub-bog occurring

334

in the Carolinas and Georgia, has been attributed to depressions created by an ancient meteor shower, although other hypotheses (e.g., wind and wave action) have challenged this notion (see Sharitz and Gibbons, 1982). Riparian wetlands are created by streams and rivers overﬂowing their banks and by channel meandering (Dollar et al., 1992; Shankman, 1993). Animals such as beavers (Castor canadensis) are well known for creating wetland habitat by building dams on streams and allowing water to spread across large expanses of land (Beard, 1953). Many coastal wetlands in North America developed when the sea level rose after the last glacial period and inundated river valleys and shallow continental shelves (Mitsch and Gosselink, 1993). Extensive wetlands have also formed in river deltas and along coasts with large sediment supplies – for instance, mangrove swamps in the Bay of Bengal in India and Bangladesh (Blasco, 1977) and estuaries of northern Australia (Woodroffe, 1990), and freshwater wetlands in the Mississippi delta (Templet and Meyer-Arendt, 1988). Sea-level rise and fall over geologic time is also an important factor controlling development and demise of wetlands (Reed, 1990; Woodroffe, 1990). Large-scale disturbances such as hurricanes, volcanic eruptions, earthquakes, landslides, and tsunamis can completely destroy large tracts of coastal wetlands (Smith et al., 1994; Viles and Spencer, 1995). Once the geomorphic setting is created for wetland development, disturbance regimes, such as ﬁre, grazing, and wind-storms operate to modify plant and animal communities over time. These processes maintain temporal and spatial heterogeneity in wetlands where physical and chemical conditions might otherwise be relatively uniform. The resulting assemblages reﬂect the heterogeneity induced by disturbance, both on a historical scale (i.e., through responses of individual organisms to variation in environmental conditions) and on an evolutionary scale (i.e., through evolution of life-history traits adapted to a particular disturbance regime: Sousa, 1984). Thus, on a geological time scale, natural disturbance has been a major factor determining the type and extent of wetlands as well as controlling structure and dynamics within any speciﬁc wetland. The effect of human-induced disturbance on wetlands is more recent, occurring on a historical timescale, but is no less important. Humans from many cultures have lived in, managed, and altered wetlands since early civilization (Maltby, 1986; Mitsch and Gosselink, 1993). Some human cultures have lived in

Karen L. McKEE and Andrew H. BALDWIN

harmony with wetlands for hundreds, even thousands, of years. People have hunted, ﬁshed, and harvested plant material for food, fuel, fodder, and shelter in wetlands. Some cultures have lived and ﬂourished within wetlands – for instance, the American Cajuns in Louisiana swamps, or the Marsh Arabs of southern Iraq (Mitsch and Gosselink, 1993). Rice and ﬁsh cultivation in domestic wetlands has provided food for humans in many countries such as China, the Philippines, and Thailand for centuries. Cranberry (Vaccinium macrocarpon) is cultivated and harvested from bogs and marshes in the northern United States (Larsen, 1982; Jorgensen and Nauman, 1994), and crayﬁsh are cultivated in swamps and rice ﬁelds in southern Louisiana (Kniffen and Hilliard, 1988). Although the foregoing activities certainly have an impact on wetlands, they do not necessarily lead to wetland loss, as can some management practices. Wetlands are managed to control ﬂooding, erosion, or insects such as mosquitos, to maintain water quality, to create buffers between urban and industrial areas, to provide habitat for ﬁsh and wildlife, to produce food, ﬁber, and fodder, to receive air- and waterborne pollutants, and to provide places for recreation, research, and education (Maltby, 1986; Mitsch and Gosselink, 1993). Some of these management practices can have dramatic impacts on the structure and function of wetlands, even causing their complete destruction. The major causes of wetland loss are conversion to agriculture, aquaculture, or mariculture and for urban or industrial expansion (Walsh, 1977; Maltby, 1986; Viles and Spencer, 1995). Even small marshes such as those in the prairie pothole region (Kantrud et al., 1989) and playas in eastern Texas (Bolen et al., 1989) have been drained and cultivated. Thus, from an historical perspective, disturbance has played an important role in creating and shaping wetland structure, ecological attributes, and values to humans. In the next section, we will review disturbance regimes in wetlands and discuss sources of variation that determine impacts on the biota.

SOURCES OF VARIATION IN DISTURBANCE

In any particular ecosystem, the sources of variation in disturbance include differences in: 1) kinds of disturbance, 2) disturbance regimes, and 3) ecosystem scale (White and Pickett, 1985). These factors determine the effect of disturbance and response of the system

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

335

Table 13.2 Agents of disturbance in wetlands and examples of natural and anthropogenic sources Type

Natural

Anthropogenic

Flood

storm surge

structural impoundment

Drought

climatic ﬂuctuation

water withdrawal

Temperature extreme

frost

thermal pollution

Burial

sediment or wrack

dredge spoil disposal

Breakage/abrasion

windthrow

logging, hay harvesting

Soil erosion

wave, tide or ice scour

canal dredging, peat mining

Fire

lightning

prescribed burning

Toxins

saltwater intrusion

oil, pesticides, heavy metals

insect, waterfowl, or mammal herbivory

domestic livestock

Bioturbation

burrowing, nest-building, animal trails

foot trafﬁc, off-road vehicles

Predation

feeding by invertebrates, birds, mammals

hunting, ﬁshing

Physical/chemical

Biotic Grazing

to the event. In the next sub-section, we review the major types of disturbance in wetlands and give some examples of both natural and anthropogenic types. The succeeding sub-sections describe disturbance regimes from a wetland perspective, and effects and responses of wetlands at different levels of organization (i.e., species, population, community, ecosystem, and landscape). Kinds of disturbance in wetlands Because they occupy a special position in the landscape, wetlands experience disturbance types that encompass those typical of both terrestrial (e.g., ﬁre or wind) and aquatic (e.g., wave action or sedimentation) ecosystems. North American wetlands occur in all the major climatic regions and are consequently exposed to disturbances ranging from breakage or abrasion from ice in boreal and temperate regions to hurricanes in tropical areas. The occurrence of wetlands along coastlines as well as in the interior of land masses means that they are also exposed to a wide range of factors that may interact with disturbance events. Salinity, for example, might interact with wind damage from a hurricane to further stress or kill organisms in a freshwater-swamp community (Conner, 1995). In Arctic salt marshes, feeding by lesser snow geese (Anser caerulescens) exposes the sediment surface, and the increased evaporation rate leads to high soil

salinities that inhibit growth of the remaining grazed plants (Srivastava and Jefferies, 1996). We have divided disturbance agents in wetlands into two major categories: natural and anthropogenic. Natural disturbances in wetlands include water-level extremes, frost, burial, erosion, ﬁre, breakage or abrasion due to wind, waves or ice, saltwater intrusion, grazing, and bioturbation. Human activities such as water-level management, logging, peat harvesting, dredging, hunting and ﬁshing, and introduction of industrial and agricultural pollutants can damage or kill wetland plants and animals, although these often differ from natural disturbances in terms of intensity and potential for organisms to adapt. Examples of wetland disturbance agents are listed in Table 13.2, and we brieﬂy review them below. Natural disturbance Although wetlands are naturally characterized by water-level ﬂuctuations, meteorological events can increase the depth and duration of ﬂooding well beyond normal diurnal or seasonal ﬂuctuation, and can increase exposure of plants and animals to other disturbances such as wave action. Fluctuations in rainfall or groundwater may lead to periods of drought in some wetland systems, such as the Everglades of Florida (Duever et al., 1994) or boreal wetlands (Hogenbirk and Wein, 1991). Because of the central role of hydrology in wetland structure and function, a change in inundation regime can have far-reaching

336

effects (Mitsch and Gosselink, 1993). Water-level changes, for example, can induce plant succession in freshwater marshes (van der Valk and Davis, 1976, 1978). Wetland animals intolerant of submergence may be drowned or displaced by unpredictable increases in water level, for instance during storms (Daiber 1982; Kinler et al., 1990). Drought conditions, on the other hand, may concentrate animals in and around drying pools of water (Duever et al., 1994). Temperature extremes, particularly frost, can kill vulnerable plants and animals in some wetlands, such as the Everglades (Duever et al., 1994). Wetlands occurring in sub-tropical and some tropical regions (e.g., southern Florida) are periodically subjected to cold or freezing temperatures, which affect local and regional distributions of species as well as structural characteristics of the vegetation. Mangroves and other tropical trees, for example, are often damaged or killed by frost at their latitudinal limits (Sherrod and McMillan, 1985; Myers, 1986). Even where these species persist, repeated freezes prune the aboveground biomass to produce stunted, multi-stemmed trees. Wetland plants and sessile animals may become buried by sediment deposition, and natural events such as winter storms and hurricanes can deposit huge amounts of sediment into a wetland in a relatively short period of time. Hurricane Andrew, for example, deposited sediment into a Louisiana estuarine marsh that was 4–11 times greater than the long-term (30 yr) annual rate (Nyman et al., 1995). Wetland areas may also be covered by accumulations of dead plant shoots called wrack, a common natural disturbance in salt marshes causing damage varying from total elimination of vegetation to decreased density of plant stems underneath the mat (Bertness and Ellison, 1987; Hartman, 1988; Valiela and Rietsma, 1995). In Florida, winter dieback of freshwater marsh plants such as arrowhead (Sagittaria latifolia), cattail (Typha spp.), maidencane (Panicum hemitomon), pickerelweed (Pontederia cordata) and spadderdock (Nuphar luteum) can also create large rafts of dead shoots (Kushlan, 1991). Deposition of wrack or sediment thus generates heterogeneity in some wetland landscapes and creates bare patches for invasion by other plant species (Ranwell, 1961; Bertness and Ellison, 1987; Hartman, 1988; Kushlan, 1991; Guntenspergen et al., 1995). Physical agents such as wind, waves, and ice may cause breakage or abrasion of plant shoots and disruption of soil structure in wetlands. Wind damage

Karen L. McKEE and Andrew H. BALDWIN

during storms or hurricanes may defoliate, break branches and boles, or uproot trees in freshwater or mangrove swamps (Putz and Sharitz, 1991; Roth, 1992; Loope et al., 1994; Smith et al., 1994; Doyle et al., 1995; Baldwin et al., 1995; Conner, 1995). In cold regions, ice rafting, ice gouging, and ice melt are important disturbance processes (McCann and Dale, 1986; Dionne, 1989; Swanson and Rothwell, 1989). Erosion of soil by ice, wind or water can expose wetland plant roots, displace animals, and alter spawning habitats (Garofalo, 1980; Letzsch and Frey, 1980; Nyman et al., 1994). Fire is a frequent disturbance in many wetlands, and is particularly important in those habitats characterized by dry seasons and accumulations of litter and/or peat (Ewel and Mitsch, 1978; Duever et al., 1994; Nyman and Chabreck, 1995). Fire has been recognized as an important agent in the development and maintenance of many wetland communities (Lynch, 1941; Ewel and Mitsch, 1978; Tallis, 1987; Maltby et al., 1990; Duever et al., 1994; Nyman and Chabreck, 1995) and for wildlife management (Cartwright, 1942; Lay and O’Neil, 1942; Nyman and Chabreck, 1995). Fire has been studied in different types of wetlands, including coastal marshes, cypress swamps and bottomland hardwood forests, freshwater marshes, pocosins, and bogs (see Kirby et al., 1988; Nyman and Chabreck, 1995). A number of studies have been conducted in the Everglades, the Okefenokee Swamp in Georgia (U.S.A.), the Great Dismal Swamp in Virginia and North Carolina, and the Delta Marsh (Kirby et al., 1988; Duever et al., 1994). Frequent ﬁre limits invasion of woody vegetation into freshwater marshes and bogs (Vogl, 1973; Ewel, 1986), affects species composition (Vogl, 1973; Chabreck, 1976; Hogenbirk and Wein, 1991; Timmins, 1992), and retards peat accumulation (Tallis, 1987; Maltby et al., 1990; Nyman and Chabreck, 1995). Fires in swamps and marshes are caused naturally by lightning (Gunderson and Snyder, 1994). In mangrove forests, a lightning strike may kill small patches of trees even if it does not initiate a ﬁre (Smith et al., 1994). Biotic disturbances such as grazing and bioturbation are also important in wetlands (Table 13.2). Grazers in wetlands include insects (Scott and Haskins, 1987; Daehler and Strong, 1995; Feller, 1995), crustaceans (Lodge, 1991), snails (Pace et al., 1979; Bertness, 1984), ﬁsh (Lodge, 1991), waterfowl (Smith and Odum, 1981; Smith, 1982; Cargill and Jefferies, 1984; Bazely and Jefferies, 1986), and mammals (Gillham,

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

1955; Wilsey et al., 1991). Although consumption of plant tissue may vary spatially and temporally, grazers can reduce macrophyte biomass (root or shoot) up to 100%. Large, unvegetated patches called “eat-outs” can be created in marshes by snow geese and nutria (Myocaster coypus) (Lynch et al., 1947; Daiber, 1982). In addition to direct removal of plant biomass by grazing, animals such as worms, molluscs, crustaceans, ﬁsh, alligators, geese, muskrats (Ondatra zibethicus), beavers, seals, and elephants may cause disturbance in wetlands through activities such as grubbing (Glazener, 1946; Lynch et al., 1947), burrowing (Edwards and Frey, 1977; Bertness and Miller, 1984), nest building (Lay and O’Neil, 1942; Lynch et al., 1947; Beard, 1953; Daiber, 1982), and trampling (Gillham, 1955; Ranwell, 1961; Shanholtzer, 1974; Reimold, 1976). Predation is often considered to be an internal, chronic disturbance in some types of ecosystems, particularly in sessile communities such as those in rocky intertidal or coral-reef habitats where predators create bare patches that can be recolonized by other species (Sousa, 1984, 1985; Connell and Keough, 1985). Predation in wetlands, however, is less distinct in time and space, does not necessarily lead to new resources or open space, and consequently is not considered to be a disturbance agent per se as it is in some other systems. However, feeding activities of some wetland animals can impact relative abundances of other organisms directly or by altering their habitat. For example, detritivores such as ﬁddler crabs (Uca pugnax) can decrease the abundance of meiofauna (e.g., nematodes, crustaceans, and annelids) living in salt-marsh sediments because they are ingested along with the marsh soil (Hoffman et al., 1984). Fiddler crabs can also impact meiofaunal populations by burrowing, which aerates the soil substrate and creates refuges from natant and benthic predators (Bell et al., 1978; Hoffman et al., 1984). Anthropogenic disturbance Because wetlands often occur in areas where humans prefer to live (e.g., along coastlines and in river ﬂoodplains), they are often drained and cleared to make room for human habitation, agricultural ﬁelds, and industrial complexes (Tiner, 1984; Maltby, 1986; Pinder and Witherick, 1993). Others are converted to agricultural wetlands (e.g., rice paddies) or to aquaculture systems for production of ﬁsh, shrimps, molluscs, or alligators (Richards, 1993; Simenstad and

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Fresh, 1995; Viles and Spencer, 1995). Some wetlands are managed through modiﬁcation and control of water movement, prescribed burning, or other activities that alter natural regimes (Lay and O’Neil, 1942; Lynch et al. 1947; Lodge, 1994) (Table 13.2). Burning of marshes is frequently conducted to promote growth of plants preferred by wildlife, to remove accumulated litter, and to increase the nutritive quality of forage (Chabreck, 1976, 1981; Angell et al., 1986; Nyman and Chabreck, 1995). Decreased sediment input into wetlands caused by damming of rivers, structural barriers to onshore and alongshore movement of sediment, and reduction of soil erosion in catchments can also constitute a disturbance to wetlands that are dependent on sediment to offset subsidence, erosion and/or sea level rise (Reed and Cahoon, 1992; Viles and Spencer, 1995). Construction of roadways and navigational canals through wetlands leads to direct destruction of wetlands through the breakage, uprooting or burial (under dredge spoil) of vegetation (Burger and Shisler, 1983), and indirect effects through altered hydrology and the introduction of herbicides and other toxins (Gordon, 1988; Gurney and Robinson, 1989; Williams, 1993). Wetlands are also exploited for materials valued by humans, such as, oil and gas, marl, peat, forage for livestock, and timber (Ranwell, 1961; Walsh, 1977; Maltby, 1986; Adam, 1990; Dijkema, 1990; Ngoile and Shunula, 1992; Boesch et al., 1994). Grazing by cattle and other domestic animals removes biomass in wetlands (Reimold et al., 1975), and trampling kills vegetation and disrupts soil structure (Turner, 1987). Human visitors to wetlands can create varying degrees and types of disturbance through activities such as hiking, horseback-riding, boating, bait-digging, camping, and use of off-road vehicles (Burger, 1991; Mercer, 1993). Because of their transitional position between terrestrial and aquatic habitats, wetlands are especially vulnerable to pollution from nutrients (Johnson, 1985), oil spills (Stebbings, 1970; Baker, 1973; Webb et al., 1985; de la Cruz, 1982; Garrity and Levings, 1993), thermal efﬂuent (Dunn and Sharitz, 1991), herbicides (Teas and Kelly, 1975; Gurney and Robinson, 1989), insecticides (Provost, 1977; Keith et al., 1994), heavy metals (DiLabio and Rencz, 1980; Martin and Beckett, 1990; Ott et al., 1993; Gambrell, 1994), and other chemicals (Broom and Pekelder, 1987).

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Fig. 13.1A. Disturbance at different spatial scales in wetlands. Bioturbation by crabs (Ucides cordatus) in a mangrove forest.

Scale and pattern of disturbance The physical, chemical, and biotic characteristics of wetlands determine the effect of a disturbance regime and subsequent rates and patterns of response. Some features may render them more vulnerable to certain types of disturbance; for instance, unconsolidated sediment and shallow plant root systems may increase susceptibility to mechanical damage and erosive forces. Other attributes, however, may moderate the impact of disturbance and promote rapid recovery of wetland sys-

Karen L. McKEE and Andrew H. BALDWIN

tems and their constituent organisms. Those attributes that distinguish wetlands from other systems are often important in determining the magnitude of disturbance and the resilience of the system. Spatial scale of disturbance The spatial scale of disturbance regimes – that is, the areal extent of disturbance to a wetland community – can vary widely. Small-scale disturbances such as windthrow in a swamp, or crab-burrowing (Fig. 13.1A), occur on spatial scales of centimeters to meters. Deposition of dead plant material, or wrack, onto salt-marsh vegetation by tides is another example of small-scale disturbance. In coastal salt marshes of New England, wrack is typically deposited in high marsh habitats, primarily in the spring and early summer, which may damage or kill underlying vegetation and create bare patches in the marsh (Bertness and Ellison, 1987; Valiela and Rietsma, 1995). Wrack deposition also affects animal distributions by creating areas of refuge and concentrations of food resources; for instance, invertebrates such as amphipods are typically found beneath Spartina wrack and other debris in infrequently-ﬂooded marsh habitats (Daiber, 1982). At an intermediate scale, disturbances such as grazing by geese or mammals may remove biomass over some portion of a wetland. Large-scale disturbances such as hurricanes, extensive ﬁres, or logging, on the other hand, may damage or kill vegetation and animals on spatial scales of kilometers (Fig. 13.1B).

Fig. 13.1B. Disturbance at different spatial scales in wetlands. Logging in a swamp in south Louisiana, U.S.A.

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

339

Fig. 13.2A. Patterns of disturbance in wetlands. Hurricane Andrew caused irregular patterns of disturbance in a marsh in southern Louisiana: (1) burial by wrack (light areas of dead vegetation mats) and (2) deep sediment deposits (dark lumps); (3) scouring and erosion (open water); and (4) salt burn (dead standing vegetation).

Fig. 13.2B. Patterns of disturbance in wetlands. Dredging and backﬁlling of canals and oil pipelines in southern Louisiana cause sharp, linear boundaries and geometrical features in the marsh landscape.

Spatial pattern of disturbance Most disturbances generate heterogeneity at some scale, and often occur in patches – relatively discrete areas that are readily distinguishable from undisturbed areas (Pickett and White, 1985b). A downed tree in a swamp, for example, creates a distinct opening in the canopy and alters light and other environmental conditions on the forest ﬂoor. The shape

and distribution of disturbed patches varies with the disturbance agent. Hurricane damage may produce irregular, discontinuous patterns of disturbance in marshes (Guntenspergen et al., 1995) (Fig. 13.2A), whereas passage of marsh buggies and other off-road vehicles leave linear, parallel ruts in marsh vegetation and soil. Other types of human disturbance such as construction of navigational canals and oil pipelines

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and urban expansion also cause linear or geometric patterns (Fig. 13.2B). Disturbance in relation to spatial gradients is of particular interest in wetlands, where strong gradients often exist. Topographical variation leads to gradients in ﬂooding and other hydro-edaphic factors in wetlands, which in turn cause segregation of species according to their abilities to perform along different parts of the gradient. Van der Valk and Davis (1976) found that extremes in the water table of a prairie pothole marsh caused some species to shift position along a ﬂooding gradient, and eliminated others. The overall outcome was an increase in species richness in the submersed and emergent zones, but a decrease in the meadow zone. Hydrological disturbance (construction of a storm-surge barrier) of a salt marsh in the Netherlands caused large-scale dieback of the vegetation, but it subsequently recovered under a new tidal regime (De Jong and Van der Pluijm, 1994). Most species shifted into zones of lower elevation, which corresponded to the original ﬂooding frequencies. In a salt marsh in southern California dominated by perennial glasswort (Salicornia virginica), changes in the salinity gradient caused by above-average rainfall and reservoir discharge allowed invasion by cattail (Typha domingensis) (Zedler and Beare, 1986). Temporal pattern of disturbance Disturbances in wetlands can occur at frequencies of days, weeks, months, years, or centuries, and they may be highly predictable or unpredictable. Events such as ﬁre, hurricanes, ice storms, and grazing by some animals occur in wetlands at distinct intervals, and have different consequences depending on frequency. Fire, for example, may have differential effects on wetland vegetation depending on frequency. In a study of the effects of ﬁre frequency in a tallgrass prairie wetland in Kansas (U.S.A.), plant species diversity was lower in annually burned wetlands than in infrequently burned wetlands (2-, 4-, and 20-yr intervals: Johnson and Knapp, 1995). Fire is also important in maintaining speciﬁc types of wetland communities, such as pitcherplant (Sarracenia spp.) bogs (Eleuterius and Jones, 1969) or forests of Atlantic white cedar (Chamaecyparis thyoides) (Buell and Cain, 1943). The frequency of disturbance may also determine if a particular wetland community returns to some “equilibrium” state. Infrequent disturbance allows sufﬁcient time for re-establishment of plant and animal communities. A study conducted in a Netherlands

Karen L. McKEE and Andrew H. BALDWIN

salt marsh showed that intervals between frost-damage events (5–7 years) were sufﬁciently long for some plant species (e.g., Atriplex portulacoides) to recover to equilibrium (i.e., pre-disturbance) conditions (De Leeuw et al., 1992). Frequent disturbance of an area, however, can alter competitive relationships, allowing other species to become established or increase in abundance. Repeated disturbance by off-road vehicles in some wetlands allows invasion by exotic plant species (e.g., in salt marshes in southern California: Zedler, 1982) or eliminates rare wetland species (e.g., in lakeshore marshes in Nova Scotia: Wisheu and Keddy, 1991). Frequent disturbance can also generate permanent features in a wetland. The female American alligator (Alligator mississippiensis) builds ponds that are small-scale, but persistent, disturbed areas in the Florida Everglades (Lodge, 1994). These “alligator holes”, which may be between 6 and 20 m in diameter, are maintained by successive generations of alligators and may be several centuries old. Predictability refers to the variability in time between disturbances – that is, the regularity of disturbance. The effects of predictability of disturbances on wetlands is not clear. In a broad sense, however, some types of wetlands are more subject to regular, predictable disturbance than others. Tidal wetlands, for instance, are regularly inundated on a predictable schedule, and this physical disturbance has undoubtedly been a major selective force in the evolution of intertidal organisms; for instance, the feeding and reproductive cycles of many marine organisms are keyed to the tidal cycle (e.g., the gastropod Melampus bidentatus: Russell-Hunter et al., 1972). In contrast to tidal cycles, events such as heavy rainstorms may suddenly and unpredictably alter water levels and/or salinities in both tidal and non-tidal wetlands. Sessile organisms (e.g., oysters and clams) cannot escape and may suffer extensive mortality subsequent to ﬂooding and salinity reduction caused by a hurricane, whereas motile organisms (e.g., ﬁsh and crustaceans) typically disperse to areas with more favorable conditions (Knott and Martore, 1991). Magnitude of disturbance Magnitude of disturbance can be divided into disturbance intensity (the physical force of a disturbance) and disturbance severity (the amount of damage caused to organisms, communities, or ecosystems) (Sousa, 1984; White and Pickett, 1985). Measures of disturbance intensity include wind speed for hurricanes,

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

Fig. 13.3A. Intensity and frequency of disturbance varies in wetlands. Insect damage on red mangrove (Rhizophora mangle) leaves in Belize, Central America.

temperature for ﬁres, and animal densities for grazing. Insect herbivory, for example, may be a low-intensity disturbance decreasing leaf biomass (Fig. 13.3). Woodboring insects (e.g., Elaphidion mimeticum) can kill large branches or entire trees, creating canopy gaps in forests of red mangrove (Rhizophora mangle) (Feller and McKee, 1999). Severity of disturbance can be measured in terms of mortality, type of structural

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damage (e.g., defoliation vs. trunk snap in a swamp) or percent of an area affected (e.g., percent of a marsh buried by sediment). In Hong Kong, for example, insect outbreaks in mangrove forests can periodically defoliate large areas (Anderson and Lee, 1995) Disturbances of a uniform intensity may differ dramatically in the severity of damage they inﬂict among different community types. For example, Hurricane Andrew caused high tree mortality (60–85%) in mangrove forests of southern Florida (Baldwin et al., 1995), whereas mortality of bald cypress (T. distichum var. nutans) in cypress domes ranged from about 1% to 4% depending on location relative to the eye of the hurricane (Noel et al., 1995). Within mangrove forests, differential ﬂood tolerance and spatial position determine the pattern of tree mortality due to impoundment (K. McKee, pers. observ.). Similarly, some species of marsh plants are relatively unaffected or even stimulated by oil spills, whereas others suffer widespread mortality (Lin and Mendelssohn, 1996) (Fig. 13.4). Many anthropogenic disturbances are too drastic to allow adjustment or adaptation by communities (Odum, 1969). In some cases, the disturbance agent may be outside the evolutionary history of the species in a community and characterized as inﬁnitely severe. Examples include permanent drainage, repeated clear-cutting, and toxic chemical releases. However, other anthropogenic disturbances may mimic natural disturbances. These include livestock grazing (Fig. 13.3B), periodic hay

Fig. 13.3B. Intensity and frequency of disturbance varies in wetlands. Livestock grazing in a salt marsh.

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Fig. 13.4. Severity of disturbance can vary depending on a number of factors. Plant species exhibit different degrees of sensitivity to petroleum hydrocarbons in a freshwater marsh in southern Louisiana (U.S.A.). (1) Sagittaria lancifolia (broad-leaved plant growing in clumps and in background) is most resistant to oiling, but other species such as (2) Leersia oryzoides and Panicum hemitomon are killed (darkened areas of dead vegetation).

harvesting, water-level manipulation, changes in sediment loading, and prescribed burns. However, an anthropogenic disturbance agent may vary in its intensity and severity. Intense disturbance by all-terrain vehicles was found to destroy vegetation as well as the seed bank in a rare wetland plant community, while moderate levels of disturbance altered community composition but did not eliminate vegetation or seed banks (Wisheu and Keddy, 1991). Interactions Disturbance events may interact to generate a greater combined effect. For example, geese are attracted to ice-scour depressions in salt marshes, and their feeding in these disturbed areas causes even greater disruption of the substrate than either agent alone (B´elanger and B´edard, 1994). Grazers are also attracted to newly burned marshes, because the regrowth following a ﬁre is often succulent and more nutritious than the original vegetation (Chabreck, 1976; Faulkner and de la Cruz, 1982; Smith and Kadlec, 1985). Grazing by nutria may increase susceptibility of some marshes to the effects of saltwater intrusion by reducing accumulation of organic matter and vertical accretion (e.g., in coastal Louisiana: Grace and Ford, 1996). Similarly, frequent or intense

ﬁre may lead to increased rates of soil erosion or greater ﬂooding due to a lowering of marsh surfaces by peat burns (Nyman and Chabreck, 1995). Turner (1988) experimentally examined effects of multiple disturbance agents in a Spartina alterniﬂora salt marsh, and found that combinations of disturbances (clipping, trampling, and burning) caused effects on standing biomass and net above-ground primary production, some of which were additive and other nonadditive (i.e., greater or less than the sum of effects). Clipping and trampling, for example, had additive effects on standing stock, but combinations of ﬁre with clipping or trampling had less effect than expected. Disturbance agents may interact with environmental factors in a wetland to affect severity of the disturbance or degree of recovery. In Louisiana freshwater marshes, a combination of saltwater intrusion and increased water levels generated by storms can cause dieback of vegetation (McKee and Mendelssohn, 1989). The recovery of these freshwater marsh communities from saltwater intrusion was determined by post-disturbance salinity and water levels (Flynn et al., 1995). In salt marshes in Western Australia, the combination of physical disturbance to the native species (Juncus

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

kraussii) and lowered salinity (associated with street drains cut into the marsh) allowed the invasion and spread of Typha orientalis, an introduced species (Zedler et al., 1990). Studies in marshes along the coast of the Gulf of Mexico have also shown that rates of revegetation after grazing by waterfowl is determined by post-herbivory hydroperiod and water depth, salinity, storm-tide timing and severity, and water turnover (Lynch et al., 1947; Miller et al., 1996). In South Carolina, regeneration of Taxodium– Nyssa swamps after Hurricane Hugo was determined by the combined effects of wind damage and saltwater intrusion (Conner, 1995). Interactions between disturbance type and timing can result in different patterns of effects and responses. Timing of prescribed ﬁres controls plant species composition in coastal marshes (Chabreck, 1981). Fall or winter ﬁres promote growth of Olney’s three-square (Scirpus americanus), whereas spring burning promotes dominance by salt meadow cordgrass (Spartina patens). In Manitoba, spring and fall ﬁres increase shoot and root biomass of common reed (Phragmites australis), but summer burning decreases it (Thompson and Shay, 1985). Timing of marsh burning also affects wildlife, since spring and summer ﬁres can kill young animals or destroy nests.

EFFECTS OF DISTURBANCE ON WETLAND BIOTA

Many disturbances directly damage or kill wetland plants and animals. Sessile organisms, such as rooted plants, mussels, oysters and barnacles, can be buried by sediment deposition, killed by sudden changes in salinity, or consumed by other wetland animals or humans (Daiber, 1982; Diaz, 1994; Viles and Spencer, 1995). Mobile animals can often escape small-scale disturbances, but dispersion may increase exposure to predators or decrease reproductive success (Burger, 1991; Fernandez and Azkona, 1993; Monda et al., 1994; Reinert and Mello, 1995). Small organisms that live on or just under the sediment surface may be dislodged by physical factors such as splashing water (Mendoza, 1982) or bioturbation, and carried by water currents or tides to unfavorable locations. Bioturbation by animals may inhibit plant establishment; for instance, the amphipod Corophium volutator prevents establishment of the halophyte Salicornia europaea at low tidal elevations

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by burying seeds and inhibiting seedling establishment (Gerdol and Hughes, 1993). Disturbances that remove plant biomass, reduce stem densities, or alter canopy structure affect animal populations by altering availability of food, refuge, and substrate (Zimmerman and Minello, 1984; Capehart and Hackney, 1989; LaSalle and Rozas, 1991; Batzer and Resh, 1992). A burned marsh or clear-cut swamp, for example, is unsuitable for birds and mammals that need the plant material for nest-building or for cover (Bray, 1984; Kinler et al., 1987; Burdick et al., 1989). Edaphic algae are important primary producers in tidal marshes, and their distribution and relative abundance are affected by humidity and light (Sullivan, 1976), which are in turn determined by dominant vegetative cover (Blum, 1968). Micro-organisms may be stimulated by disturbances such as wind-storms that cause sudden deposition of large amounts of organic matter on a wetland soil surface, thereby increasing the substrate for decomposers.

WETLAND RESPONSE TO DISTURBANCE

Responses to disturbance in wetlands may be examined at different scales of organization. In this section we consider responses at species, population, community, ecosystem, and landscape levels, which are summarized in Table 13.3. To identify general patterns of response, we surveyed the results of ﬁfty-four selected studies conducted in North America reporting ninety instances of disturbance-induced changes in vegetation cover, biomass, primary production, relative abundance of species, and/or species richness (Table 13.4). Most of these studies involved ﬁeld experiments, but a few were ﬁeld observations or laboratory experiments. The results include responses to natural, anthropogenic, and simulated disturbance regimes categorized as to type (physical, grazing, ﬁre, salinity ﬂuctuation, toxins, and interactions). Studies reporting responses in all major wetland types were included, but most observations were in salt marshes. Patterns of response to disturbance are discussed in the following subsections. Species responses About half of the disturbance studies we surveyed showed no permanent change in percent cover (i.e., recovery occurred within the study period), but the

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Table 13.3 Wetland responses to disturbance at different levels of organization Organizational level

Disturbance regime

Characteristics affected

Typical response/strategy

Species

small-scale, mild

life history strategy

Vegetative propagation

Population

intermediate-scale, moderate

suppressed juveniles, seedbank

large-scale, intense

widely-dispersed diaspores

environmental ﬂuctuation

phenotype

morphological and reproductive plasticity; physiological acclimation

genotype: multiple events

intra-population

predictable Community

Ecosystem

Landscape

inter-population

polymorphism adaptive strategies

intermediate intensity, frequency

structure

increased diversity, species richness

altered environmental conditions or selective removal of dominant species

competitive interactions

increased relative abundance of subordinate species

large-scale, intense

succession

delayed or “reset”

alternating conditions

succession

cyclic

intense and/or frequent

primary production

decrease

type (e.g., ﬁre, grazing)

carbon and nutrient dynamics

redistribution, consumption, transformation

natural

spatial conﬁguration

irregular patterns, continuous area, gradual ecotones

anthropogenic

boundaries

regular, linearized, fragmented, sharp

remainder exhibited a net decrease in vegetative cover (Table 13.4). Part of these differences may have been due to differences in duration of the observation period, which varied substantially among studies and probably contributed to this pattern. However, the majority of the studies reporting changes in vegetative cover dealt with disturbance by grazing, and were conducted in saline marshes dominated by perennial graminoids. Differences in life-history features of the dominant plant species interacting with the characteristics of the disturbance agent appear to be important factors affecting response to disturbance (McIntyre et al., 1995) and likely inﬂuenced rates and patterns of revegetation in the studies summarized in Table 13.4. Wetland plants exhibit different types of strategies (e.g., vegetative propagation, suppressed individuals, buried seeds, and widely-dispersed propagules) that determine rates and patterns of recovery from disturbance. Disturbances such as ﬁre or grazing often remove aboveground biomass, after which many perennials can resprout and quickly re-establish following disturbance (Olson and Platt, 1995; Grace and Ford,

1996). In contrast, annual species may be killed by these disturbances and depend on seedling recruitment (which may be slow if the seed source is at a distance) to repopulate disturbed areas. Many species of marsh plants colonize small disturbance-generated patches vegetatively, and some rely on physiological integration between connected ramets to cope with environmental stresses (e.g., hypersalinity or drought) associated with the patch environment. Transfer of water (Hester et al., 1994; Shumway, 1995) and carbon (Shumway, 1995) from parent to daughter ramet during invasion of a drier or saltier bare patch ensures sufﬁcient resources for growth and reduces energy expenditure for synthesis of organic osmotica (Hester et al., 1994). Wetland fauna also exhibit different life-history characteristics that have consequences for recovery from disturbance. Amphipod crustaceans, which are common to both marine and freshwater wetlands, are differentially affected by disturbance depending on speciﬁc attributes determining relative mobility, ability to ﬁnd food, and rates of recolonization of disturbed patches (Conlan, 1994). Amphipods that

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Table 13.4 Results of selected studies conducted in North America reporting effects of disturbance on wetland vegetation Reference

Disturbance Type 1

Allison (1995)

Cat. 2

Study type 3

Wetland type 4

Plant form 5

Plant type 6

Response 7 C%

Bio

NPP

RA

SR

S

BS

FE

SM

M

M

0

+

−

S

BS

FE

SM

D

P

0

−

0

Baldwin et al. (1995)

N

B/A

FO

M

W

P

−

Baldwin (1996)

S

BR

FE

BM

G

P

−

+

0

0

+

0

S

BR

FE

BM

D

P

0

0

+

0

S

B/A

FE

BM

G

P

−

−

+

+

S

B/A

FE

BM

D

P

−

−

+

−

S

S

LE

BM

M

P

+

0

S

WL

LE

BM

M

P

+

0

S

I

LE

BM

M

P

N

G

FE

SM

G

P

Beare and Zedler (1987)

S

S

LE

SM

G

P

B´elanger and B´edard (1994)

N

G

FE

SM

G

P

0

0

0

0

N

B/A

FE

SM

G

P

−

0

0

0

−

−

−

+

0

0

Bazely and Jefferies (1986)

Bertness (1984)

N

I

FE

SM

G

P

N

G

FE

SM

G

P

Bertness (1985)

N

BI

FE

SM

G

P

Burger and Shisler (1983)

A

BS

FE

SM

G

P

Burk (1977)

A

T

FO

FM

M

M

Cargill and Jefferies (1984)

N

G

FE

SM

G

P

Castaner and LaPlante (1992)

N

WL

FO

FM

M

M

−

+

+

−

+

−

+

− + 0

0

− −

+ +

Delaune et al. (1984)

A

T

FE

SM

G

P

Doyle et al. (1995)

N

B/A

FO

M

W

P

Ehrenfeld (1995)

N

B/A

FE

SW

W

P

0

N

F

FE

SW

W

P

+

N

G

FE

SM

D

A

−

N

BW

FE

SM

D

A

+

N

I

FE

SM

D

A

0

S

S

LE

FM

M

P

−

S

WL

LE

FM

M

P

0

Furbish and Albano (1994)

S

G

FE

SM

G

P

Giroux and B´edard (1987)

N

G

FE

SM

G

P

Grace and Ford (1996)

S

BR

FE

FM

D

P

0

Ellison (1987)

Flynn et al. (1995)

Guntenspergen et al. (1995)

0

0

0

0

−

0

− −

−

−

+

0

0

−

−

−

S

WL

FE

FM

D

P

0

S

S

FE

FM

D

P

0

S

I

FE

FM

D

P

N

B/A, BS

FE

BM

G

P

0

−

N

BW, E

FE

BM

G

P

−

continued on next page

346

Karen L. McKEE and Andrew H. BALDWIN

Table 13.4, continued Reference

Disturbance Type 1 Cat. 2

Hartman (1988) Hik and Jefferies (1990)

Study type 3

Wetland type 4

Plant form 5

Plant type 6

Response 7 C%

Bio

NPP

N

BW

FE

SM

M

M

N

G

FE

SM

G

P

+

+

S

BR

FE

SM

G

P

0

0

N

G

FE

SM

G

P

+

Hik et al. (1991)

S

G

LE

SM

G

P

+

Hik et al. (1992)

N

G

FE

SM

M

M

Hogenbirk and Wein (1991)

N

F

FE

FM

G

P

N

F

FE

FM

D

A

0

N

WL

FE

FM

G

P

0

N

WL

FE

FM

D

A

0

−

0

0

−

+

N

BI/S

FE

BM

M

M

Johnson and Knapp (1995)

N

F

FE

FM

G

P

Jonsson-Ninniss and Middleton (1992)

A

E

FE

B

M

M

−

Kerbes et al. (1990)

N

G

FO

FM

G

P

−

N

G

FO

SM

G

P

−

A

T

LE

SM

G

P

0

A

T

LE

BM

G

P

−

A

T

LE

FM

D

P

+

Loope et al. (1994)

N

B/A

FO

SW

W

P

Mallik and Wein (1986)

A

WL

FE

FM

G

P

+

A

F

FE

FM

G

P

A

I

FE

FM

G

P

N

S

FE

FM

M

P

McKee and Mendelssohn (1989)

SR

0

Iacobelli and Jefferies (1991)

Lin and Mendelssohn (1996)

RA

−

− −

−

+

−

0

+

+

+

0

0

−

−

−

+

−

−

−

N

WL

FE

FM

M

P

−

0

N

I

FE

FM

M

P

−

−

Mika et al. (1985)

A

T

FE

FM

M

M

Miller et al. (1996)

N

G

FO

SM

G

P

Monda et al. (1994)

S

G

FE

FM

G

P

Putz and Sharitz (1991)

N

B/A

FO

SW

W

P

Reimold et al. (1975)

A

G

FE

SM

G

P

−

Scott and Haskins (1987)

N

G

FO

FM

D

P

−

Shumway and Bertness (1992)

N

S

LE

SM

G

P

Slavin and Shisler (1983)

A

WL

FO

SM

G

P

Smith and Kadlec (1985)

N

G

FE

BM

G

P

N

F

FE

BM

G

P

0

0

N

I

FE

BM

G

P

−

−

Smith and Odum (1981)°

N

G

FE

SM

G

P

−

−

Smith (1982)

N

G

FE

SM

G

P

−

−

+

− −

+ + −

−

+

− + 0

0

− continued on next page

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

347

Table 13.4, continued Reference

Disturbance Type 1 Cat. 2

Study type 3

Wetland type 4

Plant form 5

Plant type 6

Response 7 C%

Bio

NPP

SR

+

0

Smith et al. (1994)

N

B/A

FO

M

W

P

Srivastava and Jefferies (1996)

N

G

FE

SM

G

P

−

Taylor and Grace (1995)

N

G

FE

FM

M

M

−

N

G

FE

BM

M

M

−

N

G

FE

SM

G

P

−

Thompson and Shay (1985)

A

F

FE

FM

G

P

Turner (1988)

S

BR

FE

SM

G

P

−

0

0

0

S

B/A

FE

SM

G

P

−

−

0

0

S

F

FE

SM

G

P

−

−

0

0

S

I

FE

SM

G

P

−

−

0

0

Umbanhowar (1992)

N

B/A

FE

FM

M

M

Webb et al. (1985)

A

T

FE

SM

G

P

Williams (1993)

A

E

FO

SM

G

P

Zedler and Beare (1986)

A

S

FO

SM

D

P

−

RA

0 0 0

0

0

−

− 0

0

0

− −

1

Abbreviations: N, natural; A, anthropogenic; S, simulated. Abbreviations: Cat., category; F, ﬁre; G, grazing; BS, burial by sediment; BW, wrack; BA, breakage/abrasion; E, soil erosion; BR, biomass removal; BI, bioturbation; T, toxins; S, salinity; WL, water level ﬂuctuation; I, interactions between two or more disturbance categories. 3 Abbreviations: FE, ﬁeld experiments; FO, ﬁeld observations; LE, laboratory experiments. 4 Abbreviations: SM, salt marsh; BM, brackish marsh; FM, freshwater marsh; B, bog; SW, swamp; M, mangrove. 5 Dominant plant forms were graminoid (G), dicotyledonous (D), woody (W), or a mixture (M). 6 Abbreviations: P, perennial; A, annual; M, mixture. 7 Responses reported were percent cover (C%), biomass (Bio), net primary production (NPP) (above-ground in all but one case° ), relative abundance of dominant species (RA), and/or species richness (SR). Direction of change in response is indicated as positive (+), negative (−), or none (0). 2

are highly mobile are more successful than more sedentary species in ﬁnding ephemeral food sources after a disturbance event. Benthic amphipods, however, may control post-disturbance community structure better than pelagic species, because they can modify the sediment and consume settling larvae of other amphipod species or macroalgae that compete for space. Many benthic invertebrates in salt marshes are reproductive nearly year-round (Daiber, 1982), ensuring rapid recolonization after a disturbance. In some types of wetlands, severity and scale of disturbance may determine successful regeneration strategies, primarily for sedentary or sessile organisms. Mild chronic disturbance, such as bioturbation by burrowing crabs, causes a strategy to be selected which is different from that appropriate for severe, periodic disturbance such as ﬁre. A disturbance that creates small gaps in a vegetative canopy (e.g., windthrow that downs a single tree in a swamp, or wrack deposition

in a salt marsh) is quickly closed by encroachment of the surrounding vegetation; no succession occurs, and the patch eventually resembles the original vegetation. The regeneration strategy that is successful in this disturbance regime is vegetative propagation (Allison, 1995). Reproductive strategies of wetland animals may also inﬂuence response to disturbance. Some benthic animals exhibit brood protection (e.g., polychaetes: Kendall, 1979) – a strategy that would be advantageous in mild disturbances that do not cause mortality of adults, but decrease refuges from predators for eggs and larvae. Intermediate levels of disturbance may create openings that cannot be quickly ﬁlled by lateral expansion of edge organisms, but relatively small patches may be reclaimed by regrowth from damaged individuals or by suppressed juveniles already present in the patch. In freshwater or mangrove swamps, windstorms may create canopy gaps by snapping boles of trees, but leave

348

suppressed seedlings and saplings either undamaged or with minimal damage (Smith et al., 1994; Baldwin et al., 1995). Once released from suppression by adult vegetation, the juvenile plants will begin growing rapidly, and there will be intense competition for dominance of the space (Ellison and Farnsworth, 1993). Somewhat larger patches of disturbance can be colonized by seedlings recruited from the seed bank (Bakker and de Vries, 1992; Hogenbirk and Wein, 1992; Legg et al., 1992; Bonis and Lepart, 1994; Bonis et al., 1995) or by propagules recently dispersed into them, but only when conditions allow germination and/or establishment [e.g., in oligohaline marshes (Baldwin et al., 1996) or mangrove forests (McKee, 1995)]. Depending on species composition of the seed or seedling bank and competitive interactions, succession may occur. Haukos and Smith (1993), for example, predicted vegetational composition in playa lakes in southern Texas (U.S.A.) based on lifehistory characteristics of the species represented in the seedbank. An animal strategy that would allow rapid recolonization of areas in which the vegetation, but not the soil, was disturbed is illustrated by some invertebrates, which develop directly from eggs deposited in benthic capsules (Daiber, 1982). Severe disturbance, in addition to destroying the original community, may cause two changes that impact regeneration. Juveniles and propagules may be eliminated from the disturbed area, and the distance from a source of replacements may be increased, so that recovery depends on successful dispersal and establishment of propagules. Recolonization by clonal growth, for example, was found to be ineffective in revegetation of large bare patches created by lesser snow geese (Anser caerulescens) in Hudson Bay salt marshes (Kerbes et al., 1990). If a severe disturbance destroys a large area, including potential recruits (e.g., the seed or seedling bank) and the distance from reproductive adults is great, then the site must be recolonized by species with widely-dispersed seeds or propagules at the time the patch is created. This strategy is exhibited by mangroves, which have propagules that are buoyant and potentially capable of being dispersed over long distances by tides and ocean currents (Rabinowitz, 1978). Dispersal occurs during the period of peak hurricane activity, ensuring a supply of new recruits to a forest that may be damaged or destroyed by a hurricane (Tomlinson, 1986). Some animals with lengthy reproductive seasons and pelagic larval stages (e.g., some benthic invertebrates) would

Karen L. McKEE and Andrew H. BALDWIN

be expected to recolonize a severely disturbed site more rapidly than those with different life-history features. Thus, after a severe disturbance there may be a long period of recolonization and succession, and species with widely-dispersed diaspores will dominate initial stages of recovery. The life-history characteristics of wetland organisms can be used to identify particular species that respond to disturbance by increased or decreased abundance. Certain species such as narrow-leaf cattail (Typha angustifolia) and purple loosestrife (Lythrum salicaria) may be important indicators of disturbance in wetlands (Wilcox, 1995). The appearance of plants commonly found in roadside ditches or on mudﬂats exposed by low water can thus indicate changes in a wetland’s hydrology, or some other disturbance that promotes invasion. Shifts in vegetation zonation patterns may also reﬂect the responses of individual species to disturbance (van der Valk and Davis, 1976). Population responses Organisms in wetland environments are subject to a myriad of disturbance agents occurring on widely ranging scales of space and time. Some disturbance regimes may have profound ecological and evolutionary effects on wetland plant and animal populations. Physical, chemical, and biological agents of disturbance may promote phenotypic changes in response to new conditions. Alteration of physical or biological processes by disturbance also modiﬁes selection pressures and causes changes in the structure and dynamics of populations (Urbanska, 1987). Differential mortality and reproduction may lead to population differentiation and, ultimately, speciation. Population responses to a disturbance regime may be genetic or environmentallyinduced modiﬁcations (Snaydon, 1987). Genetic responses include changes in anatomy, morphology, physiology, and biochemistry, while environmentallyinduced modiﬁcations include phenotypic plasticity and acclimation to new environmental conditions. The response depends on the magnitude or amplitude of the environmental variation with respect to the size of individuals (at least for plants and other sessile organisms) or to the lifespan of the species. The most common responses to ﬂuctuations in the environment during the lifespan of an organism are phenotypic plasticity and physiological acclimation. A wetland annual herb, Murdannia keisak, for example, exhibits morphological and reproductive plasticity to

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

disturbance by thermal efﬂuent (Dunn and Sharitz, 1991). Populations in thermally disturbed areas were shorter, ﬂowered earlier, and produced more, smaller seeds than those in revegetating or undisturbed sites. Absence of population variation under common conditions in the greenhouse indicated that most of the observed differences between disturbed and undisturbed populations were due to a phenotypic response to water depth and light. The weed-like characteristics of this species explain its rapid adjustment and dominance in disturbed habitats. Physiological acclimation to temperature extremes is also exhibited by salt-marsh snails (M. bidentatus: McMahon and Russell-Hunter, 1981) and isopods (Sphaeroma rugicauda: Marsden, 1973). Prior exposure to a speciﬁc temperature regime for a period of time allowed adjustment and improved performance during temperature extremes. Temporal changes of a magnitude similar to the life-span, or spatial variation similar in amplitude to an organism’s size, are usually associated with polymorphism within populations (i.e., intrapopulation genetic variation). Seed polymorphism is a plant trait that may reﬂect adaptation to a varying environment, or promote differential dispersal (Harper, 1977). Salicornia europaea, for example, is an annual halophyte that produces seeds of two different sizes. Ellison (1987) showed that the larger seeds germinated early in the spring, whereas smaller seeds germinated later in the summer. The large seeds would produce larger seedlings which were more capable of competing with other marsh species in a disturbance patch created by wrack deposition. Small Salicornia seeds persist in the seed bank (Philipupillai and Ungar, 1984) – a strategy that would be successful in areas experiencing multiple disturbance events. Genetic responses to disturbance may lead to changes that are adaptive to the particular agent of disturbance (although not all responses are adaptive). Pond pine (Pinus serotina), for example, produces epicormic sprouts and serotinous cones that release seeds upon exposure to ﬁre (Ewel, 1990). Salicornia europaea depends on disturbance patches for successful colonization in New England salt marshes. Differences in seed morphology explained higher Salicornia seedling densities in areas exposed to wrack deposition as compared to artiﬁcial patches protected from wrack by mesh fences (Ellison, 1987). The ﬁne hairs on the S. europaea seeds trap air bubbles, promoting ﬂotation, but also cause them to attach to

349

pieces of wrack. The seeds are thus deposited along with the wrack, which is the disturbance agent. Speciﬁc differences in stress tolerance determine community response to environmental ﬂuctuation, for instance reductions in salinity in salt marshes (Zedler et al., 1990; Beare and Zedler, 1987) or increases in salinity in freshwater or brackish marshes (Iacobelli and Jefferies, 1991; Flynn et al., 1995). However, adult plant populations in wetlands may also exhibit genetic divergence related to certain types of disturbance, such as hypersalinity, increased water level, or heavy metal pollution. Among populations of the brackish marsh species Spartina patens on the coast of the Gulf of Mexico, for example, there is intraspeciﬁc variation in lethal salinity level as well as morphological characteristics (Hester et al., 1996). Populations of S. alterniﬂora and a freshwater marsh species (Panicum hemitomon) also vary in response to salt as well as to ﬂooding (Lessmann et al., 1997). Thus, genotypic differences within a species can be important in response of wetland plants to disturbance. Despite the potential role of disturbance in inﬂuencing evolutionary processes, there are many gaps in knowledge of population response to disturbance in general, and limited information is available concerning plant and animal populations in wetlands. The few examples given above illustrate the importance of such information to an understanding of the role that disturbance plays in wetlands. Community responses Patch dynamics Disturbance can be a discrete event creating relatively distinct boundaries between disturbed and undisturbed areas. These discontinuities, gaps between what have been termed patches, provide marked differences in opportunities for establishment and growth (Pickett and White, 1985a,b; Veblen, 1992). Patchy disturbance may act to maintain or enhance existing heterogeneity within a community. Sediment deposition in a Louisiana salt marsh during Hurricane Andrew, for example, was higher in Juncus roemerianus stands than in Spartina alterniﬂora stands because of higher stem density in the former (Nyman et al., 1993). Differences in sediment loading may act to maintain the distribution of these two species by creating heterogeneity in elevation, nutrient level, or soil texture. Environmental conditions such as light, temperature, and humidity within disturbed patches often differ

350

dramatically from those in undisturbed areas (Denslow, 1980; Veblen, 1992). Because changes in environmental conditions serve as germination cues in many plant species (Fenner, 1985; Leck, 1989, 1996), disturbed patches are opportunities for recruitment of new plant species into the vegetation. Space may also be more available in disturbed patches, which is important for plants as well as sessile animal communities such as barnacles and mussels because of reduced competition (Sousa, 1984, 1985). There are several examples of the stimulatory effect of patch creation on recruitment in wetlands. In New England salt marshes, Salicornia europaea seedlings colonize disturbed patches, but cannot compete with perennials in undisturbed areas (Ellison, 1987). Seedlings of Atlantic white cedar are also more abundant in sites disturbed by ﬁre or tree blowdown than in undisturbed sites (Motzkin et al., 1993; Ehrenfeld, 1995). Vegetative recolonization by clonal growth was found to be ineffective in large bare patches created in Hudson Bay marshes by goose grubbing, but seedling recruitment was extensive (Kerbes et al., 1990). Anthropogenic disturbance by clearcutting enhanced relative abundance of red maple (Acer rubrum) in Massachusetts coastal swamps (Motzkin et al., 1993). Other factors affecting colonization in disturbance patches are shape, orientation, edaphic conditions, timing and periodicity of patch creation, size-range of patches, and within-patch environmental heterogeneity (Denslow, 1980; Veblen, 1992), although few quantitative studies of these factors exist for wetlands. Umbanhowar (1992) studied the effects of mounds of thatching ants, earthen mammal mounds, bison (Bison bison) wallows, and openings in dry marshes in a northern mixed prairie that consisted of both upland and wetland habitats. Small, abundant disturbance patches (ant and earthen mounds) were dominated by vegetatively reproducing perennials, while annuals were more abundant in larger disturbance patches that occurred infrequently (bison wallows and dry marshes). Certain wetland species were restricted to those disturbed patches that were wet (wallows or marshes). Species diversity The diversity of plant and animal species in wetlands may be altered by disturbance. In our survey of North American wetland studies, we found that half of the observations indicated no change in plant species richness after disturbance, and this pattern held across

Karen L. McKEE and Andrew H. BALDWIN

disturbance categories (Table 13.4). Most of these studies were conducted in salt or brackish marshes, suggesting that the composition of these communities was relatively resistant to the disturbance regimes studied. The remainder, which spanned both freshwater and saline habitats, showed either an increase or a decrease in species richness. There was also no consistent pattern as to effect of disturbance on relative abundance of dominant species (i.e., those dominant prior to the disturbance event) (Table 13.4). The observations were evenly divided among those showing an increase, a decrease, or no change in relative abundance. When examined by wetland type, more species of salt and freshwater marsh plants showed no change or a decrease in dominants after disturbance, whereas most of the species of brackish marshes and swamps showed an increase (Table 13.4). Physical disturbance generally caused no change or an increase in abundance of dominants, although this response varied depending on the speciﬁc type of disturbance. Grazing, salinity changes, and interactions (among two or more disturbance factors) resulted in decreases in relative abundance of dominants in some cases. Livestock grazing, however, may increase species richness or diversity, as reported in some European salt marshes (Jensen, 1985; Andresen et al., 1990). Variation in effects of disturbance on species richness and relative abundance may be related to differences in disturbance intensity. Disturbance at some intermediate level of intensity or frequency can create opportunities for establishment and persistence of several species, resulting in a more diverse community than would occur at lower or higher levels of disturbance. Low frequency or intensity of disturbance can lead to dominance by a few or even a single species through competitive exclusion. At the other extreme, very frequent or intense disturbances can eliminate most species, again creating a community of low species diversity. The concept of maximum diversity at intermediate levels of disturbance has been termed the Intermediate Disturbance Hypothesis, a classic theory formulated by Connell (1978) to explain diversity in tropical forests and coral reefs. While the Intermediate Disturbance Hypothesis has been supported by studies in a number of habitats, there have been relatively few studies speciﬁcally investigating the effects of disturbance frequency or intensity on species diversity in wetlands. In a Louisiana coastal marsh, however, highest species richness of vegetation occurred at intermediate intensities of disturbance by

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

grazing mammals, wrack, and scouring (Grace and Pugesek, 1997). Other studies, although not directly testing effects of disturbance intensity on species diversity, have also found similar patterns (Mallik and Wein, 1986; Guntenspergen et al., 1995). Intermediate levels of disturbance may also maximize faunal diversity by creating heterogeneous plant communities through patch generation, and some studies indicate highest species diversity in heterogeneous landscapes (Bowland and Perrin, 1993). Competitive interactions In general, disturbance involves the removal of biomass or alteration of resource levels. As biomass is reduced or resources are altered with increasing disturbance frequency or intensity, competitive pressures should change. The relationship between disturbance and competition is a cornerstone of Grime’s (1977) triangular model of plant strategy, which suggests that disturbed habitats (at least those with adequate resources for growth) will be colonized by weak competitors having high reproductive effort and high growth rate (“ruderal” syndrome) – that is, disturbance reduces competitive pressure. Another model speciﬁc to wetlands is the centrifugal model of community organization (Keddy, 1990), which suggests that maximum competitive intensity (as well as maximum biomass and lowest species diversity) occurs under undisturbed, high-fertility conditions. As biomass is reduced by disturbance (or by reduced fertility), competition intensity will decrease and species diversity will increase. This model is based on studies of biomass and species changes along gradients of stress and disturbance in several wetland systems. Few investigators have explicitly examined effects of disturbance on competition intensity in wetlands, although the relationship can be inferred from a number of studies. In a study of plant competitive ability along gradients of stress and disturbance in a lakeshore wetland, disturbance selectively removed competitive dominants, allowing persistence of competitive subordinates (Wilson and Keddy, 1986). Plant traits such as tall, leafy shoots that contributed to competitive dominance rendered plants more susceptible to wave damage, whereas wave-resistant morphologies such as small rosettes were characteristic of weak competitors. Disturbance may also alter competitive relationships by opening up space and/or causing stress conditions and changes in availability of resources. Fire, for example, creates bare patches in vegetation and releases

351

bound-up nutrients. These changes alter the competitive balance, for instance between saw grass (Cladium jamaicense) and maidencane (Panicum hemitomon) in Florida (Lowe, 1986) or between herbaceous and woody species in freshwater marshes (Ewel, 1986). Disturbance by wrack deposition or by grazers may cause environmental changes (e.g., increased solar radiation, soil temperatures, and salinity in salt marshes as a result of removal of the plant canopy) that in turn affect competitive interactions (Bertness and Ellison, 1987; Bertness, 1991a; Iacobelli and Jefferies, 1991; Bertness et al., 1992; Srivastava and Jefferies, 1996). Biotic disturbances such as grazing can also directly alter competitive interactions among wetland plants and animals and thereby contribute to the pattern of either increased or decreased abundance of plant dominants in response to grazing (Table 13.4). Plant community structure and species distributions in a salt marsh on Assateague Island, Maryland (U.S.A.), were affected by selective herbivory (Furbish and Albano, 1994). Simulated grazing on Spartina alterniﬂora (lowmarsh dominant) reduced its competitive ability and therefore its abundance relative to Distichlis spicata. In a brackish marsh in Louisiana, intense herbivory by nutria was found to decrease the abundance of dominant perennials and increase light penetration, which appeared to prevent competitive exclusion of other species from the community (Ford, 1996). Competition is not limited to adult plants; the dense canopy of vegetation in some wetlands can also prevent colonization by seedling recruitment (Metcalfe et al., 1986; Falinska, 1991). Disturbance in an oligohaline marsh in Louisiana (U.S.A.) dominated by Spartina patens stimulated seedling recruitment by a number of species otherwise rare in the community (Baldwin, 1996), indicating that seedling recruitment was competitively inhibited by S. patens. Succession Plant and animal communities are subjected to a wide array of disturbances and ﬂuctuating environmental conditions that may set back, deﬂect, or slow the progression to an equilibrium state (Connell, 1978). In a disturbance regime in which space, light, and/or nutrients are altered, the community can be “reset” to an earlier state of succession. In wetlands, disturbance may result in cyclic succession. In a prairie glacial marsh in Iowa, a series of events (water-level drawdown, reﬂooding, and grazing by muskrats create a cyclical succession of plants dependent on recruitment

352

from seed banks (van der Valk and Davis, 1978). Numerous species persist in the seed bank until suitable germination conditions occur, at which time they are recruited to the community. Mangrove ecosystems are often described as undergoing cyclic succession, with large-scale disturbances such as hurricanes, sedimentation, or sea-level change periodically resetting the system to an earlier state (Lugo, 1980). Some successional processes in wetlands may be controlled by human activity; for instance, ﬁre suppression and/or stabilization of hydrologic regimes by humans have resulted in species changes in freshwater marshes in the Everglades (Ewel, 1986) and in European salt marshes (ter Heerdt and Drost, 1994; De Jong and Van der Pluijm, 1994) and swamps (Segerstrom et al., 1994). Ecosystem responses Effects of disturbance on ecosystem-level processes such as net primary production and nutrient cycling have been examined in only a few wetlands, so generalizations are difﬁcult to make. However, one can make some predictions based on how different types of disturbance regimes affect organic matter and nutrients in wetlands. Physical disturbances in which biomass is killed but not consumed (e.g., during a windstorm that defoliates a forested wetland) may stimulate decomposition or secondary production of detritivores by fresh litter input. For example, estimates of dead biomass generated in mangrove forests in southern Florida by Hurricane Andrew approached 150 metric tons per hectare (Smith et al., 1994). The same hurricane caused an increase in leaf-litter ﬂux in a Louisiana forested wetland equivalent to 30–41% of the normal annual litter-fall (Rybczyk et al., 1995). Such sudden changes in organic-matter pools may have impacts on ecosystem processes different from those of low-disturbance regimes in which transformations are more gradual. Our survey of disturbance effects in wetlands showed that a decrease in standing plant biomass after disturbance occurred in most cases (Table 13.4). However, reductions in plant biomass depended on study duration, since longer-term experiments allowed more time for vegetative recovery. Also, differences among studies in disturbance intensity affected the amount of biomass initially damaged or removed as well as the amount of regrowth. There were more reports of decreased biomass in salt and freshwater marshes than in brackish marshes and in response to

Karen L. McKEE and Andrew H. BALDWIN

grazing than to other types of disturbance (Table 13.4), which may reﬂect differences in species assemblages among wetland types, or relative responses by species to particular disturbance regimes. Generalized effects of disturbance on net primary production in wetlands are more difﬁcult to state, since most of the productivity studies in our survey were conducted in saline marshes (Table 13.4). Given this limitation, the results showed that production was decreased in response to disturbance in some cases. Two or more disturbance agents caused a decrease in primary production in all reported cases. However, production remained unchanged or was increased in 59% of the cases. Differences in disturbance regimes may explain this pattern; for instance, physical disturbance generally caused no change in production, whereas grazing effects were mixed. In general, enhancement of primary productivity by disturbance could occur because of: 1) increased nutrient availability; 2) reduction in self-shading by biomass removal; 3) removal of older, less productive tissue; and/or 4) changes in competitive interactions favoring more productive species. Physical disturbances that redistribute organic matter could alter several factors affecting primary production; these effects may include shade reduction and removal of less productive tissues. The pools of nutrients and their availability to producers in wetlands could also be impacted by physical disturbances such as windstorms. Nutrient contents of downed leaves (e.g., in a Louisiana forest hit by a hurricane: Rybczyk et al., 1995) would be higher than normal senescent ones, if removal occurred before nutrient retranslocation. Release of nutrients from litter might occur relatively quickly, depending on the structural and chemical composition of the litter and changes in environmental conditions (for instance, as a result of initial leaching or increased decay rates caused by higher temperatures after canopy removal). Factors such as moisture and oxygen may still limit decomposition, however, so that nutrient releases may not remain high after the rapidly leached components are broken down. Also, a large portion of the litter may be refractory (e.g., wood in forested wetlands), which not only may have a lower nutrient content than leaves, but may decay more slowly. Thus, even though both carbon and nutrients are redistributed within the system after a physical disturbance, the bulk of the material may not be immediately transformed. Disturbances such as ﬁre not only consume plant

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

biomass and increase light penetration, but also rapidly transform nutrients. Although some nutrient elements are volatilized in the ﬁre, others are deposited in the ash. Burning may also result in greater spatial variability of phosphate phosphorus and nitrate nitrogen in wetlands (Wilbur and Christensen, 1983). Effects of winter ﬁre on nutrient pools in an irregularly ﬂooded marsh in Mississippi were studied by Faulkner and de la Cruz (1982) in black needlerush (Juncus roemerianus) and big cordgrass (Spartina cynosuroides) communities. Winter ﬁre caused a minimal input of calcium, magnesium, phosphorus, and potassium to soils, but 90% of the nitrogen and 50% of the potassium in plant material was lost. Although increased primary production in these brackish-marsh communities following marsh burning was attributed to temporary increases in nutrient availability (Hackney and de la Cruz, 1981; Faulkner and de la Cruz, 1982), it could also have been a response to removal of shaded tissues and increased light levels. Fire may not always result in nutrient increases sufﬁcient to stimulate plant production. Burning of sawgrass communities caused large releases of nutrients, but most were removed within a few hours (Forthman, 1973). Similarly, Hoffpauir (1961) found that increases in concentration of potassium and sodium in the water of a marsh in Louisiana following burning were diminished after 49 days as a result of tidal action, rainfall, and regrowth. Grazing, which can enhance primary productivity if nutrients are returned via feces and urine, would likely cause effects intermediate to that of physical disturbance and ﬁre. Hik and Jefferies (1990) found that grazing early in the growing season increased net above-ground primary production (NAPP) of Puccinellia phryganodes, but only in plots with low to moderate grazing. This was attributed to the fertilizer effect of goose feces, because clipping alone decreased growth of the sward. Other grazing experiments have shown similar results (Cargill and Jefferies, 1984). Higher grazing intensity may cause a decrease in net aboveground primary production (e.g., in salt marshes: Hik and Jefferies, 1990; Andresen et al., 1990) or aboveground biomass (e.g., in Louisiana brackish marshes: Taylor and Grace, 1995) and in extreme cases can lead to permanent loss of wetland (Miller et al., 1996). Effects of disturbance on nutrient cycling in wetlands involve a number of processes, including direct addition of nutrients to the soil (e.g., in feces and urine of grazers and deposition of sediment or

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ash from ﬁre), net mineralization of organic matter killed by the disturbance (e.g., standing dead plants, fallen plant fragments, or fecal droppings), nitrogen ﬁxation by Cyanobacteria on the sediment surface, and nitriﬁcation–denitriﬁcation processes in surface and sub-surface sediments. Bazely and Jefferies (1989) examined the effects of lesser snow geese (Anser caerulescens) on the nitrogen budget of a salt marsh. Nitrogen ﬁxation rates were higher in grazed than in ungrazed plots, because the lack of litter accumulation in the former allowed growth of cyanobacterial mats. Grazing also increased nitrogen availability for forage species (Carex subspathacea and Puccinellia phryganodes) through fecal input and acceleration of physical breakdown and decomposition of organic matter. Kemp et al. (1990) similarly showed that grazing by a snail (Littorina irrorata) affected nitrogen turnover in a salt marsh by accelerating breakdown and decomposition of the dominant grass (Spartina alterniﬂora). In both examples, the effect of grazing was to increase the rate of nutrient transformation and to alter nutrient pools. Disturbance by pollutants could inhibit or stimulate nutrient transformations depending on toxicity to micro-organisms or provision of carbon substrates such as petroleum hydrocarbons (Grant and Payne, 1982; Thomson and Webb, 1984; Lin and Mendelssohn, 1996). The long-term consequences of disturbances that remove organic matter and/or nutrients from wetlands (e.g., frequent ﬁres, livestock grazing, logging, or hay removal) depend on the speciﬁc regime and other local conditions. Intense ﬁres that damage live roots or remove layers of peat can have strong negative effects on plant regrowth in some wetlands (Lynch, 1941; Nyman and Chabreck, 1995), but in others revegetation may occur relatively quickly (although species composition is changed). Frequent grazing by animals allows some return of nutrients through fecal deposition, but practices such as logging and hay removal do not. Extensive and frequent burning or removal of plant biomass in wetlands can also negatively impact wildlife dependent on the vegetation for food, cover, and nesting materials (Bray, 1984; Burdick et al., 1989), although the consequences for higher trophic interactions is not known. Removal or consumption of biomass would also reduce litter production with consequent effects on detritivores and their predators, for instance in cattlegrazed salt marshes. Changes in organic-matter export could potentially affect detrital-based food webs and secondary production in adjacent coastal (Turner, 1977;

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Twilley, 1985, 1988; Deegan et al., 1990) and inland (Turner, 1988a,b) water bodies. Large and/or continual losses of material that would otherwise contribute to vertical accretion could ultimately affect wetland stability or alter succession in some types of wetlands (e.g., those undergoing rapid subsidence or ﬂuctuating water levels). Landscape responses Studies using a landscape approach to assess wetlands can be found in the literature, but few report landscapelevel responses to disturbance (Bedford and Preston, 1988; Klopate, 1988; Burdick et al., 1989; Childers and Gosselink, 1990; Gosselink, 1990; Brinson, 1991; Calvo et al., 1992; Childers et al., 1994; Bedford, 1996). Examination of disturbance effects on wetlands at a landscape level is important, since impacts of many types of disturbance can only be appreciated from a large-scale perspective and because wetlands are an integral part of the landscape, exchanging materials with other landscape components (Klopate, 1988; Pearson, 1994). The landscape-level perspective differs from that of the ecosystem because it considers the wetland as a component in a larger, heterogeneous landscape. Geomorphologic processes of sedimentation, erosion, wind and wave action, glaciation, and uplift generate somewhat distinct physiographic units such as coastal plains, river deltas, and basins. Within each of these units are speciﬁc types of wetlands and different disturbance regimes, such as hurricanes, channel meandering, or waterfowl grazing. A landscape perspective allows assessment of patterns that reﬂect the wetland’s geomorphologic origins as well as disturbances in adjacent ecosystems; for instance, roadways or canals may alter water movement or sediment into or out of the wetland (Bedford, 1996). Examination of disturbance in wetlands can begin at a landscape level by considering landscape features that differ among wetland types. Boundaries between wetland and non-wetland (or between different types of wetland), proportion of edge to area, and spatial continua are landscape features that differ among wetland types and that may inﬂuence effects of disturbance and response of the biota. The presence of corridors, such as river channels crossing or connecting wetland ecosystems, determine movement of water, nutrients, pollutants, and biota. Landscape-level movement of seeds of wetland species along riverine corridors, for example, can be used to assess effects of stream

Karen L. McKEE and Andrew H. BALDWIN

diversion or other disturbance regimes on the species composition of a downstream wetland community. The landscape approach is also useful when dealing with large-scale disturbances. Knowledge of how the disturbance affected adjacent systems that are linked to the wetland is helpful in predicting responses and recovery rates. A hurricane causing widespread disturbance to several different ecosystems within a landscape is a case in point. Hurricane Andrew damaged a variety of ecosystems in southern Florida: underwater reefs, seagrass beds, freshwater marshes, and mangrove, pine, cypress, and tropical hardwood forests (Pimm et al., 1994). The hurricane’s swathe was about 100 km long, and vegetation was partially or completely defoliated across a band 50 km wide. The effects of the hurricane and particularly the recovery of the wetland ecosystems depend in part on their relative position in the landscape. Landscape features such as spatial conﬁguration and areal extent of the wetland, as well as characteristics of adjacent communities, inﬂuence rates and patterns of recovery. Some wetland systems are closer to urban or agricultural centers than others, and differences in human activities affect resistance as well as resilience of the wetland communities to large-scale natural disturbance. Study of human effects on wetlands is particularly amenable to the landscape approach. Human activities in and near wetlands can be readily identiﬁed and placed into a landscape perspective. Human habitations are typically not randomly distributed, but concentrated in certain landscapes – along coastlines, lake margins, and waterways. Agricultural and industrial activities are similarly concentrated. Other human disturbances in wetlands (e.g., marsh burning, pollution, or logging) may also be concentrated in speciﬁc areas. Their landscape-level patterns may differ, however. In a natural wetland landscape, the horizontal and vertical structure is derived from geomorphology, and natural disturbances often produce irregular patterns (Fig. 13.2A) and gradual transitions between communities, although ecotones separating wetland from non-wetland may be abrupt. As human intervention increases in the wetland landscape, stronger contrasts such as sharp, linear boundaries appear (Figs. 13.1B and 13.2B). The managed marsh, for example, might retain much of its natural landscape features, but some portion may be linearized. As human activities increase in the vicinity of or within the wetland, more straight lines and geometric patterns appear, for instance, with adjacent agricultural or urban activities. The endpoint

DISTURBANCE REGIMES IN NORTH AMERICAN WETLANDS

along this gradient is a fragmented or isolated wetland situated in the middle of an urban or industrial complex. Thus, the level of human modiﬁcation on wetlands can be assessed on a landscape scale as an increase in linearization, fragmentation, or some other temporal–spatial feature. Wetland characteristics affecting response to disturbance Vegetation in wetlands can inﬂuence disturbance regimes by altering the location, amount, and form of organic matter. For example, ﬁres in some Florida swamp systems are sustained by the continuing buildup of dry litter and peat, which serve as fuel sources (Ewel, 1990). At the same time, certain swamp species are intolerant of ﬁre, while others such as pond pine (Pinus serotina), titi (Cliftonia monophylla), and swamp cyrilla (Cyrilla racemiﬂora) depend on ﬁre for regeneration (Ewel, 1990), suggesting that the ﬁre regime can determine the composition of vegetation in these swamps. The relationship between wetland characteristics and disturbance regime is further complicated by propagule availability and environmental conditions such as water level and precipitation, which can affect both the composition of the vegetation and the severity and frequency of ﬁre. The effect of environmental conditions on the disturbance regime may be illustrated in coastal marshes dominated by rhizomatous graminoids, where water level can dramatically inﬂuence the effect of ﬁre on vegetation. If the water level is above the soil surface, the plant parts above the water are burned off, but the rhizomes survive and the vegetation quickly recovers (reviewed by Nyman and Chabreck, 1995). If the water level is below the soil surface, however, the rhizomes may be killed by high surface temperatures. In some cases peat itself can ignite, consuming soil and rhizomes. Grazing is another example of a disturbance regime that is often determined by the type of vegetation. In brackish marshes of coastal Louisiana, the preferred food of the muskrat is Olney’s three-square (Scirpus americanus) (Chabreck, 1976). Other plant species, such as Spartina patens, are less attractive to these herbivores. Similarly, many insect grazers are hostspeciﬁc, feeding on particular taxonomic groups of plants (e.g., specialized insects on mangroves) (Feller, 1995). These observations suggest that herbivore pressure will be higher in areas where preferred

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plant species are abundant. Grazer abundance is also inﬂuenced by environmental conditions such as salinity and temperature, which can determine abundance and activity of animals directly by exerting physical and physiological stresses, as well as through effects on plant species composition. The same type of disturbance can have very different effects depending on the type of wetland in which it occurs. Hurricane-force winds may cause relatively little physical damage to herbaceous salt-marsh vegetation (Guntenspergen et al., 1995), but severe damage to trees in mangrove forests (Smith et al., 1994; Baldwin et al., 1995). Similarly, anthropogenic input of nutrients may have little effect in wetlands such as ﬂoodplain forests, which naturally receive an abundance of nutrients in surface or ground water (Mitsch and Gosselink, 1993). However, addition of nutrients to low-nutrient systems such as sawgrass savannas in the Everglades of southern Florida may result in invasion by species normally occurring in higher-nutrient systems (e.g., Typha sp.: Koch and Reddy, 1992). These examples illustrate the potential complexity of relationships between disturbance regime, vegetation, and environmental conditions in wetlands. In practice, effects of these factors may be reciprocal, making it difﬁcult (if not impossible) to identify causal relationships among them. WETLAND RECOVERY FROM DISTURBANCE

Type, intensity, scale, pattern, and frequency of disturbance interact with attributes of the biota to determine recovery rates in wetlands. In general, the ability of a system to recover is determined by time span between successive disturbances relative to the time required for the biota to regenerate. Pickett and White (1985b) also emphasized the role of the resource base in determining rates of regrowth after disturbances. Wetland systems vary tremendously in availability of resources. Some wetlands, such as northern bogs, are nutrient-poor (Johnson, 1985), whereas other systems that receive large amounts of alluvial sediment are relatively resource-rich. If these systems were exposed to the same type and magnitude of disturbance, they might exhibit different rates of recovery because of differences in availability of resources to support regrowth. Thus, recovery may be relatively slower in resource-limited and/or previously-stressed wetland habitats compared to more productive ones. However,

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variation in other factors such as species composition makes comparisons across wetland types difﬁcult. A study reporting the responses of coastal marshes near Atchafalaya Bay in Louisiana to the impact of Hurricane Andrew provides an excellent illustration of how recovery of plant communities depends on the type of disturbance as well as the initial condition of the wetland (Guntenspergen et al., 1995). At a landscape scale, impact of the hurricane on the vegetation depended primarily on marsh type. Oligohaline and freshwater marshes were greatly affected by salt burn, whereas no evidence of salt burn was found in salt or brackish marshes. The condition of the marsh prior to the hurricane (e.g., healthy vs. deteriorating, or grounded vs. ﬂoating) determined whether the landscape was fragmented, compressed, or unaltered. This study also demonstrated how a single event can generate different types of disturbance in close proximity to each other (Fig. 13.2A) – which, in turn, determined vegetative recovery. The areas damaged by salt burn began resprouting within six weeks after the hurricane. In contrast, scouring, deep sediment burial, and wrack deposition caused the greatest amount of disturbance to the vegetation, and most of the areas affected showed no signs of recovery during the ﬁrst months after the hurricane. However, after deep sediment deposition, the vegetation recovered relatively faster and showed rapid succession. Vegetative recovery was much slower in the sites with wrack deposition, and some areas with thick deposits (up to 20 cm deep) were still devoid of plants 6 to 8 months after the hurricane. In the scoured areas, there was little possibility for recovery, and permanent loss of wetland was predicted. Individual species may exhibit different abilities to recover from disturbance. Disturbance by sediment deposition into a salt marsh at Bolinas Lagoon in northern California was studied to examine the ability of marsh plants to recover (Allison, 1995). Distichlis spicata and Salicornia virginica were most able to recover. Frankenia salina and Jaumea carnosa only recovered if disturbance occurred early in the growing season. Recovery was by vegetative growth; seedling establishment was rare and did not inﬂuence recovery signiﬁcantly. Lasting changes in vegetative composition and cover occurred following sediment deposition as a result of elimination or reduction in cover of sensitive species and changes in the substrate reducing seedling recruitment. These are but a few examples of how different types of disturbance and wetland characteristics determine

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rates and patterns of recovery. Much more work is needed in a variety of wetlands before generalizations can be made about what factors are most important in controlling rates and patterns of recovery from disturbance.

CONCLUSIONS

Considering the highly dynamic nature of wetlands, it is surprising that so few studies have been conducted to examine speciﬁcally the role of disturbance in wetland communities. However, if we are to identify unifying principles of organization in wetlands, then one must carefully consider all processes that potentially contribute to variation in heterogeneity within and across wetland types. Although some wetlands may experience little disturbance, other systems may be strongly affected by disturbance, either through natural or anthropogenic agents. Comparisons of systems with different kinds, rates, and magnitudes of disturbance may lead to a greater understanding of patterns and processes in wetlands. Even those wetland types that appear to be relatively unaffected by natural disturbance may in fact be subject to subtle, lowlevel disturbance that has as great an impact as more dramatic events such as ﬁres or hurricanes. An understanding of wetland structure and function thus requires consideration and experimental examination of the role of disturbance. Also, wetland systems are ever vulnerable to human-induced disturbance. From a practical standpoint, an understanding of disturbance processes could lead to stronger management practices that minimize wetland loss and/or unwanted changes in wetland structure and function.

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Chapter 14

MINING John A. COOKE

INTRODUCTION

Table 14.1 Minerals produced by the mining industry

The products of the mining industries (Table 14.1) are the mineral resources which are essential to modern civilisation and on which the generation of wealth often depends. Mineral resources provide energy, construction materials, transport, and many items which improve the quality of life and deﬁne the norms for standards of living in developed countries. Even with increasing usage of oil and natural gas, coal, the largest mining industry with 4.5 billion (109 ) t yr−1 of coal mined throughout the world, presently fuels 40% of world electricity production (Buchanan and Brenkley, 1994). If usage is expressed in terms of the range and quantity of products, the consumption of the major non-ferrous metals is considerable in applications such as batteries (cadmium, lead, nickel), electrical conductors (aluminium, copper), car body shells, domestic appliances and structural steel (zinc), other alloys (nickel), and rooﬁng (copper, lead, zinc) (Barbour, 1994). Mining has been undertaken since the beginning of civilization. In eastern Britain and elsewhere there is evidence of Neolithic ﬂint mines before the age of metal tools; copper and zinc have been mined since the Bronze Age, and 5000 years ago in Egypt and the Sudan gold was being mined; in early Roman times lead was obtained from Greece and in the later Christian era from Britain, Sardinia, and Spain. Following the decline of the Roman Empire, many of the mining skills were lost except in Germany, which led the revival of mining in the Middle Ages. For example, German mining expertise helped European countries to meet the demand for lead for rooﬁng and piping during the great periods of building of castles and churches in the 12th century. However, it was the growth in industrialization in the early

Type

Main minerals

Mineral fuels

coal, lignite, oil-shale, peat, uranium

Construction materials

gravel, sand, slate, stone

Metals precious

gold, silver, platinum group

ferrous

iron, manganese

non-ferrous

aluminium, cadmium, chromium, copper, lead, nickel, tin, titanium, zinc

Other minerals

asbestos, barytes, china clay, diamonds, ﬂuorspar, graphite, gypsum, perlite, phosphate, potash, salt, vermiculite

Table 14.2 World production of some metals (103 t yr−1 ) over 150 years 1 Year

Aluminium

1835

−2

Copper 33

Lead

Zinc

91

14

1900

88

495

871

400

1950

8418

2488

1686

2210

1984

78 800

8230

3350

6350

1 2

Data from Mining Annual Review (1985). Insigniﬁcant production in 1835.

19th century which increased the demand for most minerals. For example, the mining and production of copper, lead, and zinc have increased 250, 37, and 450 times, respectively, over the 150 years since 1835 (Table 14.2). Aluminium is an example of a metal not used historically; but the mining of its ore, bauxite,

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and its production has increased considerably during the 20th century (Table 14.2). The exponential increase in production of many minerals has been accompanied by shifts in the main geographic centres of mining and its globalization from its beginnings in a few countries to now, when over 150 countries are involved in mining in a substantial way (Mining Annual Review, 1995). The history of copper production (Mining Annual Review, 1985; Table 14.2) illustrates this point. In the 19th century Britain was the largest producer of copper (15 000 t yr−1 in 1835 – about half the world production), originally using the ore mined in Cornwall and Devon but later with imported ore. Copper from the United States then took over the world market, rising from 10 000 t yr−1 in 1867 to being the main producer with 250 000 t yr−1 in 1900. By the mid-1950s Zambia and Chile were major contributors. Chile is now the world’s largest producer, having increased its mining production from 1.3 106 t yr−1 in 1984 to 2.2 106 t yr−1 in 1994 (Mining Annual Review, 1995). Mining covers a large range of activities which include primary phases (extraction and quarrying) and secondary phases (milling, processing, and waste disposal). The direct impacts of mining are largely local in extent, but such activities can cause extreme disturbance to land surfaces with the direct removal of all previous soils, plants, and animals or their burial beneath waste-disposal facilities. On a global scale, the area of land directly degraded by mining industries is less signiﬁcant than the other forms of disturbance such as overgrazing, deforestation, agriculture, overexploitation for fuelwood, and other industrial uses (Daily, 1995). However, within an individual country the area of land disturbed by mining can be large. For example, China has 2 106 ha disturbed by mining out of a total area of 930 106 ha (Guo et al., 1989), and in South Africa the area of land directly affected by mining has been estimated at 1 106 ha, which is 1% of the total area (Wells et al., 1992). Mining can also lead to indirect disturbance of the surrounding land and environment through the alteration of drainage patterns and the pollution of surface waters, groundwaters, and the atmosphere. These may have much greater spatial impact than direct disturbances and are often difﬁcult to assess and quantify. Land disturbance caused by mining, like that caused naturally by glaciation or volcanic activity, leads to new land surfaces and opportunities for the development

John A. COOKE

of ecosystems through succession (Bradshaw, 1983). This succession can be facilitated and directed by reclamation efforts (Bradshaw, 1990a). This chapter discusses the nature of disturbance of land by mining, and some of the post-mining ecosystems which have developed both through natural succession and following reclamation designed to promote such development.

MINING AND DISTURBANCE

Mining methods and land disturbance There are many methods of mining involving either underground or surface extraction. Which method is used often depends on the nature of a particular deposit and not just on the nature of the mineral itself. Table 14.3 summarizes the main methods of mining and the nature of the associated land disturbances. Mining, in the human consciousness, is associated with tunnelling underground; in the pre-Industrial age, this was in relatively shallow workings. Shallow underground mining has largely been superseded by the development of massive machinery involved in strip mining from the surface. However, there are a considerable number of abandoned, small-scale, shallow workings and associated spoil heaps in most of the main centres of mining throughout the world. The history of the underground mining industries has involved the development of deeper and deeper mines, and this has led to increasing land disturbance, mainly because of the need to dispose of waste rock and spoil on the surface. By the beginning of the 18th century, large steam engines and pumps were developed which could drain deep workings and mechanically ventilate deeper and deeper shafts and levels. For example, in 1763, a shaft of 180 m was sunk at Walker-on-Tyne in northeastern England, but 100 years later shafts of 600 m were being used in this and other coal mines (Galloway, 1882). By 1913, when coal mining was at its peak in Britain, a small and densely populated country, there were 2789 working deep coal mines (Duckham, 1969). In South Africa, gold is mined from a “reef ” of quartzite, at even greater depths and in an arc of 47 km, on the Witwatersrand near Johannesburg. There are currently 40 deep mines in this area; one of these, Western Deep, has a shaft to a depth of over 3.35 km, which is soon to be increased to over 4.1 km. (South

MINING

367

Table 14.3 Main types of mining and land disturbance Mining method

Brief description

Main land disturbance

Shallow underground mining; e.g. coal, metal ores

Seams up to 60 m deep; using bord and pillar, levels and adits in valleys, bell pits, etc.

Surface spoil heaps; subsidence and collapsed old workings often left derelict

Deep underground mining; e.g. coal, gold

Seams deep underground, accessed through shafts

Subsidence; surface waste disposal – spoil heaps, tailings and slurry lagoons usually remain

Strip-mining (opencast), e.g. coal, mineral ores, diamonds

Horizontal or sloping seams usually up to 60 m below surface, taken out from surface

Removal of vegetation and stockpiles of overburden-rock, topsoil, and subsoil, temporary if using progressive backﬁlling; mineral processing and waste-disposal facilities

Dredge mining, e.g. heavy minerals such as ilmenite and rutile

Alluvial and mineral deposits throughout bulk of mined materials; mining ponds created with ﬂoating dredger and concentrator

Removal of vegetation and stockpiling of topsoil and tailings from concentrator, usually progressively backﬁlled

Quarrying, e.g. hard rock, sand, and gravel

Rock and aggregate taken out from surface working; most materials sorted and processed on site and removed

Little left to backﬁll; steep faces and quarry ﬂoor left

Open-pit mining, e.g. metal ores, china clay

Ore body near surface usually steeply dipping seam or pipe; ore taken out by blasting or hydraulically

Little left to backﬁll; steep faces and pit ﬂoor left; also rock dumps and waste-disposal facilities such as tailings lagoons remain

African Mining, 1996). Wastes associated with these mines have been dumped on the surface since 1886, usually close to main urban centres including the city of Johannesburg, and cover about 8000 ha (Thatcher, 1979). Coal production is currently expanding at over 2% a year, world recoverable reserves being of the order of 850–1000 109 t (Buchanan and Brenkley, 1994). Much of this increase has been achieved by increases in surface-mining, particularly using strip-mining techniques. Strip mining involves the removal of all the surface layers above the mineral, including topsoil, overburden and rock, and the consequent disturbance of aquifers and drainage patterns. The United States currently strip-mines about 60% of the 935 106 t yr−1 of coal it produces (Mining Annual Review, 1995). In some states such as Ohio, strip-mining for coal has a long history, large-scale operations having been undertaken since the advent of the power shovel in 1914 (Riley, 1957). Since that time, large dragline operations have been developed which are able to remove over 300 t in one bucket load. By 1965 it was reported that 112 000 ha of land in Ohio had been disturbed by surface mining, with a ﬁgure for the United States

as a whole of > 1.25 106 ha (Thirgood, 1978). Of this total, 64% of the strip-mined land was still in need of reclamation. In fact, in the United States there was little incentive to reclaim strip-mined lands; vast areas were left abandoned, with topsoil and overburden mixed and cuts not ﬁlled, and with pyrite-contaminated spoil generating acid drainage waters (Bradshaw and Chadwick, 1980). The scale of the potential land disturbance by strip mining in the United States can be appreciated from estimates of strippable coal reserves, which were given as 4 106 ha in 1981 (United States National Research Council, 1981). In other types of surface mining (Table 14.3), such as quarrying for aggregates or open-pit mining, the quarry or pit itself may be left open, and pose problems for reclamation. A distinction needs to be made between the changes in scale between old and modern operations. Rock and stone have been quarried for centuries; but in earlier years the quarries were small, serving only local construction needs. The development of cement and concrete by the Romans changed the pattern of demand to a universal one for limestone, clay, and aggregates (Bradshaw and Chadwick, 1980). This demand, together with the use of machinery to

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Fig. 14.1. Waste production in non-ferrous metal mining, milling, and processing (based on Barbour, 1994).

process the quarried materials on site, has led to very large modern quarries (faces over 100 m deep with a working life of >50 yr). Similarly, the scale of open-pit workings has reached massive proportions, as demonstrated by the Kennecott Copper Mine in Bingham (Utah, U.S.A) with an open pit 3 km wide and 800 m deep (Williamson et al., 1982). Mining wastes: direct and indirect disturbances In many of the forms of surface and deep mining (Table 14.3) it is the production and disposal of waste which can cause the most extensive and long-lasting disturbance to land. For example, the legacy from a declining industry using deep mining for coal in Britain was that, in 1972, the National Coal Board owned 2000 spoil heaps containing 2 109 t of mining waste

on an area of 15 000 ha, with about an equal number of spoil heaps being associated with other coal-mining companies and operations (Thomson and Rodin, 1972; Glover, 1978). Much of this colliery spoil had become extremely acidic over the years, with the oxidation of pyrite releasing sulphuric acid into the surface layers (Palmer, 1978); thus, even though many of the heaps had been abandoned for many years, there had been little natural colonization by vegetation. The relationships between the primary and secondary phases of mining and waste production are shown for non-ferrous metal production in Fig. 14.1. The disposal of rock and overburden, the construction of impoundments (dams) for the ﬁne tailings (<0.1 mm) produced from the milling operations, and to a lesser extent the disposal of slag from the smelting and reﬁning stages, can involve large areas of local

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land. Whereas the ore quality may be very high for iron and manganese (concentrations of 60–70%), nonferrous metal ores are low in metal and, consequently, generate large volumes of waste. Gold, for example, occurs in an extremely low ratio (a few grams per tonne) to the associated rock which must be mined. However, the situation is worsening for many metals, in that their production is increasingly from lowergrade ores and the volumes of waste are escalating. For example, the average ore grade for copper, has decreased from 4% in 1900 to 0.5% in 1975, with a considerable increase in the tailings produced, from, approximately, 17 to 290 106 t yr−1 worldwide over the same period (Williamson et al., 1982). Although many tailings dams are relatively small in size – typically between 5 and 20 ha behind a wall up to 30 m in height – they also may be extremely large structures such as one in Utah (U.S.A.) which has an impoundment area of 2000 ha (Williamson et al., 1982). This is of great concern because the tailings produced often have high levels of potentially toxic residual chemicals or acid-generating sulphides, or may be radiobiological hazards. The toxic elements most commonly found world-wide are similar to the list given for such sites in the United States: arsenic (As), cadmium (Cd), chromium (Cr), copper (Cu), lead (Pb), Mercury (Hg), nickel (Ni), selenium (Se), silver (Ag), and zinc (Zn) (Pierzynski et al., 1994). Extreme acidity may also occur when the tailings contain signiﬁcant amounts of pyrite (iron disulphide), and this acidity can increase the solubility and toxicity of many metals. Uranium ore is extensively mined in Australia, Canada, South Africa, and the United States, and the tailings represent radiobiological as well as toxic hazards (Ritcey, 1989). The low concentrations of uranium in ore (about 0.05%) mean that large volumes of tailings are produced (106 t y−1 from a single mine is common). Radioactive elements such as thorium and radium (and their breakdown products) occur in the tailings. Radium gives rise to radon gas, which may reach high concentrations in the air in the vicinity of the waste disposal site. Because of the very long halflives, these radioactive hazards can remain essentially undiminished for 1000 years without treatment for mitigation (Ritcey, 1989). The production of mining waste can lead to impacts and disturbance to other parts of the environment. Although less common than impoundments on land, discharge of tailings can occur directly into rivers

369

and the sea. Ellis (1988) has provided examples from eleven countries of coastal mines which undertake such disposal and some of the consequences in the ecosystems affected. Other liquid efﬂuents, usually treated before disposal, are generated from the milling of metal ores (Fig. 14.1). They include potential pollutants from: ﬂotation processes (fatty acids, oils, propylene glycol, tannins, xanthates, etc.); chemical treatments from processing of bauxite (caustic soda), gold (cyanides) and uranium (sulphuric acid); and supernatants and recycling water from the disposal of tailings slurry (Barbour, 1994). Acid mine drainage is perhaps the most serious impact mining has on aquatic ecosystems. It is very common, being associated with the mining of coal, metals, and other minerals which occur as sulphides or with sulphide gangue materials. Surface leaching from the physical workings, rock dumps, spoil heaps and tailings, occurs and such drainage can last almost in perpetuity without treatment (Ferguson and Erickson, 1988). The cumulative length of streams and rivers world-wide affected by acid mine drainage is considerable, though difﬁcult to estimate. Contamination of groundwater can occur from storage of mining waste on land. This can seriously affect water extraction from wells and aquifers. Even with the most elaborate interception and isolation technologies to prevent groundwater coming into contact with wastes, most such efforts will fail over the long term (Sengupta, 1993). Air pollution from the processing and smelting of ore has been of major concern in some of the main mining centres of the world. The Sudbury area of Ontario (Canada), with rich deposits of nickel and copper, had 13 mines, 6 concentrators, 4 smelters, and two iron-ore recovery plants in the 1970s. This area has a well-documented history of air pollution from the wet and dry deposition of sulphur and metals, which have had considerable impact on the surrounding ecosystems in an area of >1000 km2 (Hutchinson and Whitby, 1974). A more recent air-pollution concern may have considerable global impact in the future. It is the emission of methane, an important greenhouse gas, from the activities of underground and surface mining themselves. These emissions are geographically spread around the world, and have been estimated at 45 1012 g yr−1 (Williams and Mitchell, 1994).

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Table 14.4 Severity of the main factors which inﬂuence plant establishment on some wastes and substrates of mining origin 1 Waste type

Adverse texture/stucture 2

Low nutrients

Extreme pH

Toxicity

High salinity

Coal spoil

2

2

0–2

0–1

0–1

Oil-shale

1–2

2

1–2

0

0–1

Iron-ore mining

1

1

0

0

0–1

Heavy metal wastes

2

2

2

1–2

0–1

Gold wastes

2

2

2

0–2

0–2

Bauxite (red mud)

2

2

2

2

2

Acid rocks

2

1

1

0

0

Calcareous rocks

2

2

1

0

0

Sand and gravel

1

1

0

0

0

China clay wastes

2

2

1

0

0

Strip-mining (coal)

1–2

1–2

0–2

0

0–1

1

1–2

0

0

0–1

Coastal sands 1

Adapted from Bradshaw and Chadwick (1980). Key to importance of limiting factors (difference in severity due to variation in materials and situations): 0, negligible; 1, moderate; 2, severe. 2

ECOSYSTEMS OF MINED LAND

Succession and reclamation Mining is a temporary land-use. It is not sustainable at any one place because the mineral deposit is ﬁnite and eventually exhausted. The land surface is changed by mining activity, and the disturbance may persist for a long time. Reclamation is the process whereby the land surface is returned to some form of beneﬁcial use. The terms restoration, rehabilitation, and replacement represent the goals of the reclamation process to achieve the pre-mined state or some other new land use (Bradshaw, 1990b). Restoration, although used more generally in the earlier literature (e.g., Johnson and Bradshaw, 1979), now is sometimes used exclusively to refer to restoration of the original (pre-mining) ecosystem with all its structural and functional aspects. Rehabilitation is the term used for the progression towards the reinstatement of the original ecosystem, and replacement is the creation of an alternative ecosystem (Bradshaw, 1990b). Recovery of mined land occurs when the land is largely left to natural processes after disturbance. In practice, the terms recovery, restoration, rehabilitation, and replacement may all be described as resetting the ecological clock (Cairns, 1991). Despite the lack of unifying theories of succession

(Miles, 1987), it is important to understand the basic processes concerning ecosystem development in the contexts both of abandoned derelict mine sites and of the practice of ecological restoration of mined land. These processes include the mechanisms of plant and animal colonization, the establishment of species populations, the development of ecosystem structure and function, and the limits and time-scales involved. Ecosystems of abandoned mined land In natural succession where there is usually no soil, resource availability and resource demand during colonization and early succession are particularly important (Vitousek and Walker, 1987). Compared to normal soils, mining substrates, which may have been derived from deep in the earth, or are wastes produced from the processing of the minerals or rock surfaces left after extraction, may present extreme challenges to colonization by plants and the formation of any kind of self-sustaining ecosystem (Table 14.4). Mining substrates show considerable variation in their physical and chemical nature, as indicated in Table 14.4. Severe values (Category 2) of the factors mentioned are likely to inhibit natural colonization by most plant species for many years although often a few species (which may be particularly tolerant or have tolerant ecotypes or populations) may form a sparse vegetation cover.

MINING

Physical texture may be very coarse, as in some quarry wastes and coal spoils, or very ﬁne, as in milled tailings. Fine texture without organic matter leads to high bulk densities, extreme compaction, low water inﬁltration rates, and surface waterlogging. Nearly all mine substrates have very low levels of macronutrients (especially nitrogen, phosphorus, and potassium). Low pH is a problem in wastes containing iron pyrite which on weathering will generate sulphuric acid and (if there is no capacity in the waste to neutralize acids) causes pH values of <2.0. Toxicity, especially of aluminium, zinc, and other metals in acidic wastes, can be a signiﬁcant problem for plant growth. High salinity can be caused by natural weathering, acid neutralization by carbonates, additions of chemicals in the milling and concentration processes, and by evaporation from the surface in warm climates. These constraints on plant growth have been discussed extensively by Bradshaw and Chadwick (1980) and Williamson et al. (1982). The next sections describe abandoned sites associated with limestone quarries, metalliferous mine spoils, and strip mining for coal, which have provided opportunities for detailed study into natural colonization and ecosystem development over sufﬁcient time-scales. Limestone quarries Abandoned rock quarries consist of a series of habitats: steep cliffs, terraces, rock screes (blast piles), and ﬂat bottoms which may be wet or dry (Usher, 1979). Acidic rocks such as granites are hard, weather slowly, and have very little ﬁne material. Such quarries are inhospitable places for plant growth and succession. Limestones, however, are softer rocks which can weather, even if it is over a long time period, to produce shallow soils with a coarse physical texture, which can be colonized by plants. The limitations for plant growth in these materials largely concern their low levels of nitrogen, phosphorus, and potassium and the general unavailability of some nutrients at pH values over 7. However, there are many so-called calcicole species which are adapted to such soil conditions, and are represented among the colonizers of abandoned quarry habitats (Bradshaw and Chadwick, 1980). Abandoned limestone quarries have gained much ecological interest and conservation importance in Britain and western Europe and elsewhere, for their species-rich plant and animal communities (especially calcareous grassland) and the occurrence of rare species which may not be found in many “natural”

371

localities (Ratcliffe, 1974; Davis, 1979). This conservation importance is enhanced because of the considerable destruction of much calcareous grassland by agricultural improvement and urbanization (Willems, 1990). Consequently, limestone quarries provide some of the best-studied examples of ecosystem development through natural succession on mined land. In a survey of over 200 abandoned calcareous quarries in Britain (Davis, 1979), a wide range of plant species was found including many rare species in the Orchidaceae such as the bee, ﬂy, man, and musk orchids (Ophrys apifera, Ophrys insectifera, Aceras anthropophorum and Herminium monorchis). Four basic types of plant communities were described: open ground becoming herb-rich open grassland; closed grassland; scrub; and woodland (Davis, 1979). These communities represent a successional progression of many decades of ecosystem development (Usher, 1979). The reasons for slow vegetation development can be divided into the following main categories: low input and retention of seed, low seedling establishment and survival, and biotic checks including grazing and competition (Davis et al., 1985). Some of the factors involved in these categories are discussed below. The surrounding land-use is important as the main seed source (Hodgson, 1989). Relict patches of calcareous species around the quarry may facilitate colonization [e.g., blue moor grass (Sesleria albicans) left around the edges of magnesian limestone quarries, in northeastern England (Richardson et al., 1980)]. Long-distance transport of seed and other propagules is possible – for example, the dust-like seed of the orchid species referred to above, which can be carried in the air for large distances. However, there is a random element determining whether such seed reaches a particular quarry and ﬁnds suitable microsites for germination within it. This probably explains many of the observations that quarries only a few kilometres apart contain different sets of the rarer species (Davis, 1979). The age of the quarry is also important in terms of length of time needed for the arrival of propagules. The presence of a population of a plant species will depend on its ability to germinate, survive, and reproduce. It is essential that suitable substrates and microsites exist. Older abandoned quarries usually have more sites favourable for plant establishment than modern fully-mechanised quarries, because of a more heterogeneous topography and range of substrates, including softer materials such as clays and overburden (Davis, 1979; Humphries, 1980). Plant establishment

372

can be affected by interactions between limiting factors. For example, the lack of a sufﬁcient supply of nutrients can lead to poor root growth and the death of seedlings due to drought (Bradshaw et al., 1982). In most accounts of the plant species involved in the early stages of succession, legumes are particularly signiﬁcant. For example, in Wharram Chalk Quarry in the Yorkshire Wolds (Britain), Lotus corniculatus, Medicago lupulina, and Ononis repens were important members of the pioneer community (Usher and Jefferson, 1990) and in Jamesville Limestone Quarry near Syracuse (New York, U.S.A.) Melilotus alba was one of the principal species colonizing the bare rocky substrates (Klemow and Raynal, 1981). The importance of nitrogen-ﬁxing species is a key factor in any ecosystem development and can be rate-limiting, because these species are only sparsely present in the seed rain, and seedling mortality can be very high. Through the slow accumulation of nitrogen by such species, together with other inputs of nutrients through the weathering of rock and from the atmosphere, skeletal soils develop which can support an open grassland with low-growing forbs and grasses, often with 30–40 species in a square metre. The closure of the grassland sward is accompanied by a decline in species richness when more competitive grasses (e.g., Arrhenatherum elatius, Dactylis glomerata, and Poa pratensis) become established (Usher, 1979; Usher and Jefferson, 1990). The slow process of soil eutrophication may trigger the development of a closed grassland with competitive, faster-growing species, followed by the development of scrub and woodland. Interestingly, in the studies in Wharram Quarry (Usher and Jefferson, 1990), the peak and then decline in plant species richness after the open grassland phase was not followed by a similar trend in the soil animal community. Here, mites of the groups Cryptostigmata and Mesostigmata increased in both number of species (from 9 to 38) and numbers of individuals per square metre; for Collembola the number of individuals increased 25-fold, but the species richness remained stable. Large, modern quarry operations are faced with many difﬁculties in the choice of suitable reclamation strategies. Whereas the quarries described above were generally small, modern operations are both too large to abandon without reclamation, and too big to ﬁll to their original contours (e.g., with domestic household waste). Some current ideas on ecological restoration, developed out of the work on old abandoned quarries,

John A. COOKE

are summarized in Table 14.5 and illustrated in Fig. 14.2 (also see Davis, 1982; Buckley, 1989; Land Use Consultants, 1992). Metalliferous mining wastes Centuries of mining of metalliferous ores and their processing have created many derelict areas throughout the world, with substrates characterized by high concentrations of metals. These areas have often developed plant communities through natural colonization which are distinctive and have been characterized by metal type and soil metal concentration. In some mining areas, such as the 20 000 km2 Shaban (Katanga) Copper Arc in Zaire, there is a mix of metalliferous areas derived from: natural outcropping of metalliferous rock, ancient mining, and modern mining and oreprocessing activities (Brooks and Malaisse, 1985). In this area of south-central Africa, copper and cobalt are the main metals which have been mined, and there is a well-described copper and cobalt ﬂora (Wild, 1978; Brooks and Malaisse, 1985). Figure 14.3 illustrates a generalized transect across an area of mineralization and mine waste and shows the basic changes in vegetation physiognomy with increasing copper and cobalt concentrations in the soil: from open woody savanna, to shrubby steppe often with stunted trees, to plant communities dominated by grasses and forbs with some dwarf shrubs. Where the concentrations of copper range from 0.2% up to as high as 1.5%, the species present become increasingly metal tolerant; about 50 species are recognised as being endemic metallophytes largely restricted to soils with high copper and/or cobalt concentrations. Some of these are shown in Fig. 14.3 (vegetation type C) and include monocotyledons and dicotyledons, annuals and perennials, therophytes, geophytes, chaemaephytes, and hemicryptophytes. The high level of endemism within the metallophyte ﬂora of this area of Africa probably reﬂects the ﬂora’s very long history of speciation in undisturbed metalliferous sites, allowing the evolution of distinct metal-tolerant species. For example, Becium homblei, the copper ﬂower, probably evolved from the non-metallophyte Becium obovatum (Brooks and Malaisse, 1985). In comparison, endemism at the species level is very low in Europe. Perhaps the only example is Viola calaminaria, which is characteristic of the zinc ﬂora of western Europe. The term “metallophyte” can be applied to any plant which shows an association with or restriction to metalliferous soils. The degree of restriction of a

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373

Table 14.5 Promotion of ecosystem restoration in large, modern limestone quarries Task

Possible solutions

Problems

Improvement of site and substrate

Restoration blasting to replicate natural cliff forms; buttresses, benches, and rock screes

Need to improve quarry ﬂoor if left as hard rock

Keeping of ﬁner materials (e.g. 3 mm dust) and returning to scree blast piles

Cost too high and may be impractical

Use of mulches, sewage sludge or similar materials applied to cliff faces, benches and scree

Possible introduction and promotion of weedy species, and addition of too much nitrogen to the system

Hydraulic seeding 1 of scree piles, including wild species, with only low levels of fertilizers

Cost and lack of availability of local wild seed

Use of commercial seed to establish nurse plants to improve conditions for plant growth

Possible inability of wild species to colonise in a realistic time-scale

Planting of native trees on benches and bottom of scree slopes

Non-availability of local provenances

Use of direct seeding of wild species, spreading of seed-rich soil, or transplanting turf from natural areas around edge of quarry

Non-availability of wild seed, soil, or turf in sufﬁcient amounts; nature of the substrate of quarry ﬂoor

Promotion of colonization by indigenous wild species

1 Hydraulic seeding (sometimes called “hydroseeding”) is a technique in which seed and nutrients are sprayed over the ground in the form of an aqueous slurry. The jet of the spray can travel over 60 m [see Bradshaw and Chadwick (1980), pp. 83–87].

Fig. 14.2. Restoration at Tunstead Limestone Quarry, Derbyshire, England. Restoration blasting was used to create a natural landform with rock buttresses and scree piles. A seed mix containing 27 local species of plants (5 grasses and 22 forbs including 2 legumes) was sown hydraulically (see footnote, Table 14.5) onto a 30 cm cover material of crushed limestone, sand, and peat in 1990. For further details see Gunn et al. (1992) (original photograph J.A. Cooke, taken in 1992).

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John A. COOKE

Fig. 14.3. A generalized transect across mineralization and mine tailings in Shaba Province, Zaire (based on Brooks and Malaisse, 1985). (A) Open miombo forest dominated by the tree Brachystegia ﬂoribunda with normal soil copper and cobalt levels below 50 mg kg−1 . (B) Shrubby steppe with Loudetia simplex, Protea goetzeana, and Tristachya helenae; natural mineralization, with copper levels up to 1000 mg kg−1 and cobalt levels of 500 mg kg−1 . (C) Lower-growing vegetation with herbaceous species on mine spoil and tailings with copper levels between 2000 and 15 000 mg kg−1 and cobalt between 500 and 2000 mg kg−1 . Different vegetation types consisting of metallophyte species such as the grasses Eragrostis boehmii and Rendlia cupricola, the sedge Bulbostylis cupricola, the legume Crotalaria cobalticola, and the labiates Becium homblei, Haumaniastrum katangense, and H. robertii. Table 14.6 Metallophytes tolerant of lead or zinc, deﬁned in terms of their distribution on metalliferous mine sites, with European examples 1 Classiﬁcation

European examples

Absolute metallophytes (found on metal-rich sites throughout their distribution)

Viola calaminaria (zinc violet); Thlaspi caerulescens (alpine pennycress)

Local metallophytes (found only Minuartia verna (spring on metal-rich sites for part of their sandwort); Armeria maritima range) (sea thrift); Cochlearia pyrenaica (scurvy grass) Pseudometallophytes (found on Agrostis capillaris (e) both metal-rich and non-metal-rich (common bent) Plantago sites in same region: includes lanceolata (i) (plantain) electives (e) such as tolerant populations of more common species, indifferents (i) which occur regularly, and accidentals which occur sporadically with reduced vigour) 1

Adapted from Antonovics et al. (1971).

species has given rise to a classiﬁcation into absolute, local, and pseudo-metallophytes. Table 14.6 outlines this classiﬁcation and gives examples of European species in each category. In a recent listing of absolute and local metallophyte species in Britain (United Kingdom Department of the Environment, 1994) there

were 13 angiosperms, 2 ferns, 11 bryophytes, and 31 lichens. In Britain, these species are often restricted to metal-rich mine sites, which are now of considerable conservation importance. An important group in any development of vegetation on these sites is the elective pseudometallophytes. These are species which are able to evolve metal-tolerant populations in a few generations through natural selection of seed from neighbouring non-tolerant populations (McNeilly, 1990). Most species which have evolved such populations are outbreeding grasses with the necessary genetic variation. These grass species are usually those of lowfertility soils such as Agrostis spp. and Festuca spp., and are not competitive species on nutrient-rich soils (Ernst, 1988). The role of metals and metal toxicity in determining the characteristics of plant communities which develop on metalliferous mine sites is complex, and few researchers have found simple relationships between metals and vegetation (Smith and Bradshaw, 1979; Baker and Proctor, 1990). In many of the old mine sites, spoil and overburden (often mixed) were discarded near abandoned workings and have remained undisturbed ever since. These sites represent a wide range of edaphic conditions, and there is often extreme heterogeneity within sites at the microscale. This variation depends upon the nature of the country rocks and gangue materials (especially whether

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375

Fig. 14.4. Two abandoned lead mine sites in the Pennines, England. Left: Langdon Head, County Durham showing an entrance to an old level with acidic metalliferous spoil mounds in the background with species-poor Agrostis capillaris–Festuca ovina grassland (original photograph J.A. Cooke). Right: Gratton Dale, Derbyshire with calcareous, high-zinc spoil with a cover of relatively herb-rich grassland with Minuartia verna locally abundant (original photograph A.J.M. Baker).

acidic or calcareous), types and availability of metals, availability of nutrients, and the physical structure (water-holding capacity and surface drought). All these factors are likely to affect plant establishment and interact with the physiological and genetic tolerance or sensitivity of the plant species or populations. In a number of surveys of old lead and zinc mines, even with total soil concentrations of lead and zinc up to maxima of 7.5% and 10%, respectively, the major factor in determining ecosystem development was soil pH and the characteristics associated with it (Johnson et al., 1978; Smith and Bradshaw, 1979; Cooke and Morrey, 1981; Morrey et al., 1988). Acidic wastes characteristically have few species, and over many years only sparse discontinuous cover may

have developed, consisting usually of metal-tolerant pseudometallophytes such as Agrostis capillaris and Festuca ovina (Fig. 14.4). Moorland communities which include heather, Calluna vulgaris, also can occur on such sites in northern temperate regions. In contrast, less acidic and calcareous sites typically have a continuous grass turf with a much greater species richness (25–30 species), including grasses and most of the absolute and local metallophytes such as Minuartia verna, Silene vulgaris ssp. maritima, and Thlaspi caerulescens (Fig. 14.4). Critical threshold levels of pH for the change in species composition appear to be between 5.5 and 6.5. Soil pH is an important determinant of metal mobility in soils, the availability of other elements especially

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John A. COOKE

Table 14.7 Promotion of ecosystem restoration on metalliferous wastes based upon relative toxicity 1 Waste characteristics

Possible solutions

Problems

Low metal content and toxicity; no major acidity or alkalinity problems

Apply lime, fertilizer, and organic matter as required; seed with commercial or wild seed; use turf or thin layer of native soil as an inoculum if available; plant indigenous trees

Probable commitment to long-term maintenance; metals and other trace elements may reach toxic levels in vegetation and animals

Medium metal content and toxicity

Apply lime fertilizer and organic matter as required; sow metal-tolerant commercial varieties or seed of wild metallophytes

Few species/varieties available; cost of collecting wild seed may be very high

High metal content and toxicity

Treat surface with 10–50 cm of innocuous material such as overburden; lime fertilize and seed with indigenous species

Regression may occur through upward movement of soluble toxins

Very high metal content, extreme acidity, toxicity or salinity

Cover surface with 30–100 cm barrier layer such as unmineralized rock and cover with suitable rooting medium

High cost, and may be subject to regression through drought or root penetration through barrier layer

1

From Johnson et al. (1994).

nutrients, and microbial activity. Metal availability is higher in acidic soils, and metal solubilities are reduced at higher pH. In calcareous spoils the increased levels of calcium will also reduce the toxicity of metals, especially lead, through, for example, antagonistic effects between Ca2+ and metal ions (Simon, 1978). The availability of nitrogen and phosphorus is affected by pH. Phosphate may have a role in regulating the toxicity of metals such as lead and zinc, and thus determine vegetation structure, composition, and cover (Alvarez et al., 1974; Smith and Bradshaw, 1979; Morrey et al., 1988; Baker and Proctor, 1990). Other soil factors which are of signiﬁcance are the level of organic matter and the presence of mycorrhizal fungi which are known to be efﬁcient at binding metal ions. Lower levels of activity of soil micro-organisms and higher metal content in leaf litter decrease decomposition rates (Williams et al., 1977) and probably slow down soil formation and successional changes. In natural succession, successful plants must either tolerate or avoid combinations of stress factors likely to cause death. In heterogeneous metalliferous mine sites, avoidance is possible by colonizing less “severe” microsites. Where rapid evolution of tolerance in pseudometallophyte grasses with suitable genetic variation occurs, the level of tolerance is usually related to the metal levels in the site from which they were collected. An interesting study by Hogan et al. (1977a,b) of waste from copper and nickel mines in Canada showed that the success of the species Agrostis gigantea was a result

of employing both avoidance and tolerance strategies. Non-tolerant plants formed “islands” in areas where pH was higher and water-extractable metals lower than in non-vegetated areas. However, experimental studies of copper tolerance showed that, in two vegetated areas with high copper and acidic pH values, the plants of Agrostis gigantea present were copper-tolerant. The ecosystems which have developed on metalliferous mining wastes through unmanaged natural succession are mainly open or closed herbaceous swards with low-growing species and low productivity. In most modern mining situations, however, it is necessary to develop ecosystems which can provide the basic goals of erosion control and improvement of landscape value in a much shorter time-scale. A summary of possible approaches for accelerating ecological restoration of metalliferous wastes is given in Table 14.7. Coal strip-mines There are two common types of abandoned sites associated with strip-mining. One is associated with area strip-mining in ﬂat or undulating land, and the other with contour strip-mining in hilly terrain. Area strip-mining can give rise to saw-tooth ridge topographies where the topsoil and overburden is mixed and not re-contoured to the original. In contour stripmining, the mineral deposits are removed from a short distance into the hill with the cut following the contour line along the hill. The overburden is thrown down the hill leaving a highwall–bench–outslope topography

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(Bradshaw and Chadwick, 1980). Studies of these two types of abandoned sites in the United States, after strip-mining for coal, have provided insights into ecosystem development. In most of the coal stripmining areas of the United States, few reclamation procedures were standard practice until 1977, when federal legislation, the Surface Mining Control and Reclamation Act, was passed (Daniels and Zipper, 1988; Safaya and Wali, 1992). In Oklahoma, 12 000 ha of old area-stripped sites were abandoned (unreclaimed) before 1971 (Johnson et al., 1982). A typical site is 100–200 m wide and up to 1 km long, with a high ridge on one side due to the overburden being put on unmined land for the ﬁrst cut. Subsequent cuts were backﬁlled but left as a series of parallel spoil ridges and valleys; the elevation difference from crest to valley was 5– 25 m. In a survey of 49 such sites 10–70 years old (Johnson et al., 1982; Gibson et al., 1985), there was considerable variation in substrate variables, typical of abandoned sites where there had been a mixing of surface and overburden layers. Substrate factors regarded as signiﬁcant in ecosystem development were: pH (3.5–8.0); clay content (5.4–43.3%); calcium (197– 7069 mg g−1 ); nitrogen (906–3272 mg g−1 ); and iron (49–1315 mg–g−1 ). The vegetation of the Oklahoma mines was typically a medium to dense cover of small trees, shrubs, and woody vines, with 59 tree and 77 woody understorey species. Two species dominated the tree canopy, Celtis laevigata and Ulmus americana, and four other species were common: Diospyros virginiana, Populus deltoides, Salix nigra, and Ulmus alata. The understorey was variable in species composition. Table 14.8 provides a comparison of some spoil and vegetation factors found on two contiguous sites 17 and 55 years old with very similar pH values. There was a clear development in vegetation structure and changes in some key substrate variables with time. In this study the main age-related substrate factors, increases in clay content and amounts of calcium and potassium, were associated with the weathering of soft shales, which were the predominant material of the spoil banks. In addition, nitrogen increased through biological activity. As forest vegetation developed on the mine spoils, the species composition was different from that of the oak–hickory (Quercus–Carya) upland forests in the region and more like “bottomland” forests on moist

377 Table 14.8 Spoil properties and vegetation structure of two contiguous unreclaimed coal strip-mine sites of different ages in Oklahoma 1 17 years old

55 years old

Vegetation structure Trees ha−1 Basal area m2 ha−1 Tree species Vegetation height index 2 (m) Woody understorey density 3 Grass cover (%)

148 0.47 12 5.4 2147

510 12.69 19 13.8 17497

12.2

2.0

7.1

7.2

Spoil properties pH Clay (%)

11.8

24.6

Ca (mg g−1 )

2089

2986

K (mg g−1 )

197

280

Fe

(mg g−1 )

214

99

Total N (mg g−1 )

1139

2914

Total P (mg g−1 )

500

362

1

From Johnson et al. (1982). “Vegetation height index” is the mean height of the tallest plant in each of the replicate quadrats sampled. 3 Plants per hectare. Saplings, tree seedlings, shrubs, woody vines and cacti are included if their stem diameter is less than 5 cm at a height of 140 cm. 2

and well-drained alluvium (Gibson et al., 1985). This resemblance was explained in part by the differing dispersal abilities of individual species, and in part by variation in edaphic and microclimatic factors. The species with mammal-dispersed seed were poorly represented on the abandoned mines. The main species on the mined land were present in sites of all ages, suggesting that species replacement is not a dominant mechanism of vegetation development on these sites. Contour strip-mining was typically the approach in the Appalachian Plateau region of the United States, where hundreds of thousands of hectares of land have been disturbed since the early part of the 20th century. A typical site consists of natural vegetation on an upper, non-mined slope with a steep high wall beneath it where the cut was taken to leave a bench with a steep outslope below. The bench may have compacted overburden restricting plant colonization, with permanent swampy areas towards the base of the wall. The overburden removed from the cut was

378

bulldozed over the outslope, often generating unstable slope conditions (Daniels and Zipper, 1988). Holl and Cairns (1994) surveyed eighteen such contour strip-mined sites in Virginia and grouped them into four age classes together with ﬁve reference sites in surrounding unmined sites bearing mixed hardwood forest. All of the mine sites had some kind of initial planting. The vegetation on the bench and south-facing outslopes was sampled. A detrended correspondence analysis (DCA) of the plant community composition was used to explore the trends in the data set. The ﬁrst axis of the DCA explained 20% of the variation and was strongly correlated with site age. The next two axes explained only a further 10% of the variation. Site age was particularly related to the increase in species richness of trees and shrubs and tree cover. However, in the oldest mined stands (25–30 yr) the tree basal area was only half that in the reference sites; of the total of 44 trees and shrubs recorded in all sites, 16 were found only in the reference sites. These included some of the characteristic species of undisturbed upland forests such as Acer saccharum, Fagus grandifolia, Juglans nigra, and Quercus alba. In both these case studies of abandoned strip mines in Oklahoma and Virginia, and in other states, such as New Mexico (Wagner et al., 1978), Pennsylvania (Brenner et al., 1984; Schuster and Hutnik, 1987), and Texas (Gorsira and Risenhoover, 1994), it is likely that the time scale for development of forest ecosystems is more than 50 years. These studies highlight the “temporal paradox” for ecological restoration, when reclamation legislation such as the Surface Mining Control and Reclamation Act (SMCRA) in the United States demands the achievement of speciﬁc postreclamation criteria in just a few years. The SMCRA requires the use of topsoil, mandates that original contours be established, that soils are stabilized rapidly and that 90% of the original site productivity is restored within 5 years (Daniels and Zipper, 1988; Holl and Cairns, 1994; Chaney et al., 1995). These legal stipulations have lead to practices which can have negative effects on the promotion of species colonization and natural succession. To achieve rapid site stabilization the use of commercial species of grasses and legumes to give rapid erosion control and cover is common, but this can prevent the entry of volunteer species and arrest succession (Luken, 1990). Similarly, the use of fast-growing species of trees can be either beneﬁcial or detrimental depending on the species planted. The extensive use of black

John A. COOKE

locust, Robinia pseudoacacia, has been found to encourage further colonization whereas eastern white pine, Pinus strobus, can be detrimental by excluding species and delaying succession (Schuster and Hutnik, 1987; Leopold and Wali, 1992). It is clear that the needs of legislation (and the mining industry) for shortterm reclamation achievements can conﬂict with, and so prevent, the longer-term establishment of stable, diverse, and self-sustaining ecosystems. Restoration of ecosystems on mined land The best practice in modern mining demands the full integration of environmental considerations into the planning of all the stages of a mining operation. To avoid the “temporal paradox” and successfully restore natural ecosystems, this is essential. Before mining starts baseline research on the pre-mining ecosystem structure needs to be undertaken and a detailed restoration plan produced. The ecological considerations needed for such restoration planning are given in Table 14.9. These are discussed further in relation to two case studies with restoration histories going back to the late 1970s: dredge mining of coastal sand for heavy minerals in South Africa, and area stripmining for bauxite in Australia. As noted in Table 14.3, the land disturbance of the primary extraction phases of both strip and dredge mining can be progressively restored and reasonable restoration success can be achieved if based on ecological considerations. Dune mining in South Africa On the eastern seaboard of South Africa in KwaZulu– Natal near Richards Bay, dredge mining in coastal dunes for heavy minerals such as rutile, ilmenite, and zircon has been taking place since 1977 (Camp, 1990). Ecosystem restoration is a fundamental part of the mining operation. Mining entails the removal of the dune forest in a prescribed mining path through the dunes. Topsoil is then stored and an artiﬁcial pond is created in the ore-bearing dune, and the sand is mined using a ﬂoating dredger, which pumps the sand in a slurry to a gravity separator where the heavy minerals are separated from the sand. The mined sand is then pumped back as tailings behind the mining path, dewatered and stacked to resemble the topography prior to mining. Over 90% of the sand bulk is returned after mineral extraction.

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379

Table 14.9 Basic ecological considerations in planning for the restoration of pre-mining ecosystems 1

Table 14.10 Ecosystem restoration after dune mining in KwaZulu–Natal, South Africa: a comparison of two stands of different ages with a mature forest reference stand for the number of animal species 1

Consideration

Possible questions

1.

Land-use objectives

Ecologically possible? sustainable?

2.

Clearing existing vegetation

Needed as seed source, mulch, erosion control? piled material as habitat for animals?

Millipedes i) sweep net

Topsoil handling

Double-strip for seed-bank retention? timing of stripping, stockpiling, and replacement? reinstatement alternatives if unsuitable?

ii) ﬂight intercept

Species

3.

4.

Earthworks

Final landforms and erosion control?

5.

Revegetation

Species and provenance selection? means of plant establishment? seed-bed preparation – fertilizers and soil amendments needed?

6.

Nutrient accumulation and cycling

Use of amendments, nurse species, inoculation with symbionts and mycorrhizal fungi necessary?

7.

Indigenous volunteer Direct introduction, or facilitation of species of plants and natural colonization? animals

8.

Management

Replanting or reintroduction? erosion control, ﬁre management, pest and weed control, fertilizer application, watering and irrigation, further amendments?

9.

Monitoring

What should be monitored? sampling – how, when, where? what management should be undertaken?

10. Success criteria

What criteria? over what time scale?

1

Based on Australian Environment Protection Agency (1995).

To promote the establishment of indigenous dune forest, the topsoil is re-spread on the non-toxic tailings to a depth of about 10 cm. Then, artiﬁcial windbreaks (shade-cloth fences) are erected, and the area is sown with a mixture of seeds consisting mainly of fast-germinating species (Sorghum spp., Pennisetum americanum, Crotalaria juncea). This mix acts as a nurse crop which protects the slower-germinating indigenous species present in the seed bank from the high surface temperatures and winds. After this nurse crop has died off the vegetation is dominated by Acacia karoo, which is regarded as the major pioneer species in the succession in this area (Camp and Weisser, 1991). Over 400 ha have been reclaimed in this way since 1978, providing a chronosequence of reclaimed dunes. A summary of vegetational changes

5–8 years

11–16 years

Ref. stand

3

6

11

54

80

116

74

107

188

1

3

4

Beetles caught by

Rodents 1

From Van Aarde et al. (1996).

and numbers of some animal taxa are given in Tables 14.10 and 14.11. A reference or benchmark site of a mature dune forest without mining disturbance is given for comparison. Vegetation development on the mined dunes appears to follow succession on natural disturbances and will lead to dune forest typical of the area. Species richness increases with time as the pioneer tree species (predominantly Acacia karoo) are replaced by tropical and sub-tropical broad-leaved shrubs and trees (Fig. 14.5; see also Mentis and Ellery, 1994; Van Aarde et al., 1996). Functionally, the nitrogen-ﬁxing capability of Acacia karoo promotes an increase of soil nutrients and soil development (Lubke et al., 1993), which also facilitates the establishment of non-pioneer species. The changes in vegetation structure were accompanied by increases in the number of animal taxa (Tables 14.10 and 14.11). Similarity coefﬁcients comparing the rehabilitating stands of different ages and the reference stand increased with increasing age since rehabilitation (Van Aarde et al., 1996). Through the death of older trees and the creation of gaps in the dense Acacia karoo stands, colonization by secondary tree-canopy species is facilitated (Fig. 14.5). However, it is very likely that these secondary tree species were not recruited from the seed banks of the original replaced topsoil, but were dispersed into the older rehabilitated areas by fruit-eating birds and vervet monkeys (Cerceopithecus aethiops) (Foord et al., 1994). Thus, the presence of nearby pristine forest, even as small pockets or strips, could be a key aspect of restoration success. The spatial relationship of patches of undisturbed natural ecosystems to the mined area is an often neglected facet of restoration planning which should be addressed before mining starts.

380

John A. COOKE

Table 14.11 Ecosystem restoration after dune mining in KwaZulu–Natal, South Africa: a comparison of two stands of different ages with a mature forest reference stand for vegetation 1 5–8 years

11–16 years

Reference stand

Acacia karoo (sweet thorn) scrub 1.5–3.0 m high, sparse middle layer of Brachylaena discolor (coast silver oak) and Vepris lanceolata (white ironwood); herb layer mainly grasses Panicum maximum and Digitaria diversinervis

Acacia karoo 9–12 m high with some secondary dune-forest species such as Celtis africana (white stinkwood), Trema orientalis (pigeon wood) and Trichilia emetica (Natal mahogany); herb layer mainly Digitaria diversinervis

Secondary dune forest with canopy 12–15 m or higher with main species: Allophylus natalensis (dune false currant), Celtis africana, Mimusops caffra (coastal red milkwood), Ochna natalitia (Micky Mouse bush) and Teclea gerrardii (Zulu cherry-orange); herb and shrub layer dominated by the shrub Isoglossa woodii and the fern Microsorum scolopendria

1

From Van Aarde et al. (1996).

Fig. 14.5. Dune restoration after dredge-mining for heavy minerals at Richards Bay, KwaZulu Natal, South Africa (original photograph Richards Bay Minerals). Left: A white stinkwood (Celtis africana) growing under a canopy of Acacia karoo, 10 years after top soil re-instatement. Right: Four or ﬁve years after the re-instatement of topsoil with a dense cover dominated by Acacia karoo.

Bauxite mining in Western Australia Bauxite mining in the northern jarrah (Eucalyptus marginata) forest in Western Australia currently requires the restoration of 450 ha of forest per year

(Baker et al., 1995). The vegetation is dominated by the relatively slow-growing Eucalyptus marginata, which attains a height of 30–40 m. Other eucalypt tree species present include E. megacarpa (bullich),

MINING

E. patens (blackbutt) and E. calophylla (marri), and an understorey of many species including the cycad Macrozamia riedlei. The ore is mined from numerous discrete shallow deposits 3–6 m deep and 10–50 ha in area. The soils are typically nutrient-poor laterite gravels that overlie a pallid zone of kaolinitic clay (Tracey and Glossop, 1980). The basic mineral extraction is technically simple, but large areas must be mined. First the forest is cleared, with some timber kept to provide animal habitat on the restored area. The top 50 cm of topsoil is stripped separately to maintain the seed bank (350– 1500 germinable seeds m−2 ). If possible it is not stockpiled, as this can reduce the number of germinable seeds, cause the loss of nutrients, and reduce aerobic microbial activity (Tracey and Glossop, 1980). The jarrah forest is a nutrient-poor ecosystem and most of the native species have vesicular-arbuscular mycorrhizae and ectomycorrhizae, which are rapidly lost in the stockpiling of topsoil, leading to low levels of re-infection in the early years after restoration (Jasper et al., 1987; Baker et al., 1995). Overburden above the caprock is then removed and stockpiled. The caprock is blasted and the ore removed down to the pallid zone of the original laterite soil proﬁle. Ore extraction leaves shallow pits with vertical faces 2–5 m high and a compacted clay ﬂoor. The entire pit surface is then deep-ripped along the contours, breaking up the clay ﬂoor to allow root penetration and prevent waterlogging. The pit walls are battered down and reshaped, and the overburden and topsoil are replaced in the correct sequence. Further surface-ripping then occurs to improve drainage, as waterlogging is very undesirable because it encourages the soil-borne fungus Phytophthora cinnamomi causing die-back in jarrah and some understorey species. In fact, much emphasis is placed in the restoration on the prevention of spread of dieback, so that infected soils are stripped and stored separately, and vehicles are cleaned if they move from an infected to an uninfected area (Baker et al., 1995). As in the South African dune-forest restoration, the topsoil seed bank is the major source of seed for the developing ecosystem. However, in the restoration of the jarrah forest this is considerably augmented by a seed mix, which is sown by hand across the entire restored pit surface. This mix includes more than 60 different native species, including understorey legumes and Macrozamia riedlei. In a ﬁre-regulated system such as the jarrah forest, the main restoration

381

objective is to establish a large proportion of the 141 common plant species as soon as possible. This includes achieving a minimum success criterion for plant establishment of 2000 eucalypt seedlings per hectare and 2 legume understorey species per square metre after nine months. These objectives have required speciﬁc research into the reproductive biology of many plant species (e.g., vegetative propagation of species with recalcitrant seeds) (Baker et al., 1995). Successful plant establishment has been increased by sowing immediately after ripping of the topsoil. This allows seed to become lodged in microsites before surface crusts develop. The crusts prevent good seed–soil water contact. Monitoring and evaluation of the restored areas over the last 15 years has shown satisfactory ecosystem development compared to the unmined areas of jarrah forest. The likely time-scales for the achievement of various success criteria are: plant species richness (<2 yr), litter biomass (4–10 yr), nitrogen capital (15 yr), 80% of animal groups including birds, mammals, reptiles, and invertebrates (4–6 yr) (Nichols et al., 1989; Baker et al., 1995).

CONCLUSIONS

The result of mining disturbance is usually complete destruction of natural ecosystems for more than 60 years and possibly for millennia, as in the case of large tailings impoundments. The scale of the direct disturbance to land may be large (>2000 ha per mine site), with even greater geographical impacts possible through pollution of air and water. Other multiplier effects can occur through the fragmentation of original natural ecosystems and alteration in surface and groundwater drainage patterns. In general, disturbance is unlikely to recur; this may happen, however, where mining wastes are regarded as secondary ore bodies (e.g., the re-working of old spoil heaps and tailings dams for metals). The likelihood of the original ecosystem developing through natural recovery is very low, as most abandoned mined land is left with severe physical and chemical constraints to plant and animal colonization and survival. Ecosystems which have developed naturally on mined land often remain in an arrested successional stage such as open herb-rich grassland, and provide opportunities for rare species and species populations which are stress-tolerators but

382

poor competitors. Such ecosystems may have persisted for many years with little change, and many have become important for the conservation of genetic diversity. Successional processes on abandoned mine land may lead to woodland or forest, but these usually take a long time (>50 yr). Also, the ecosystems which do develop are often different from the typical climax vegetation of the region because of differences in the soils of mined land. Unless speciﬁc restoration goals are integrated into the mine planning, it is unlikely that the pre-mined ecosystem will develop naturally. Where local topsoil can be replaced quickly, as in the examples of progressive reclamation during strip or dredge mining, and where ecological principles, developed through research, guide the restoration planning, then success may be achieved. In many mining situations, more modest goals may have to be accepted, simple ecosystems being established and maintained for aesthetic and safety reasons. However, it is important to try to maintain future options for longer-term ecological sustainability.

ACKNOWLEDGEMENTS

The author would like to thank: Dehn von Ahlefeldt (University of Natal), Alan Baker (University of Shefﬁeld) and Ronwyn Stander (Richards Bay Minerals) for help with the photographs; and the South African Foundation for Research Development, and the University of Natal Research Fund for ﬁnancial support.

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John A. COOKE Processing, Conference, Sydney. Australasian Institute of Mining and Metallurgy, pp. 43–53. Barbour, A.K., 1994. Mining non-ferrous metals. In: R.E. Hester and R.M. Harrison (Editors), Mining and its Environmental Impact. Issues in Environmental Science and Technology. Royal Society of Chemistry, Letchworth, England, pp. 1–15. Bradshaw, A.D., 1983. The Reconstruction of Ecosystems: Presidential address to the British Ecological Society, December 1982. J. Appl. Ecol., 20: 1–17. Bradshaw, A.D., 1990a. Restoration: an acid test for ecology. In: W.R. Jordan, M.E. Gilpin and J.D. Aber (Editors), Restoration Ecology: A Synthetic Approach to Ecological Research. Cambridge University Press, Cambridge, pp. 23–29. Bradshaw, A.D., 1990b. The reclamation of derelict land and the ecology of ecosystems. In: W.R. Jordan, M.E. Gilpin and J.D. Aber (Editors), Restoration Ecology: A Synthetic Approach to Ecological Research. Cambridge University Press, Cambridge, pp. 53–74. Bradshaw, A.D. and Chadwick, M.J., 1980. The Restoration of Land: The Ecology and Reclamation of Derelict and Degraded Land. Blackwell, Oxford, 317 pp. Bradshaw, A.D., Marrs, R.H. and Roberts, R.D., 1982. Succession. In: B.N.K. Davis (Editor), Ecology of Quarries: the Importance of Natural Vegetation. Institute of Terrestrial Ecology, Abbots Ripton, pp. 47–52. Brenner, F.J., Werner, M. and Pike, J., 1984. Ecosystem development and natural succession in surface coal mine reclamation. Miner. Environ., 6: 10–22. Brooks, R.R. and Malaisse, F., 1985. The Heavy Metal-Tolerant Flora of Southcentral Africa. A.A. Balkema, Rotterdam, 199 pp. Buchanan, D.J. and Brenkley, D., 1994. Green Coal Mining. In: R.E. Hester and R.M. Harrison (Editors), Mining and its Environmental Impact. Issues in Environmental Science and Technology. Royal Society of Chemistry, Letchworth, England, pp. 71–95. Buckley, G.P., 1989. Biological Habitat Construction. Belhaven Press, London, pp. 363. Cairns, J.J., 1991. The status of the theoretical and applied science of restoration ecology. Environ. Prof., 13: 186–194. Camp, P.D., 1990. Rehabilitation after dune mining at Richards Bay Minerals. S. Afr. Min. World, October: 34–37. Camp, P.D. and Weisser, P.J., 1991. Dune rehabilitation, ﬂora and plant succession after mining at Richards Bay, South Africa. In: D.A. Everard and G.P. Von Maltitz (Editors), Dune Forest Dynamics in Relation to Land-Use Practices. Foundation for Research Development, Pretoria, pp. 106–123. Chaney, W.R., Pope, P.E. and Byres, W.R., 1995. Tree survival and growth on land reclaimed in accord with public law 95–87. J. Environ. Qual., 24: 630–634. Cooke, J.A. and Morrey, D.R., 1981. Heavy metals and ﬂuoride in soils and plants associated with metalliferous mine wastes in the Northern Pennines. In: P.J. Say and B.A. Witton (Editors), Heavy Metals in Northern England: Environmental and Biological Aspects. University of Durham, Durham, pp. 153–164. Daily, G.C., 1995. Restoring value to the world’s degraded lands. Science, 269: 350–354. Daniels, W.L. and Zipper, C.E., 1988. Improving coal surface mine reclamation in the Central Appalachian region. In: J. Cairns Jr (Editor), Rehabilitating Damaged Ecosystems. CRC Press, Boca Raton, pp. 139–162.

MINING Davis, B.N.K., 1979. Chalk and limestone quarries as wildlife habitats. Miner. Environ., 1: 48–56. Davis, B.N.K., 1982. Ecology of Quarries: the Importance of Natural Vegetation. Institute of Terrestrial Ecology, Abbots Ripton, 77 pp. Davis, B.N.K., Lakhani, K.H., Brown, M.C. and Park, D.G., 1985. Early seral communities in a limestone quarry: an experimental study of treatment effects on cover and richness of vegetation. J. Appl. Ecol., 22: 473–490. Duckham, F., 1969. Introduction. In: R.L. Galloway (Editor) A History of Coal Mining in Great Britain. David and Charles, Newton Abbot, pp. 273, reprint of 1882 edition. Ellis, D.V., 1988. Case histories of coastal and marine mines. In: W. Salomons and U. Forstner (Editors), Chemistry and Biology of Solid Waste: Dredged Material and Mine Tailings. SpringerVerlag, Berlin, pp. 73–100. Ernst, W.H.O., 1988. Response of plants to mine tailings and dredged materials. In: W. Salomons and U. Forstner (Editors), Chemistry and Biology of Solid Waste: Dredged Material and Mine Tailings. Springer-Verlag, Berlin, pp. 54–69. Ferguson, K.D. and Erickson, P.M., 1988. Pre-mine prediction of acid-mine drainage. In: W. Salomons and U. Forstner (Editors), Environmental Management of Solid Waste: Dredged Material and Mine Tailings. Springer-Verlag, Berlin, pp. 24–43. Foord, S.H., van Aarde, R.J. and Ferreira, S.M., 1994. Seed dispersal by vervet monkeys in rehabilitating coastal dune forests at Richards Bay. S. Afr. J. Wildl. Res., 24: 56–59. Galloway, R.L., 1882. A History of Coal Mining in Great Britain. David and Charles, Newton Abbott, 273 pp., reprinted 1969. Gibson, D.J., Johnson, F.L. and Risser, P.G., 1985. Revegetation of unreclaimed coal strip-mines in Oklahoma. II. Plant Communities. Reclam. Revegetation Res., 4: 31–47. Glover, H.G., 1978. The disposal of coal mine spoil in the United Kingdom. In: G.T. Goodman and M.J. Chadwick (Editors), Environmental Management of Mineral Wastes. Nato Advanced Study Institutes Series. Series E: Applied Science. Sijthoff and Noordhoff, Alphen aan den Rijn, pp. 35–69. Gorsira, B. and Risenhoover, K.L., 1994. An evaluation of woodland restoration on strip-mined lands in east Texas. Environ. Manage., 18: 787–793. Gunn, J., Bailey, D. and Gagen, P., 1992. Landform Replication as a Technique for the Reclamation of Limestone Quarries. Department of the Environment, HMSO, London, 38 pp. Guo, H., Wu, D. and Zhu, H., 1989. Land restoration in China. J. Appl. Ecol., 26: 787–792. Hodgson, J.G., 1989. Selecting and Managing Plant Materials Used in Habitat Reconstruction. In: G.P. Buckley (Editor), Biological Habitat Construction. Belhaven Press, London, pp. 45–67. Hogan, G.D., Courtin, G.M. and Rauser, W.E., 1977a. The effects of soil factors on the distribution of Agrostis gigantea on a mine waste site. Can. J. Bot., 55: 1038–1042. Hogan, G.D., Courtin, G.M. and Rauser, W.E., 1977b. Copper tolerance in clones of Agrostis gigantea from a mine waste site. Can. J. Bot., 55: 1043–1050. Holl, K.D. and Cairns Jr., J., 1994. Vegetational community development on reclaimed coal surface mines in Virginia. Bull. Torrey Bot. Club, 121: 327–337. Humphries, R.N., 1980. The development of wildlife interest in limestone quarries. Reclam. Rev., 3: 197–207. Hutchinson, T.C. and Whitby, L.M., 1974. Heavy metal pollution

383 in the Sudbury mining and smelting region of Canada. 1. Soil and vegetation contamination by nickel, copper, and other metals. Environ. Conservation, 1: 123–132. Jasper, D.A., Robson, A.D. and Abbott, L.K., 1987. The effect of surface mining on the infectivity of vesicular-arbuscular mycorrhizal fungi. Aust. J. Bot., 35: 641–652. Johnson, F.L., Gibson, D.J. and Risser, P.G., 1982. Revegetation of unreclaimed coal strip-mines in Oklahoma. 1. Vegetation structure and soil properties. J. Appl. Ecol., 19: 453–463. Johnson, M.S. and Bradshaw, A.D., 1979. Ecological principles for the restoration of disturbed and degraded land. Adv. Appl. Biol., 4: 141–200. Johnson, M.S., Putwain, P.D. and Holliday, R.J., 1978. Wildlife conservation value of derelict metalliferous mine workings in Wales. Biol. Conservation, 14: 131–148. Johnson, M.S., Cooke, J.A. and Stevenson, J.K., 1994. Revegetation of metalliferous wastes and land after metal mining. In: R.E. Hester and R.M. Harrison (Editors), Mining and its Environmental Impact. Issues in Environmental Science and Technology. Royal Society of Chemistry, Letchworth, England, pp. 31–48. Klemow, K.M. and Raynal, D.J., 1981. Population ecology of Melilotus alba in a limestone quarry. J. Ecol., 69: 33–44. Land Use Consultants, 1992. Amenity Reclamation of Mineral Workings: Main Report. Land Use Consultants, Department of the Environment, HMSO, London, 242 pp. Leopold, D.J. and Wali, M.K., 1992. The rehabilitation of forest ecosystems in the eastern United States and Canada. In: M.K. Wali (Editor), Ecosystem Rehabilitation. Volume 2. SPB Academic Publishing, The Hague, pp. 187–231. Lubke, R.A., Moll, J.B. and Avis, A.M., 1993. Rehabilitation Ecology. In: C.E. Services (Editors), Environmental Impact Assessment, Eastern Shores of Lake St. Lucia (Kingsa/Trojan Lease Area). CSIR, Pretoria, pp. 251–302. Luken, J.O., 1990. Directing Ecological Succession. Chapman and Hall, London, 251 pp. McNeilly, T., 1990. Evolutionary lessons from degraded ecosystems. In: W.R. Jordan, M.E. Gilpin and J.D. Aber (Editors), Restoration Ecology: A Synthetic Approach to Ecological Research. Cambridge University Press, Cambridge, pp. 271–286. Mentis, M.T. and Ellery, W.N., 1994. Post-mining rehabilitation of dunes on the north-east coast of South Africa. S. Afr. J. Sci., 90: 69–74. Miles, J., 1987. Vegetation succession: past and present perceptions. In: A.J. Gray, M.J. Crawley and P.J. Edwards (Editors), Colonisation, Succession and Stability. Blackwell, Oxford, pp. 1– 29. Mining Annual Review, 1985. Mining Annual Review. Mining Journal Ltd, London, 556 pp. Mining Annual Review, 1995. Mining Annual Review. Mining Journal Ltd, London, 248 pp. Morrey, D.R., Baker, A.J.M. and Cooke, J.A., 1988. Floristic variation in plant communities on metalliferous mining residues in the northern and southern Pennines, England. Environ. Geochem. Health, 10: 11–20. Nichols, O., Wykes, B.J. and Majer, J.D., 1989. The return of vertebrate and invertebrate fauna to bauxite mined areas in southwestern Australia. In: J.D. Majer (Editor), Animals in Primary Succession: the Role of Fauna in Reclaimed Lands. Cambridge University Press, Cambridge, pp. 397–422.

384 Palmer, M.E., 1978. Acidity and nutrient availability in colliery spoil. In: G.T. Goodman and M.J. Chadwick (Editors), Environmental Management of Mineral Wastes. Nato Advanced Study Institutes Series. Series E: Applied Science. Sijthoff and Noordhoff, Alphen aan den Rijn, pp. 85–126. Pierzynski, G.M., Schnoor, J.L., Banks, M.K., Tracy, J.C., Licht, L.A. and Erickson, L.E., 1994. Vegetative remediation at Superfund sites. In: R.E. Hester and R.M. Harrison (Editors), Mining and its Environmental Impact. Issues in Environmental Science and Technology. Royal Society of Chemistry, Letchworth, England, pp. 49–69. Ratcliffe, D.A., 1974. Ecological effects of mineral exploitation in the United Kingdom and their signiﬁcance to nature conservation. Proc. R. Soc. London Ser. A., 339: 355–372. Richardson, J.A., Davis, B.N.K. and Evans, M.E., 1980. Disused quarries. In: T.C. Dunn (Editor), The Magnesian Limestone of Durham County. The Gilpin Press, Houghton le Spring, pp. 61– 68. Riley, C.V., 1957. Reclamation of coal strip-mined lands with reference to wildlife plantings. J. Wildl. Manage., 21: 402–413. Ritcey, G.M., 1989. Tailings Management: Problems and Solutions in the Mining Industry. Vol. 6. Process Metallurgy. Elsevier Amsterdam, 970 pp. Safaya, N.M. and Wali, M.K., 1992. Applicability of U.S. environmental laws in the developing countries: an analysis of ecological and regulatory concepts. In: M.K. Wali (Editor), Ecosystem Rehabilitation, Vol. 1. SPB Academic Publishing, The Hague, pp. 143–155. Schuster, W.S. and Hutnik, R.J., 1987. Community development on 35-year-old planted minespoil banks in Pennsylvania. Reclam. Revegetation Res., 6: 109–120. Sengupta, M., 1993. Environmental Impacts of Mining: Monitoring, Restoration and Control. Lewis, Boca Raton, 494 pp. Simon, E., 1978. Heavy metals in soils, vegetation development and heavy metal tolerance in populations from metalliferous areas. N. Phytol., 81: 175–188. Smith, R.A.H. and Bradshaw, A.D., 1979. The use of metal tolerant plant populations for the reclamation of metalliferous wastes. J. Appl. Ecol., 16: 595–612. South African Mining, 1996. South African Mining; Coal, Gold, and Base Minerals. February. Thomson Publishing, Randburg, p 53. Thatcher, F.M., 1979. A Study of the Vegetation Established on the Slimes Dams of the Witwatersrand. PhD, University of the Witwatersrand, Johannesburg, 518 pp. Thirgood, J.V., 1978. Approaches to land reclamation in Britain and North America. In: G.T. Goodman and M.J. Chadwick (Editors), Environmental Management of Mineral Wastes. Nato Advanced Study Institutes Series. Series E: Applied Science. Sijthoff and Noordhoff, Alphen aan den Rijn, pp. 1–18. Thomson, G.M. and Rodin, S., 1972. Colliery Spoil Tips – after Aberfan. The Institution Civil Engineers, London, 60 pp.

John A. COOKE Tracey, W.H. and Glossop, B.L., 1980. Assessment of topsoil handling techniques for rehabilitation of sites mined for bauxite within the Jarrah forest of Western Australia. J. Appl. Ecol., 17: 195–201. United Kingdom Department of the Environment, 1994. The Reclamation and Management of Metalliferous Mining Sites. Department of Environment, HMSO, London, 168 pp. United States National Research Council, 1981. Surface Mining: Soil, Coal, and Society. National Research Council, National Academy Press, Washington D.C., 233 pp. Usher, M.B., 1979. Natural communities of plants and animals in disused quarries. J. Environ. Manage., 8: 223–236. Usher, M.B. and Jefferson, R.G., 1990. The concepts of colonization and succession: their role in nature reserve management. In: S.H. Hillier, D.W.H. Walton and D.A. Wells (Editors), Calcareous Grasslands – Ecology and Management. Bluntishham Books, Huntingdon, pp. 149–153. van Aarde, R.J., Ferreira, S.M., Kritzinger, J.J., van Dyk, P.J., Vogt, M. and Wassenaar, T.D., 1996. An evaluation of habitat rehabilitation on coastal dune forests in northern KwaZulu–Natal, South Africa. Restoration Ecol., 4: 334–345. Vitousek, P.M. and Walker, L.R., 1987. Colonization, succession and resource availability: ecosystem level interactions. In: A.J. Gray, M.J. Crawley and P.J. Edwards (Editors), Colonization, Succession and Stability. Blackwell, Oxford, pp. 207–224. Wagner, W.L., Martin, W.C. and Aldon, E.F., 1978. Natural succession on strip mined lands in northwestern New Mexico. Reclam. Rev., 1: 67–73. Wells, J.D., van Meurs, L.H. and Rabie, M.A., 1992. Terrestrial minerals. In: R.F. Fruggle and M.A. Rabie (Editors), Environmental Management in South Africa. Juta, Cape Town, pp. 337–379. Wild, H., 1978. The vegetation of heavy metal and other toxic soils. In: M.J.A. Werger (Editor), Biogeography and Ecology of Southern Africa. W. Junk, The Hague, pp. 1303–1332. Willems, J.H., 1990. Calcareous grasslands in continental Europe. In: S.H. Hillier, D.W.H. Walton and D.A. Wells (Editors), Calcareous Grasslands – Ecology and Management. Bluntishham Books, Huntingdon, pp. 3–10. Williams, A. and Mitchell, C., 1994. Methane emissions from coal mining. In: R.E. Hester and R.M. Harrison (Editors), Mining and its Environmental Impact. Issues in Environmental Science and Technology. Royal Society of Chemistry, Letchworth, England, pp. 97–109. Williams, S.T., McNeilly, T. and Wellington, E.M.H., 1977. The decomposition of vegetation growing on metal mine waste. Soil Biol. Biochem., 9: 271–275. Williamson, N.A., Johnson, M.S. and Bradshaw, A.D., 1982. Mine Wastes Reclamation. The Establishment of Vegetation on Metal Mine Wastes. Mining Journal Books, London, 103 pp.

Chapter 15

DISTURBANCE ASSOCIATED WITH MILITARY EXERCISES Stephen DEMARAIS, David J. TAZIK, Patrick J. GUERTIN and Eric E. JORGENSEN

INTRODUCTION

Disturbances to ecosystems caused by military activities must be distinguished according to whether they occur during times of war or peace. In war, environmental planning is not a predominant concern. During peacetime, disturbances may accompany military training activities (Lanier-Graham, 1993), often with an opportunity to document, minimize, and mitigate their negative impacts on the environment. Wartime disturbances often are catastrophic and at large spatial and temporal scales. Disturbance impacts may cross regional and national boundaries, particularly for such conﬂicts as World Wars I and II. During wartime battles, with victory the primary concern, there is little concern for adverse effects on the environment; environmental planning and management play little part. Impacts are not studied in advance, and documentation may not be undertaken at all until political arenas are stabilized. Defeat of a nation or group of peoples can be facilitated by eliminating their environmental resources. General Sherman of the Union Army practiced the “scorched earth” approach as he made his way to capture Atlanta, Georgia, during the Civil War in the United States during the 1860s, burning all resources of potential use by his enemy. During World War II, large areas of reclaimed lands in the Benelux countries were ﬂooded with sea water to impede the German invasion (Barrow, 1991). During the Vietnam War, up to 40% of the land area of Vietnam, and 44% of all forests, were sprayed with defoliants (Gradwohl and Greenberg, 1988). This defoliation, combined with effects from bombs and heavy equipment, destroyed forest vegetation over an estimated 22×106 ha in South and North Vietnam (Barrow, 1991). Recovery of these

defoliated forests has been slow and soil erosion prevalent on steep slopes (Freedman, 1989). Direct impacts of military munitions can be signiﬁcant. In Vietnam, an estimated 25×106 bomb and shell craters displaced about 3×109 m3 of soil (Gradwohl and Greenberg, 1988). A 227 kg high explosive bomb can form a crater 14 m in diameter and 9 m deep, and the crater may still be plainly visible 25 years later. Roughly 3.5×106 such bombs were dropped in Vietnam during 1968 and 1969 (Stanford Biology Study Group, 1971). Turning to military activities in peacetime, the management of military lands involves facilitating use of those lands for training. Inevitably, use leads to disturbance. The focus of this chapter is on disturbance caused by military training activities, and management scenarios that reduce or mitigate such disturbance. Military training, while not as dramatic a disturbance as the direct effects of war, can have a widespread inﬂuence on the land. Most of our information comes from lands administered by the United States Department of Defense (DoD). We will discuss the impacts of army training in four countries (Australia, Canada, Germany, and the United States). It is the policy of the United States Army to maintain training lands in a condition which closely mimics the natural conditions under which actual warfare would be conducted, as well as for wildlife habitat and other natural values (Hinchman et al., 1990; Goodman, 1996). In particular, it is the goal of the Department of Defense to reduce or avoid long-term impacts on natural resources caused by military training (Prose, 1985). Military land-use is frequently intensive, especially where maneuvers are conducted with tracked vehicles (Shaw and Diersing, 1990). For instance, at Pinon Canyon Maneuver Site, Fort Carson, Colorado, after

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two years (six training rotations), a line-transect study in three out of ﬁve training areas, with observations at one-meter intervals, showed that 40% of the surface area had been impacted by tracked vehicles (Shaw and Diersing, 1990). Under such training regimes, military land-managers are responsible for maintaining and rehabilitating training lands (Diersing and Severinghaus, 1984). Importantly, many habitats and species are protected from private development by the presence of military installations (Owens, 1990; Creswell, 1994; Goodman, 1996). Also, base longevity and maintenance of realistic habitat conditions for training require proactive resource management (Hinchman et al., 1990; O’Neil et al., 1990; Pearson et al., 1990). Therefore, it is good policy for the military to manage its lands in ways consistent with sound stewardship. Disturbances resulting from military activities are multi-faceted, affecting ecosystems at several points; soil structure is damaged through compaction causing erosion, and vegetation is damaged or destroyed causing modiﬁcation and loss of wildlife habitat. Such damage to the soil and the vegetation compromise the realism of the training exercise, creating an incentive to train in alternative undisturbed locations. This expansion of the training grounds creates further disturbance (Hinchman et al., 1990; Trumbull et al., 1994). The management approach on lands of the United States Department of Defense has increasingly incorporated the concept of ecosystem management. This concept allows the Department to consider the full array of natural resources on lands under their stewardship (Goodman, 1996). For some time, it was a priority to identify “indicator species” which could be used to indicate early stages of habitat disturbance (e.g., Diersing and Severinghaus, 1985). More recently, documentation and monitoring of long-term change has received emphasis. The approach has thus focused on development of repeatable methodologies, appropriate in many ecosystems (Tazik et al., 1992b). Description of military lands Management of military lands is a unique challenge: testing of and training in the use of advanced weapon systems that are of longer range and more devastating requires a more extensive land area. Not surprisingly, the Department of Defense is the third largest steward of land in the government of the United States, with management authority over 10×106 ha

(Goodman, 1996). Nearly half of this land (4.8×106 ha) is controlled by the United States Army (Shaw and Diersing, 1990); consequently, the Army has the largest impact and the most control over the disturbance of military lands. Worldwide, armed forces control between 7.5×105 and 1.5×106 ha of the earth’s land area (Thomas, 1995). Military lands exist in almost every ecosystem covered in this volume. Diversity is the predominant factor in their choice. Thus, many of the disturbance effects and disturbed ecosystems discussed in this volume ﬁnd expression on military lands. For example, lands for training in ground combat within the United States managed by the Army, Marine, and Army National Guard occur in lands dominated by boreal forest (13.2%), chaparral–oak woodlands (1.3%), eastern deciduous forest (10.4%), grasslands (5.0%), mesquite grasslands (0.3%), montane woodland brush (10.7%), northern desert (12.4%), northern hardwood–conifer forest (4.5%), oak savanna (1.3%), Paciﬁc rainforest (0.9%), pinyon–juniper–oak woodland (2.5%), southeast evergreen forest (11.8%), southern desert scrub (25.5%), and tropical vegetation (0.2%) (Smith, 1986; Evinger, 1995). This chapter reviews the mechanisms and the impacts associated with military activities. Emphasis is placed on military training activities because these are the impacts most frequently studied and documented. To understand the impacts of military disturbance better, we describe the principal mechanism of disturbance, the training exercise. Military impacts on soil, hydrology, plants, and animals and the larger concepts of community and ecosystem are reviewed. Management efforts to minimize and mitigate environmental impacts are described and examples are presented.

TRAINING DISTURBANCE REGIMES

The importance of both spatial and temporal factors and their interaction in characterizing land-disturbance regimes has long been recognized (Moloney and Levin, 1996). The temporal components include both frequency and return interval. Spatial components include distribution, and area or size. Additionally, factors of magnitude, such as intensity and severity, along with the occurrence of other disturbances, are important to consider in the relationships between the temporal and spatial factors (Pickett and White, 1985). Disturbance regimes derived from military activities are likely

DISTURBANCE ASSOCIATED WITH MILITARY EXERCISES

to include several agents of disturbance including intense vehicle trafﬁc, explosive munitions, and spills of petroleum, oil, and lubricants (Conrad et al., 1994). These disturbances are rarely independent of each other and often occur as part of a single military event (Department of the Army, 1988b, 1991). Consequently, deﬁning military disturbance would include several regimes. Although not standard military terminology, ﬁve regimes that can be recognized are mechanizedmaneuver training, infantry-maneuver training, liveﬁre training, command and support, and combat engineering. The mechanized maneuver regime consists of unit exercises designed to simulate actual mechanized combat situations (Department of the Army, 1984). These exercises usually occur in open terrain with slopes less than 20% (Krzysik, 1994). Maneuvering with tracked vehicles is the major agent of impact on training lands (Conrad et al., 1994). The primary environmental disturbances caused during mechanized maneuvers are disturbance to the soil and vegetation from vehicle tracks. Depending on environmental conditions, tracked-vehicle trafﬁc can result in soil compaction, comminution of surface particles, and upheaval, crushing and/or uprooting of vegetation (Krzysik, 1994; Thurow et al., 1995; Wilson, 1988). Constructed defenses often include excavations deep enough to accommodate an armored vehicle. Anti-tank ditches are wide, deep trenches designed to prohibit vehicle crossing. Excavations are usually ﬁlled in at the conclusion of an exercise (Department of the Army, 1988c,d). Infantry-maneuver training. Infantry units are trained to ﬁght in dispersed formations (Department of the Army, 1986) where a platoon of roughly 30 to 40 soldiers would have a frontage of 100 to 150 m in offensive actions and 200 m in defensive actions. Common to these training exercises are cutting vegetation for camouﬂage and digging fox-holes, which usually must be reﬁlled. Live-ﬁre training. Live-ﬁre weapon systems include riﬂes, tank guns, anti-tank missile systems, and selfpropelled and towed howitzers, multiple-launch rocket systems, and mortars. Munitions vary with the system, but include high explosives, chemical obscurants, ball ammunition, illumination, and kinetic penetrating antiarmor projectiles (Department of the Army, 1993). The large quantities of these munitions expended over an extended period of time can lead to potential environmental contamination (Getz et al., 1996). Impact of

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munitions can also cause soil displacement, damage to vegetation, ﬁres, incidental killing of wildlife, and changes in wildlife behavior (Severinghaus et al., 1980; Gese et al., 1989; Dinkins et al., 1992; Tazik et al., 1992a). The result of repetitive use of a limited number of ﬁring points can best be described as similar to the damage to soils and vegetation occurring on a construction site. The command and support regime is a conglomeration of units which include headquarters, supply, maintenance, and other non-combat units in static or semi-static bivouac positions. Environmental damage from bivouac positions can vary with the type and size of unit, but usually consists of compacted soils, loss of lower vegetation layers, damage to trees (Trumbull et al., 1994), and excavation for fortiﬁcation of the positions. The engineering regime represents the activities of combat engineering units, both in conjunction with combat units and separately (Department of the Army, 1988a). Major activities include obstacle development and destruction, gap-crossing, and construction of emplacements which involve earth moving. Engineers often use explosives for demolitions and related work (Department of the Army, 1989, 1993). Temporal and spatial considerations The spatial characteristics of training activities vary among the disturbance regimes. For example, a heavily mechanized infantry battalion including 100 tracked vehicles requires up to 24 800 ha to conduct a “move to contact” exercise and 13 800 ha for a “defensive operation”. A light infantry company made up of 108– 120 soldiers requires 7000 ha for a “move to contact” exercise and 1600 ha for a “defensive operation”. Additionally, topography, vegetation, and geographic shape inﬂuence the types of training activities. Wooded and hilly areas are used for bivouac and other static activities. Flat or rolling grasslands and similar open terrain is often used for mechanized-maneuver activities. Given these considerations, the location of activities within an area is often predetermined by topography and vegetation conditions. Temporal features of military training are controlled by a complex mixture of variables, including available land, topography, and vegetation. Impacts from training in any particular area depend greatly on the types of units stationed there. Most United States training areas large enough to support mechanized-maneuver

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training are located west of the Mississippi River, and have the potential of being impacted by the most environmentally damaging of military training activities (Conrad et al., 1994). The land base available for training activities has a strong inﬂuence on intensity and frequency of use; installations with a high concentration of military units per unit area of suitable land, as in some European countries, often experience high intensity of use.

EFFECTS OF TRAINING DISTURBANCE REGIMES

Immediate effects versus long-term cumulative effects Disturbance effects from military activities may occur in the short term, but may often persist for decades (Wilshire and Nakata, 1976; Webb and Wilshire, 1980; Lathrop, 1983; Prose, 1985) especially in ecosystems with low productivity. Cumulative effects are common in military training because of the insular nature of training facilities (surrounded by land not subject to military impact). Maneuvers (e.g., vehicle and troop movement, camping) occur repeatedly on the same sites (Trumbull et al., 1994). For instance, relatively minor short-term trampling events can be repeated over many years, causing a cumulative detrimental impact to soils and plants (Trumbull et al., 1994). Such long-term use of sites without rest may limit the potential for recovery and its rate, and ultimately impact small mammal and avian populations dependent upon speciﬁc vegetation and soil conditions. For instance, Severinghaus et al. (1980) determined that burrowing mammals were very sensitive to soil changes induced by maneuvers. Disturbance effects are so variable in their dependence on season, ecosystem, and substrate that generalizations concerning the short-term or cumulative nature of impacts are difﬁcult. For instance, in Canadian prairie the season of impact was found to be a more important determinant of disturbance than the number of vehicle passages (Wilson, 1988). Certainly, some types of disturbance will impact plants, animals, and their habitat so as to change community composition and modify the physical characteristics of the habitat. Of course, other species will take their place unless the disturbance is severe. The desirability or acceptability of these changes needs to be assessed on a case-bycase, site-by-site basis.

Soil effects Disturbance of soil structure is a common consequence of military training (Goran et al., 1983; Diersing and Severinghaus, 1984; O’Neil et al., 1990; Pearson et al., 1990). Erosion potential is increased on disturbed soils (Diersing and Severinghaus, 1984; O’Neil et al., 1990; Pearson et al., 1990; Trumbull et al., 1994) and inﬁltration rate is decreased (Trumbull et al., 1994). Compaction is a well-studied consequence of soil disturbance (Becher, 1985). Compaction varies with the moisture content at the time of impact, parent material, vegetation type, and the characteristics of the vehicle (United States Bureau of Land Management, 1980; Adams et al., 1982; Becher, 1985; Thurow et al., 1993). Timing of training exercises relative to ecosystem moisture regimes therefore may be expected to inﬂuence observed effects (Wilson, 1988; Thurow et al., 1993). Compaction effects are more pronounced and longer-lasting when vehicles pass over wet soil (Thurow et al., 1993). Compaction can reduce aeration and inhibit root growth, nutrient uptake, and seedling emergence (Chancellor, 1977). Thus, plant-community effects and recovery time are related to the extent of compaction (Webb and Wilshire, 1980; United States Bureau of Land Management, 1980; Prose, 1985; Thurow et al., 1993). Potential for wind erosion generally increases on soils subject to vehicle passages, especially when they are dry. Wind was cited as a factor at such dissimilar locations as Fort Bliss in the Chihuahuan Desert of New Mexico (Marston, 1986; Gillespie, 1987) and at Fort Lewis in the Paciﬁc Border Province of Washington (Pearson et al., 1990). After the Gulf War, wind erosion increased where moving vehicles had exposed ﬁne substrate materials after breaking through the “desert pavement” (a layer of pebbles and cryptogamic crust left behind after ﬁne materials have blown away) (El-Baz, 1992). Wind erosion causes dune formation and dust storms (Krzysik, 1985; McDonald, 1995). Heavy metals and other contaminants have been deposited into soils because of training exercises (Peters and Miller, 1993; Hinsenveld, 1995; Freese and Riesbeck, 1995) and because of activities collateral to military missions such as oil-well ﬁres (Sadiq et al., 1992) and construction (Hagarty et al., 1993). These contaminants have the potential to enter food webs.

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Hydrology effects Water inﬁltration tends to be lowered in compacted sites, runoff thereby being increased and erosion accelerated (Hillel, 1980; Prose, 1985). However, prairie soils at Fort Hood, Texas, were compacted only when vehicles passed over wet soils (Thurow et al., 1993). Decreased inﬁltration rates appeared to promote dominance by early-successional plants, even after some soil physical properties (e.g., bulk density) had recovered to their pre-disturbance condition (Thurow et al., 1993). Removal of vegetation and damage to soil causes increases in runoff rates and the amount of transported sediment by lowering ﬁeld capacity (e.g., Riggins et al., 1989). This change increases the frequency and severity of ﬂooding, including ﬂash ﬂoods, and causes siltation. Somewhat paradoxically, imprints from the treads of tracked vehicles store water, thereby lowering erosion potential for a time (Riggins et al., 1989). Effects on plants Effects on plants include direct crushing and killing; which ultimately causes reduced stem density (Trumbull et al., 1994). Root systems are directly impacted by vehicles, as a result of alteration of bulk density and the rate of water inﬁltration (Trumbull et al., 1994). These effects can produce site characteristics inconsistent with the autecological requirements of the species, and may reduce individual vigor. The relative cover of dominant plant species remained altered in the Mojave Desert 36 years after a military maneuver (Lathrop, 1983). In the same ecosystem, training activities have been found to disturb vegetation by reducing the density and cover of creosotebush (Larrea tridentata) and other plants (Krzysik, 1985). Cool-season grasses and warm-season forbs replaced warm-season grasses in locations subject to trackedvehicle impact over two years at Fort Carson, Colorado, apparently because of competitive interactions associated with spring rain events (Diersing and Severinghaus, 1984; Shaw and Diersing, 1990). Also, woody plants were damaged and killed, allowing increased cover of undesirable and disturbance-related snakeweed (Gutierrezia sarothrae) and kochia (Kochia scoparia) (Diersing and Severinghaus, 1984). Pitting (forming small basins or pits in the soil to catch and hold water and store moisture for plant use) promoted

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the establishment of cool-season grasses and forbs to reduce erosion potential on sites subject to heavy vehicle impact (Berlinger and Cammack, 1990). Wind erosion can transport seeds, expose roots, and decrease soil fertility (Marston, 1986). Ultimately, these effects would result in establishment of weedy invaders and habitat perturbation [permanent modiﬁcation of a habitat from one condition to another: White and Pickett (1985)]. Effects on animals Physical modiﬁcation of habitat resulting in changed levels of available resources is the primary disturbance affecting vertebrate populations on military installations. Modiﬁcations include clearing of woodland and understory, the mixing or removal of soil and detritus, modiﬁcation or removal of food resources, and general degradation of habitat (Severinghaus et al., 1980; Severinghaus and Severinghaus, 1982). Some animal species are adversely affected, some beneﬁt, and others are not impacted (e.g., O’Neil et al., 1990; Pearson et al., 1990). This habitat modiﬁcation can lead to species replacement (Diersing and Severinghaus, 1984), for example, by an increase in abundance of early-sere species at the expense of uncommon climax species and endemics. Severinghaus et al. (1980) found that small mammals, particularly those associated with the soil surface and sub-surface, were adversely affected by maneuver activities through clearing and compacting of the soil, vegetation disturbance, and resultant erosion. Moderate habitat modiﬁcation can increase the abundance of white-footed mice (Peromyscus leucopus), possibly in response to invasion by weedy, early-successional forbs (Diersing and Severinghaus, 1984). In the Mojave Desert, loss of shrub cover was related to lower relative abundance of the little pocket mouse (Perognathus longimembris) and southern grasshopper mouse (Onychomys torridus) (Krzysik, 1985). Birds are generally adversely affected by activities related to maneuvers. The frequency with which 17 out of 19 bird species were observed was lower on shortterm and long-term training areas compared to a control area at Fort Knox, Kentucky (Severinghaus et al., 1980). Bird species richness in woodlands decreased at Fort Carson, Colorado, following disturbance (Tazik, 1991). Habitat modiﬁcation from woodland to open woodland or forest-edge communities initially was beneﬁcial or neutral for some bird species (edge specialists,

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seed-eaters) whereas continued habitat modiﬁcation had an adverse effect on all species (Severinghaus et al., 1980). At the National Training Center in the Mojave Desert, California, disturbance (removal of shrub cover) had an adverse effect on sparrows (Amphispiza belli, A. bilineata, Spizella breweri, S. passerina and Zonotrichia leucophrys), meadowlarks (Sturnella neglecta), and thrashers (Toxostoma lecontei) (Krzysik, 1985). Bird biomass decreases as training activities increase, except in mature forested habitats where large trees tend to be resistant to maneuver-induced damage (Severinghaus and Severinghaus, 1982). Prairie species seem to be more resilient to disturbance impacts than woodland species (Tazik, 1991), perhaps because they are adapted to a climax habitat with less vertical structure and complexity. Changes in the physical structure of the habitat may have an impact on animal populations. In anaerobic marsh sediments at Fort Richardson, Alaska, mortality of waterfowl (Anas sp. and Cygnus sp.) was linked to ingestion of white phosphorus, which resists environmental breakdown (Racine et al., 1992). Blowing dust can be detrimental to desert fauna (Marston, 1986; El-Baz, 1992), but long-term impacts on animal populations and communities are not known. Community effects Species replacement (loss of one species while gaining another with no apparent change in net diversity) is an important issue on military lands (Severinghaus and Severinghaus, 1982; Diersing and Severinghaus, 1984), particularly in regard to climax species. Replacement of these frequently rare or sensitive species by common mid- and early-successional species reduces landscape diversity (Scott et al., 1996). For example, earlysuccessional annuals replaced perennial grasses at Fort Hood, Texas (Johnson, 1982; Thurow et al., 1993). Exotics invaded sites subject to vehicle impact during spring in Manitoba (Wilson, 1988). Generally, vehicle impact decreases cover and allows invasion by annual grasses and forbs. Disturbance of climax-type vegetation frequently increases biodiversity (e.g., Lathrop, 1983). These changes can last for many years (Lathrop, 1983). However, continued periodic disturbance ultimately will degrade habitats and their ability to support biodiversity. For instance, at Fort Leonard Wood, Missouri, plant species richness decreased on sites subject to periodic encampments for 20–40 years (Trumbull et al.,

1994). These observations are consistent with Connell’s (1978) intermediate-disturbance hypothesis. Ecosystem effects Military bases frequently are virtual islands of relatively natural habitat surrounded by properties with a variety of contrasting land uses (Tennesen, 1993; Creswell, 1994; Goodman, 1996). The potential for disturbance to these regional ecosystems is of four types: (1) impacts on endemic, particularly threatened and endangered, biota (Creswell, 1994); (2) habitat fragmentation allowing exotic or weedy species to become established (Wilson, 1988); (3) impacts which move beyond the boundaries of the base (McDonald, 1995); and (4) cases where the protection provided by the base results in it being the last and best example of the native ecosystem and its habitats in an otherwise highly developed area (e.g., the Mediterranean ecosystem of southern California). Toxic contaminants may disturb surrounding ecosystems through escape from the conﬁnes of a local military base. Substantial effort has been expended in research on methods for immobilizing and cleaning contaminants (Hagarty et al., 1993; Peters and Miller, 1993). Heavy metals and other contaminants deposited into water and soils purposely in the form of spent munitions or accidentally by spills, holding ponds, or ﬁre (Deneke et al., 1975; Sweazy et al., 1977; Machin and Ehresmann, 1985) have the potential to enter food webs (Sweazy et al., 1977; Peters and Miller, 1993; Genskow, 1994; Hinsenveld, 1995; Richter and Franke, 1995; Freese and Riesbeck, 1995). This impact may occur through the contaminants being taken up by plants, or through transport in ground and surfacewater (Peters and Miller, 1993; Freese and Riesbeck, 1995).

MANAGEMENT EFFORTS

Regulatory context The United States military has become increasingly challenged by a dramatic expansion in environmental compliance requirements (Conrad et al., 1994). During 1980–1990, the code of national environmental regulations doubled to nearly 10 000 pages (Butts, 1990). The Department of Defense manages over 800 installations covering more than 10×106 ha, each of which must

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comply with national, state, and local regulations. The policy of the Department of Defense requires any activities that it controls outside the United States to follow generally accepted environmental standards adopted for similar situations in the United States (Department of Defense, 1991). A variety of United States environmental laws currently exist affecting the nature, frequency, and extent of ground disturbance caused by military operations [see review by Conrad et al. (1994), and also Eckert and Carroll, Chapter 30, this volume]. The Endangered Species Act which requires that adverse impacts on threatened or endangered species and their critical habitat should be avoided, and that programs for their conservation should be put in place, has had the greatest impact of any legislation on military training in the United States. The National Environmental Policy Act (NEPA) requires the Department of Defense to consider the environmental consequences of their actions and to document these considerations. Assessments like those for the National Environmental Policy Act are also required on overseas bases. The Clean Water Act requires control of soil erosion that results in sediment deposition in surface waters. Section 404 of the Act requires permits for dredge and ﬁll operations and the delineation of wetlands. The Resource Conservation and Recovery Act imposes comprehensive regulations on hazardous wastes. The Sikes Act requires that installations should be managed so as to ensure sustained multiple use of natural resources. The Migratory Bird Treaty Act authorizes the Department of the Interior to regulate activities affecting virtually all avian species in the United States, and the United States Fish and Wildlife Service has become aggressive in its enforcement. Recently, the United States Army has modiﬁed training activities in response to a documented loss of migratory waterfowl killed by ingestion of residue from white phosphorus rounds ﬁred into wetlands (Racine et al., 1992). Protection of sections of rivers under the Wild and Scenic Rivers Act has required the modiﬁcation of ﬂight corridors for aircraft. The Environmental Conservation Program of the Department of Defense describes natural resource policy on lands under the control of the Department in the United States, its Territories, trusts, and possessions (Department of Defense, 1996). Integrated plans for natural-resource management are developed in accordance with principles of ecosystem management. Activities of the Department of Defense must promote

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conservation of biological diversity when practicable and consistent with the military mission. The goal of the ecosystem-management program of the Department of Defense is to ensure that military lands support present and future training requirements while preserving and enhancing ecosystem integrity. This goal includes restoring and maintaining native ecosystems, reestablishing and maintaining viable populations of native species, maintaining evolutionary and ecological processes (e.g., disturbance regimes, hydrological process, and nutrient cycles), and managing the sites over time periods compatible with ecosystem dynamics. The challenge lies in balancing these potentially competing resource uses, while recognizing the primacy of the military mission and the goal of ecosystem maintenance. Case studies Several case studies are presented below to illustrate the context, nature, and range of ecosystem management issues faced by military land managers. This discussion is not intended as a comprehensive exposition on the subject. Eglin Air Force Base (Department of the Air Force, 1993) Eglin Air Force Base is located in northwestern Florida, United States. Major ecological associations include sandhills, wetlands and riparian communities, sand pine (Pinus clausa), ﬂatwoods including the Pinus palustris association, pine and mixed hardwood forests, and barrier islands. Eglin is the largest forested military reservation in any country of the North Atlantic Treaty Organization, totaling over 187 500 ha. The primary military mission of Eglin is the development and testing of conventional munitions and sensor tracking systems. Other training activities include ground-troop maneuvers and ﬂight missions. Other uses include the conservation of unique natural resources (of regional, national and international signiﬁcance), recreation (including ﬁshing, hunting, camping, picnicking, etc.), and the development of forest products. Eglin’s natural-resources management plan recognizes several major issues and management concerns. Land-management activities must, to the extent possible, be compatible with the testing and training mission. However, federal laws, such as the Endangered Species Act, mandate certain management actions that may constrain the mission. For example, Eglin encompasses

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the entire geographical range of several species, and has 89 endangered or potentially endangered species. Among these, the red-cockaded woodpecker (Picoides borealis) is an endangered species of some notoriety on military lands throughout the southeastern United States, because of its impact on land-management decisions. Eglin also harbors over one-third of the estimated 2000 ha of remaining old-growth longleaf pine (Pinus palustris) in the southeastern coastal plain. Training activities on Eglin are compatible with the ecosystem-management goals of the Department of Defense and can be achieved with only minor restrictions or mitigation to minimize training-related disturbance. Physical damage from mechanical equipment is infrequent and on a small scale only. A program is in place to close unnecessary roads, correct erosion problems, and restore forested areas to a natural state. Existing old-growth forests are managed so as to perpetuate them. Canadian Forces Base Shilo (Stewart et al., 1987) Canadian Forces Base Shilo is a 39 511 ha facility in Manitoba, Canada. At the base, the armies of Canada and Federal Republic of Germany train in the use of tank and armored personnel carriers on a 21-day rotation during the ﬁve summer months. Each rotation involves 600 troops. The activities and artillery exercises conducted by the Canadians result in damage to plant cover and soils. The major issues of training-related disturbance involve vegetation, soils, and wildlife. Impacts on vegetation and soils include reduction in desirable native species, increased soil compaction, and encroachment of undesirable species such as leafy spurge (Euphorbia esula). Negative impacts on elk (Cervus elaphus) are the primary wildlife concern. Disturbance impacts are mitigated by a variety of approaches. Harrowing, seeding, and fertilizing have been applied to help recovery. Harrowing and a twoyear rest period proved most effective in restoring native species. General range-management practices employed to minimize soil and vegetation damage include use of designated roads only, minimizing sharp turns of vehicles, reducing range ﬁres, rotation of training areas, delay of training start-up until early-season perennials have established growth, and restricting access in areas of recognized ecological value. Mitigation activities for elk have been limited

largely to establishing areas of restricted access and seasonal avoidance of areas where the elk calve and rest, and have their winter range. Hohenfels Combat Maneuver Training Center (Sullivan et al., 1996) Hohenfels Combat Maneuver Training Center occupies 16 200 ha in Bavaria, Germany. The vegetation consists of mixed forest and grasslands in a topography of rolling hills and valleys. Hohenfels has been in operation since 1951, and has been used as a combat maneuver training center (CMTC) since 1989. One of its major training missions currently is training for United Nations operations such as peace-keeping. The Federal Forestry Department has managed CMTC Hohenfels woodlands to maximize environmental protection. Consequently, the CMTC has become a sanctuary for many plant and animal species whose habitats have been lost elsewhere. Training-related disturbance impacts are associated with repeated and prolonged use of heavy vehicles. Impacts include loss of vegetation and associated soil erosion. Hohenfels is home to a variety of threatened and endangered species, making it a site of national, or even European, signiﬁcance. Mitigation and management efforts are designed to support training while conserving environmental resources. Severely damaged areas are rehabilitated using a variety of measures, including reseeding, replanting, building structures for the control of water and sediment, and improving training-area design. Environmentally sensitive areas are designated as offlimits to training. Communication between trainers and environmental managers is fostered by the development of videographic simulation of alternative training-area designs. Semi-dry and dry meadows contain numerous threatened plant species. This habitat is protected by use of native grasses only in the seed mixtures used for reseeding these areas. Spreading of trees and shrubs is controlled by cutting and by sheep grazing. Shoalwater Bay Training Area (Tunstall, 1993) Shoalwater Bay Training Area occupies 270 000 ha of land together with 14 000 ha of mangroves in central coastal Queensland, Australia, 80 km north of the Tropic of Capricorn. Vegetation is a diverse mix of eucalypt forest on hilly and mountainous areas, with eucalypt and paperbark (Melaleuca) woodland on the plains. Mangroves are extensive on marine sediments, while sand dunes are covered with various forms of heath.

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Shoalwater is managed by the Australian Army as a joint service facility for training exercises up to divisional scale, including vehicle maneuvers. Because of its high conservation status, an Australian Commission of Inquiry has recommended that conservation have equal status with military training. Primary training-related disturbances result from engineering activities, movements of tracked vehicles, and ﬁre. Defensive positions include construction of tank traps and other earthworks, sandbag and timber constructions, vegetation clearing, felling trees for obstacles, barbed wire emplacements, and road demolition. Vehicle trafﬁc results in soil compaction and denudation. Lands adjacent to major camps and defensive positions are typically cleared of woody vegetation using bulldozers and hand tools. Maneuvers by tracked vehicles typically are limited to areas adjacent to roads and tracks, but result in crushing vegetation, soil compaction, and soil displacement. Fire is a natural disturbance event, the frequency of which has increased as a result of training activities. Its unpredictability and potentially dramatic effects make ﬁre a major problem. Although much of the current vegetation (except mangroves) is ﬁre-adapted, an unnaturally high ﬁre frequency could have detrimental effects. Additional damage results from camp sites, tracked vehicles, bombing, naval demolition, and timber harvest. Mitigation and management of disturbance impacts include a rather extensive land-management plan. The plan’s objective is to specify land-management practices appropriate to both military use and conservation. It attempts to contain exercise damage to the minimum necessary to achieve the objectives of the exercise, in order to sustain viability for military use over the long term. Each component of the plan (addressing, for instance, ﬁre, exercises and maneuvers, engineering works, and conservation of the biota) speciﬁes management according to objectives, strategy, priority actions, and performance indicators. The impact of maneuvers involving tracked vehicles is limited by rotational use of training areas. Nontactical movement is limited to roads. Trees are not to be felled indiscriminately, wet soils are avoided, and areas are to be rested when extensively damaged. Areas used for defensive exercises are refurbished so as to minimize environmental damage. Tank traps and other earthworks are developed under strict guidelines, pits and trenches are backﬁlled and mounded, rubbish is carted off the maneuver area rather than being buried,

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and all defense stores are recovered. In other tactically signiﬁcant areas, trees are not to be felled, and tracks are graveled where severe rutting and bogging occurs. Fire is managed by construction and maintenance of perimeter ﬁre breaks, use of prescribed burning to protect life and assets, varying the frequency, timing, intensity, and extent of burns to yield a patchy mosaic, and reducing the frequency and extent of prescribed burning to protect vegetation and wildlife.

CONCLUSIONS

Military lands are widespread and diverse, and include virtually every ecosystem. Therefore, the type and scope of their disturbance are nearly limitless. Nonetheless, there are commonalities that link disturbances occurring on military lands. The common features of disturbance in military lands derive from their use for training maneuvers. Disturbance impacts may result from maneuvers involving tracked vehicles, infantry and live ﬁre, and activities of command and support personnel and engineers, among others. Training maneuvers encompass thousands of activities, any of which could produce persistent disturbance. Historically, tracked-vehicle maneuvers have been investigated in most detail, probably because of their visual impact. Less striking, but potentially of equal importance, are the impacts of less visible activities. Activities such as bivouacking and vehicle fueling may have minimal short-term but signiﬁcant long-term cumulative impacts. Although a variety of land forms are susceptible to military disturbance, it is noteworthy that maneuvers tend to occur with high frequency in particular favored land forms. Conditions of slope, soil composition, vegetation, and drainage interact to concentrate activities in some areas while adjacent areas remain unaffected. Maneuvers may impact soils and vegetation, thereby altering physical properties, hydrology, structure, and species composition of the surface and sub-surface. Maneuvers may cause retrogression to early and midsuccessional stages. From an ecosystem perspective, these communities are undesirable because they promote common species at the expense of rare latesuccession species. Ecosystem management increasingly is becoming the context for management of military land. In this context, species replacement is an important issue on military lands. As a rule, access to military land is

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controlled and development is carefully planned. Thus, military lands frequently support the last and best examples of intact ecosystems, and the importance of their contribution to regional ecosystem management is signiﬁcant. The military’s interest in maintaining ecosystems and their processes goes beyond regulatory necessity. Realistic training potential is important, and cannot occur on severely degraded lands. Nonetheless, the area available is usually ﬁxed, with minimal potential for acquisition of new land. Thus, training occurs repeatedly at the same locations. To prevent severe degradation of these locations, restoration-oriented research and management are required. Training impacts and resultant ecosystem restoration occur over large spatial and temporal scales. Thus, documentation of the military’s long-term cumulative impacts will remain a priority. The planning emphasis must remain on longterm management approaches at the landscape and ecosystem level.

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Overseas Installations. Department of Defense, Washington, D.C., 5 pp. Department of Defense, 1996. DoD Instruction 4715.3: Environmental Conservation Program. Department of Defense, Washington, D.C., 16 pp. Department of the Air Force, 1993. Natural Resources Management Plan: Eglin Air Force Base, Florida, 1993 to 1997. Department of the Air Force, Washington, District of Columbia, 81 pp. Department of the Army, 1984. Army Field Manual 25–4: How to Conduct Training Excercises. Department of the Army, Washington, D.C., 179 pp. Department of the Army, 1986. Army Field Manual 7–70: Light Infantry Platoon/Squad. Department of the Army, Washington, D.C., 291 pp. Department of the Army, 1988a. Army Field Manual 5–100: Engineer Combat Operations. Department of the Army, Washington, D.C., 166 pp. Department of the Army, 1988b. Army Field Manual 25–100: Training the Force. Department of the Army, Washington, D.C., 45 pp. Department of the Army, 1988c. Army Field Manual 71–1: Tank and Mechanized Infantry Company Team. Department of the Army, Washington, D.C., 222 pp. Department of the Army, 1988d. Army Field Manual 71–2: The Tank and Mechanized Infantry Battalion Task Force. Department of the Army, Washington, D.C., 222 pp. Department of the Army, 1989. Army Training and Evaluation Plan 5–145–31: Mission Training Plan for the Engineer Company. Department of the Army, Washington, D.C., 220 pp. Department of the Army, 1991. Army Training Circular 25–1: Training Land. Department of the Army, Washington, D.C., 27 pp. Department of the Army, 1993. Army Pamphlet 350–38: Standards in Weapons Training. Department of the Army, Washington, D.C., 188 pp. Diersing, V.E. and Severinghaus, W.D., 1984. The Effects of Tactical Vehicle Training on the Lands of Fort Carson, Colorado – An Ecological Assessment. United States Army Construction and Engineering Research Laboratory, Technical Report N-85/03, 46 pp. Diersing, V.E. and Severinghaus, W.D., 1985. Wildlife as an Indicator of Site Quality and Site Trafﬁcability During Army Training Maneuvers. United States Army Construction and Engineering Research Laboratory, Technical Report N-86/03, 19 pp. Dinkins, W.C., Lochmiller, R.L., Bartus, W.S., DeYoung, C.A., Qualls Jr., C.W. and Fulton, R.W., 1992. Cause-speciﬁc mortality of white-tailed deer as inﬂuenced by military training activities in southwestern Oklahoma. J. Wildl. Dis., 28: 391–399. El-Baz, F., 1992. Preliminary observations of environmental damage due to the Gulf War. Nat. Resources Forum, 16: 71–75. Evinger, W.R., 1995. Directory of U.S. Military Bases Worldwide. Oryx Press, Phoenix, Arizona, 412 pp. Freedman, B., 1989. Environmental Ecology: The Impacts of Pollution and Other Stresses on Ecosystem Structure and Function. Academic Press, London. Freese, D. and Riesbeck, F., 1995. Reclamation of military areas polluted with traces of ammunition and heavy metals, using a new technology and subhydric sediments. In: W.J. van den Brink, R. Bosman and F. Arendt (Editors), Contaminated Soil. Kluwer Academic Publ., Netherlands, pp. 1207–1208.

DISTURBANCE ASSOCIATED WITH MILITARY EXERCISES Genskow, K.D., 1994. Nonpoint Source Pollution: Implications of Clean Water Act Revisions on Army Combat Training and Land Management. Master’s Thesis, Univ. Illinois at Urbana-Champaign, 89 pp. Gese, E.M., Rongstad, O.J. and Mytton, W.R., 1989. Changes in coyote movements due to military activity. J. Wildl. Manage., 53: 334–339. Getz, L.L., Reinbold, K.A., Tazik, D.J., Hayden, T.J. and Cassels, D.M., 1996. Preliminary Assessment of the Potential Impact of Fog Oil Smoke on Selected Threatened and Endangered Species. United States Army Construction and Engineering Research Laboratory, Technical Report 96/38, 44 pp. Gillespie, B.M., 1987. The Impact of Military Maneuvers on Eolian Transport and Soil Compressive Strength in South Central New Mexico. Master of Arts Thesis, University of Wyoming, Laramie, 178 pp. Goodman, S.W., 1996. Ecosystem management at the Department of Defense. Ecol. Appl., 6: 706–707. Goran, W.D., Radtke, L.L. and Severinghaus, W.D., 1983. An Overview of the Ecological Effects of Tracked Vehicles on Major US Army Installations. United States Army Construction and Engineering Research Laboratory Technical Report N-142, 75 pp. Gradwohl, J. and Greenberg, R., 1988. Saving the Tropical Forests. Earthscan Publications, London, 214 pp. Hagarty, E.P., Dee, P.E., Kikkeri, S.R. and Wilcher, J.L., 1993. Remediation of contaminated soil at a military installation site. In: J. Hager, B. Hansen, W. Imrie, J. Pusatori and V. Ramachandran (Editors), Extraction and Processing for the Treatment and Minimization of Wastes. The Materials and Minerals Society, pp. 441–459. Hillel, D., 1980. Fundamental Soil Physics. Academic Press, New York, 486 pp. Hinchman, R.R., McMullen, K.G., Carter, R.P. and Severinghaus, W.D., 1990. Rehabilitation of Military Tracked Vehicles at Fort Carson, Colorado. U.S. Army Construction and Engineering Research Laboratory, Technical Report N-91/01, 62 pp. Hinsenveld, M., 1995. Remediation strategies for contaminated (former) military sites. In: W.J. van den Brink, R. Bosman and F. Arendt (Editors), Contaminated Soil. Kluwer Academic Publishing, Netherlands, pp. 97–98. Johnson, F.L., 1982. Effects of tank training activities on botanical features at Fort Hood, Texas. Southwest. Nat., 27: 309–314. Krzysik, A.J., 1985. Ecological Assessment of the Effects of Army Training on a Desert Ecosystem: National Training Center, Fort Irwin, California. United States Army Construction and Engineering Research Laboratory, Technical Report N-85/13, 139 pp. Krzysik, A.J., 1994. Biodiversity and the Threatened/Endangered/ Sensitive Species of Fort Irwin, California. United States Army Construction and Engineering Research Laboratory Technical Report EN-94/07, 114 pp. Lanier-Graham, S.D., 1993. The Ecology of War: Environmental Impacts of Weaponry and Warfare. Walker, New York, 185 pp. Lathrop, E.W., 1983. Recovery of perennial vegetation in military maneuver areas. In: R.H. Webb and H.G. Wilshire (Editors), Environmental Effects of Off-Road Vehicles; Impacts and Management in Arid Regions. Springer-Verlag, New York, pp. 266–277.

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Machin, J.L. and Ehresmann, J., 1985. Remediating a ﬁre site. Civ. Eng., 60: 55–56. Marston, R.A., 1986. Maneuver-caused wind erosion impacts, south central New Mexico. In: W.G. Nickling (Editor), Aeolian Geomorphology. Proc. 17th Annual Binghamton Geomorphology Symp., 17: 273–290. McDonald, K.W., 1995. The Effects of Military Maneuvers on Soil Structure Breakdown and Wind Erosion at the National Training Center, Fort Irwin, California. Master of Science Thesis, Western Kentucky University, Bowling Green, 63 pp. Moloney, K.A. and Levin, S.A., 1996. The effects of disturbance architecture on landscape-level population dynamics. Ecology, 77: 375–394. O’Neil, L.J., Waring, M.R., Hughes, H.G., Landin, M.C., Pearson, M.L., Morris, P.A. and Larson, R.J., 1990. Proposed 9th Infantry Division Force Conversion; Maneuver Damage, Erosion and Natural Resources Assessment Yakima Firing Center, Washington. United States Army Waterways Experiment Station, Technical Report EL-90–9, 114 pp. Owens, S., 1990. Defense and the environment: the impacts of military live ﬁring in national parks. Cambridge J. Econ., 14: 497–505. Pearson, M.L., Morris, P.A., Larson, R.J., O’Neil, L.J., Waring, M.R., Hughes, H.G. and Ladin, M.C., 1990. Proposed 9th Infantry Division Force Conversion; Maneuver Damage, Erosion and Natural Resources Assessment Fort Lewis, Washington. United States Army Waterways Experiment Station, Technical Report GL90–13, 115 pp. Peters, R.W. and Miller, G., 1993. Remediation of heavy metal contaminated soil using chelant extraction: feasibility studies. Proc. 48th Industrial Waste Conf., 48: 141–167. Pickett, S.T.A. and White, P.S., 1985. The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, San Diego, California, 472 pp. Prose, D.V., 1985. Persisting effects of armored military maneuvers on some soils of the Mojave Desert. Environ. Geol. Water Sci., 7: 163–170. Racine, C.H., Walsh, M.E., Roebuck, B.D., Collins, C.M., Calkins, D., Reitsma, L., Bucjli, P. and Goldfarb, G., 1992. White phosphorus poisoning of waterfowl in an Alaskan salt marsh. J. Wildl. Dis., 28: 669–673. Richter, M. and Franke, C., 1995. Distribution and mobilisation of nitroaromatic compounds in a former military shooting area. In: W.J. van den Brink, R. Bosman and F. Arendt (Editors), Contaminated Soil. Kluwer Academic Publishing, Netherlands, pp. 411–412. Riggins, R.E., Holge, W., Lacey, R.M. and Ward, T.J., 1989. Sediment Control at Army Training Areas Case Study: Hohenfels, Federal Republic of Germany. United States Army Construction and Engineering Research Laboratory Technical Report N-89/08, 25 pp. Sadiq, M., Al-Thagaﬁ, K.M. and Mian, A.A., 1992. Preliminary evaluation of metal contamination of soils from the Gulf Activities. Bull. Environ. Contamination Toxicol., 49: 633–639. Scott, J.M., Ables, E.D., Edwards Jr., T.C., Eng, R.L., Gavin, T.A., Harris, L.D., Hauﬂer, J.B., Healy, W.M., Knopf, F.L., Torgerson, O. and Weeks Jr., H.P., 1996. Conservation of biological diversity: perspectives and the future for the wildlife profession. Wildl. Soc. Bull., 23: 646–657.

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Severinghaus, W.D. and Severinghaus, M.C., 1982. Effects of tracked vehicle activity on bird populations. Environ. Manage., 6: 163–169. Severinghaus, W.D., Riggins, R.E. and Goran, W.D., 1980. Effects of tracked vehicle activity on terrestrial mammals and birds at Fort Knox, Kentucky. Trans. K. Acad. Sci., 41: 15–26. Shaw, R.B. and Diersing, V.E., 1990. Tracked vehicle impacts on vegetation at the Pinon Canyon Maneuver Site, Colorado. J. Environ. Qual., 19: 234–243. Smith, R.L., 1986. Elements of Ecology. Harper and Row, New York, 677 pp. Stanford Biology Study Group, 1971. Destruction of Indochina. In: J.P. Holdren and P.R. Ehrlich (Editors), Global Ecology: Readings Towards a Rational Strategy for Man. Harcourt, Brace, Jovanovich, New York, pp. 146–154. Stewart, J.A., Downs, A.T. and Stones, G.A., 1987. The impact of military training in Canada on indigenous ﬂora and fauna. In: Proceedings: NATO CCMS Seminar Blue Book 159, Preservation of Flora and Fauna in Military Training Areas, Conference Proceedings N-87/09. US Army Construction Engineering Research Laboratory, pp. 107–124. Sullivan, R.G., Hatton, P.J. and Boehm, A., 1996. Environmental Management Programs at Combat Maneuver Training Center Hohenfels. Multimedia CD-ROM. Argonne National Laboratory, Argonne, Illinois. Sweazy, R.M., Rose, F.L. and Baugh, C.L., 1977. Toxic Effects of Military Wastewater Efﬂuent. Water Resources Center, Texas Tech University, Lubbock, Texas, 71 pp. Tazik, D.J., 1991. Effects of Army Training Activities on Bird Communities at the Pinon Canyon Maneuver Site, Colorado. United States Army Construction and Engineering Research Laboratory Technical Report N-91/31, 113 pp. Tazik, D.J., Cornelius, J.D., Herbert, D.M., Hayden, T.J. and Jones, B.R., 1992a. Biological Assessment of the Effects of Military Associated Activities on Endangered Species at Fort Hood, Texas. Special Report EN-93/01, 139 pp. Tazik, D.J., Warren, S.D., Diersing, V.E., Shaw, R.B., Brozka, R.J., Bagley, C.F. and Whitworth, W.R., 1992b. U.S. Army Land

Condition-trend Analysis (LCTA) Plot Inventory Field Methods. United States Army Construction and Engineering Research Laboratory Technical Report N-92/03, 62 pp. Tennesen, M., 1993. Can the military clean up its act? Natl. Wildl., 31: 14–19. Thomas, W., 1995. Scorched Earth: The Military’s Assault on the Environment. New Society Publishers, Philadelphia, Pennsylvania, 227 pp. Thurow, T.L., Warren, S.D. and Carlson, D.H., 1993. Tracked vehicle trafﬁc effects on the hydrologic characteristics of central Texas rangeland. Trans. Am. Soc. Agric. Eng., 36: 1645–1650. Thurow, T.L., Warren, S.D. and Carlson, D.H., 1995. Tracked Vehicle Trafﬁc Effects on the Hydrologic Characteristics of Central Texas Rangeland. United States Army Construction and Engineering Research Laboratory Tech. Man. EN-95/02, 10 pp. Trumbull, V.L., Dubois, P.C., Brozka, R.J. and Guyette, R., 1994. Military camping impacts on vegetation and soils of the Ozark Plateau. J. Environ. Manage., 40: 329–339. Tunstall, B., 1993. Environmental Impact Assessment – Shoalwater Bay Training Area (Draft). Commonwealth Scientiﬁc and Industrial Research Organization, Division of Water Resources, Canberra, Australian Capital Territory, 81 pp. United States Bureau of Land Management, 1980. The Effects of Disturbance on Desert Soils, Vegetation, and Community Processes with Emphasis on Off-road Vehicles: A Critical Review. Riverside, California, Desert Planning Staff, 190 pp. Webb, R.H. and Wilshire, H.G., 1980. Recovery of soils and vegetation in a Mojave Desert ghost town, Nevada, U.S.A. J. Arid Environ., 3: 291–303. White, P.S. and Pickett, S.T.A., 1985. Natural disturbance and patch dynamics: an introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, San Diego, California, pp. 3–13. Wilshire, H.G. and Nakata, J.K., 1976. Off-road vehicle effects on California’s Mojave Desert. Calif. Geol., 29: 123–132. Wilson, S.D., 1988. The effects of military tank trafﬁc on prairie: a management model. Environ. Manage., 12: 397–403.

Chapter 16

DISTURBANCE IN URBAN ECOSYSTEMS Herbert SUKOPP and Uwe STARFINGER

INTRODUCTION

Urban ecosystems differ from natural or rural ones in many obvious ways. Human activities, such as building, trafﬁc, or industrial production affect the quality of air, water, and soil which impacts ecosystems in many ways. Plants can be destroyed, their production reduced, animals can get killed or scared away. The results are altered population dynamics, species composition, and energy and matter ﬂuxes in urban ecosystems. In plant ecology, disturbance is frequently deﬁned as a mechanism that limits plant biomass by causing its partial or total destruction (Grime, 1979). In order to include disturbance effects on ecosystems and on animals, we follow White and Pickett (1985) who deﬁne disturbance as a relatively discrete event in time that disrupts ecosystem, community, or population structure. This can include the killing, displacement, or damaging of individuals, and create an opportunity for new individuals to become established (Sousa, 1984). The study of urban ecosystems is a relatively recent phenomenon in ecology, because most ecologists have been, and still are, interested mostly in natural ecosystems. Urban ecology as a scientiﬁc discipline is being practiced in Europe (mainly western and central Europe) more than in other regions of the world. It was the subject of a few recent books, such as those of Sukopp et al. (1990, 1995), both with an international perspective; those by Gilbert (1991) and Sukopp and Wittig (1993) are more strongly focused on European countries. In North America the ecology of urban forests has been studied in more detail (Rowntree, 1984, 1986). As a consequence, the majority of our data, examples, and references, concern cities in western and central Europe. In this paper, we want to show how the multitude

of human activities inﬂuences ecological conditions in cities, and how ﬂora and vegetation react. After describing some typical urban habitats and their disturbance regimes, we will give accounts of plant and animal species in cities and their successional changes, in order to discuss these data in light of general approaches to naturalness and the degree of disturbance.

DISTRIBUTION OF URBAN HABITATS

Even though whole cities can be seen as large ecosystems, especially with regard to energy and matter ﬂuxes, structurally and functionally they form complexes of various interconnected ecosystems (Wittig and Sukopp, 1993; Rebele, 1994). Contrary to common expectations, cities may be quite rich in plants and animals, both quantitatively and qualitatively. Urban ecosystems and the composition of urban plant and animal communities are greatly dependent on human activities causing disturbance. The extent of these impacts varies in time and space. Typically, cities show a mosaic of habitats with increasing degrees of human impact on a gradient from the outskirts to city centers, representing different ages or time spans of human impacts (Table 16.1). A common way to study the speciﬁc urban biota and their habitats is by investigation of their history (Aey, 1990) and comparison along rural–urban gradients (McDonnell and Pickett, 1990). Some features of habitats vary clinally with distance from city centers to the suburbs, or with time. The organisms and communities in these habitats react to human inﬂuences in various ways, and are consequently different for each structural unit of the city. Understanding of the distribution of habitat types in cities is

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Table 16.1 Some features of habitat types with different ages and locations on the country–city gradient 1 (examples from Central Europe) Forest

Field

Suburban garden

Inner city block

Climate

forest climate: low daily temperature maxima, high relative humidity

open land climate: higher temperature, lower humidity

mean annual temperature 1.5ºC higher than forest

highest temperature, low humidity

Relief

low, rolling hills

erosion, increased steepness

linear steep banks and dams

ﬂattened

Hydrology

unchanged

increased run-off, higher levels in depressions

lower groundwater levels

lower groundwater levels

Soil

forest soils, brown earth, moist

plowed ﬁeld soils, more nutrients (N, P, K), higher pH and humus content than forest, moist

surface sealing <50%, soils deeply dug up, N, P, K, pH and humus higher than ﬁeld, moist – wet

surface sealing >50%, garden and rubble soils, nutrient rich, compacted, dry – wet

Vegetation

Quercus–Pinus forests

crops, ﬁeld-weed communities

fruit trees, ornamentals, Euphorbia peplus–garden-weed communities

Lolium sward, Aegopodium–Urtica community, Hordeum community

Fauna 2

large mammals, tree bats

hare, partridge, ﬁeld lark, ﬁeld mouse

rabbit, stone marten

house mouse, Norwegian rat, feral pigeon, house sparrow

1

Sukopp (1990). ﬁeld lark, Alauda arvensis; ﬁeld mouse, Microtus spp.; hare, Lepus europaeus; house mouse, Mus musculus; house sparrow, Passer domesticus; Norwegian rat, Rattus norvegicus; partridge, Coturnix coturnix; pigeon, Columba livia; rabbit, Oryctolagus cuniculus; stone marten, Martes foina.

2

an important prerequisite for nature protection and town planning. This information is widely available for many European cities (Starﬁnger and Sukopp, 1994). To understand present urban biota and their ecosystems it is necessary to see them as a result of historical development. In central Europe, the process of redevelopment of a forest vegetation after the last Ice Age was not completed when human inﬂuence began to cause disturbances on a local scale. Large-scale disturbance, however, only began in medieval times with clear-cutting of extensive areas for agriculture. At the same time, human impact on the hydrology of the landscape increased. Historically, towns and cities were almost free of both spontaneously-growing and cultivated plants, due to limited space and the attitude of the inhabitants (Trepl, 1992). Residents of cities have fought nature back to create a cultural, artiﬁcial environment as opposed to the more natural environments prevailing outside (Trepl, 1992). Today, however, cities usually consist of a mixture of densely settled areas in the historic centers, remnants of agro-ecosystems, and even near-natural areas in urban forests, parks, and nature reserves.

Urban forests Even in areas adjacent to cities, and more so within the cities, forests are disturbed by urban activities such as building of houses and roads, recreation, emissions, etc. In a comparison of urban with rural forest stands in southeastern Wisconsin (U.S.A.), Sharpe et al. (1986) found that urban sites were disturbed by houses and yards, dumping, footpaths, and other factors that were not evident in the rural sites. Housing areas The vegetation in housing areas is subjected to catastrophic disturbances every time buildings are demolished and rebuilt. The disturbance regime and, consequently, the species composition of housing areas is closely linked to the age of the housing area. Aey (1990) studied the typical properties of soils and ﬂora in housing estates of different ages in L¨ubeck, Germany. The soils of the oldest parts were richer in humus and in nutrients (nitrogen, phosphorus). The plants of the oldest parts were typically native forest plants, which had high requirements for nutrients and humidity. On the more recently disturbed soils of

DISTURBANCE IN URBAN ECOSYSTEMS

the youngest housing area (~25 years old), the ﬂora contained a high proportion of non-native annuals. Streets The ecology of roads and their verges has been mostly studied in rural areas (Ullmann and Heindl, 1989). Some of the effects of car trafﬁc are similar in urban areas, and often they are more pronounced in cities due to high trafﬁc levels. Major disturbances include earth movements during the construction of streets, soil compaction due to trampling and vehicles, eutrophication and rise of pH values, as well as mowing and herbicide application in existing streets (Fig. 16.1). Salinity is a major factor in the soils of urban streets due to de-icing salt, which is widely used in city streets in climates with cold winters. (Today its use is restricted in many European countries.) Urban waste-land Once land is urbanized, it usually remains in urban use. However, some time may pass between the demolition of old buildings and the construction of new ones. In recent decades, an increasing number of inner-city sites have been left vacant after either industrial or trafﬁc (railway) uses were discontinued due to economic or political reasons. The soils, climate, and water regime of these sites are generally strongly altered by human inﬂuence, but the vegetation can often develop with relatively little disturbance for some time. Landﬁlls Landﬁlls can be distinguished by the presence or absence of organic material (for example, communal garbage-disposal landﬁlls, and construction debris, respectively). The latter resemble urban vacant lots in many respects. The climate of landﬁlls is primarily inﬂuenced by the form of the landﬁll, which may impact temperatures and precipitation. Blowing dust leads to eutrophication and higher pH values in the soil up to a distance of 100 m. Sewage farms The construction and use of sewage farms marked the beginning of large-scale waste-water treatment in the second half of the 19th century. Sewage farms are usually located in the outer suburbs or areas

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adjacent to cities. As more efﬁcient methods of wastewater treatment were developed, the use of most sewage farms was discontinued. While in operation, the soils of these farms were permanently moist and nutrients were added in large quantities. Heavy-metal contamination, however, reached high concentrations, so that cultivation of vegetables had to be stopped. The open landscapes of sewage farms with their mosaic of ﬁelds, maintenance roads, hedges, etc., however, were rich in species. In one operating sewage farm in Berlin, the number of vascular plant species was slightly higher than in an adjacent agricultural landscape (Sukopp, 1990). Mammals and especially birds were present in high species numbers, including species rare and threatened in other parts of the city. The effects on the soil persist for a long time after the sewage-farm operation ceases. DISTURBANCE AND ABIOTIC FACTORS

Urban climate To maintain a functioning city or town, large inputs of materials and energy are needed. The resulting solid, liquid, and gaseous wastes greatly alter the city and its surroundings (cf. Duvigneaud and Denayer-de Smet, 1977). The result is a speciﬁc urban climate, as ﬁrst described by Kratzer (1937), and subsequently detailed by Landsberg (1981), Oke (1987), and Kuttler (1993). Horbert et al. (1983) have described general characteristics of the urban climate, as compared to the surrounding non-urban areas, by using data from Berlin and numerous published data from other cities: (1) Higher air pollution: in the urban climate, gaseous pollution is 5–25 times higher; condensation nuclei are about 10 times more abundant. (2) Altered radiation: the urban climate has 5–15% fewer hours of sunshine; 20–25% less direct solar radiation (even 50% in winter); ~10% less surface albedo; 12% more reﬂected radiation leads to an increased net radiation of 11% at noon or 47% in the evening. (3) Wind speeds are reduced by 10–20%, times without wind are increased by 5–20%. (4) The ecologically most important result of these effects is a raised temperature. The difference between temperatures within and outside a city depends on its size. It can reach 9ºC on clear days in Berlin, and 7ºC in Aachen (western Germany). The yearly mean temperature in cities

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Fig. 16.1. Soil impacts of a street through a forest in the city of Berlin (Blume et al., 1977, modiﬁed).

DISTURBANCE IN URBAN ECOSYSTEMS

in central Europe is 0.5 to 2.0ºC above that of the surrounding countryside. (5) The annual mean precipitation is increased by up to 20% (Berlin). (6) Due to increased temperature, the relative humidity is between 2% (winter) and 10% (summer) lower; on clear days in Berlin this difference can become 30%. Climatic conditions within a city can vary considerably, depending on type of construction, paving, location in the city and, especially, the distance to large vegetated areas. Depending on these effects, different climatic zones that are more or less concentric can be distinguished. The inﬂuence of areas of vegetation on the urban climate were investigated by von St¨ulpnagel et al. (1990). He found a reduction in temperature not only in a green area but also up to 1.5 km away from it. This climatic inﬂuence grew with the size of a green area, but was reduced where the area was divided by a road. Soils Soils of urban areas usually show very heterogeneous qualities, because the human impact in cities adds changes in soil qualities to the natural variation present before the city was built. Typically, this impact increases on a rural–urban gradient: even in areas adjacent to cities, soils may be indirectly inﬂuenced by emissions and air-borne pollutants. In the forests at the outer fringe of the Berlin metropolitan area, pH values of the soil decreased by 1.1 between 1950 and 1981 (Grenzius, 1984). Even if part of this decrease may be natural, it offers evidence for anthropogenic acidiﬁcation of soils. Especially in sandy soils, acidiﬁcation leads to leaching of nutrients. In more central locations in cities the deposition of acidic emissions (SO2 , NOx ) is more than compensated by deposition of dust and fertilization, so that the pH tends to be higher than under natural conditions. Soils of urban forests in New York City were extremely hydrophobic and showed much lower rates of nitrogen mineralization than those of rural sites, suggesting that hydrocarbons may limit the activity of soil microbes. Possible synergistic effects of heavy metal contamination and soil compaction would further reduce the nitrogen mineralization in the urban forest (White and McDonnell, 1988). In the densely built-up parts of the city, many soils are completely destroyed by excavation or are covered

401

with buildings or concrete paving. Consequences of this “surface sealing” for the ecosystem are habitat loss for all plants (except some lichens or mosses) and most soil organisms, and a reduction of groundwater regeneration. The percentage of sealed surfaces in large German cities is between 40 and 60%; individual blocks can have values as high as 98% and most inner city blocks have less than 10% of the area left for vegetation to develop (B¨ocker, 1985). Soil excavation for buildings and compaction by heavy vehicles can destroy soil horizons and mix topsoil with less weathered subsoils. The decomposition processes in the soil of young landﬁlls emit heat. Shortly after the deposition of garbage in German landﬁlls, soil temperatures up to 88ºC were found. Two years later, soil temperatures ranged from 15º to 45ºC (Kunick and Sukopp, 1975). Prominent features of landﬁll soils are a low bulk density (0.2 kg dm−3 to 0.9 kg dm−3 ), and high levels of organic material (20–30%; Blume et al., 1979). The decomposition of the organic fraction leads to the production of elementary nitrogen, carbon dioxide, and later, under completely anoxic conditions, to hydrogen and methane. These gases not only determine the growing conditions for plants on the landﬁll itself but also permeate neighboring sites. In addition, soluble inorganic and organic substances are carried away by moving groundwater. In a study of a landﬁll in Berlin, Blume et al. (1979) showed that lateral diffusion of methane into a neighboring stand of 90-year-old oaks (Quercus robur) damaged and killed trees (Fig. 16.2). The effect reached trees 75 m from the landﬁll, the trees closest to it being usually most heavily damaged, although no continuous gradient was observed. Treering analyses showed that growth reduction only began after the landﬁll was covered with topsoil, which caused soil processes to become anaerobic. Individual trees were killed > 80 m away, presumably as an effect of groundwater contaminated with heavy metals (Blume et al., 1979). Damage to herbaceous vegetation was not detected in this study. Schlenther et al. (1996) reported on extensive growth failures in afforestation on a sewage farm in Berlin. Heavy metals and organic pollutants were present in high concentrations, but did not seem to be sufﬁcient to explain damages to the trees. Instead, water retention capacity, a factor connected to organic matter content of the topsoil, and consequently water availability to plants, was a key factor. Heavy metals, however, may become more of a problem in the future when

402

Herbert SUKOPP and Uwe STARFINGER

Fig. 16.2. Impacts of a landﬁll on the surroundings (Blume et al., 1979, modiﬁed). x-axis: distance from center of landﬁll; y-axis: altitude (m above sea level).

soil development leads to decreased pH values and subsequent higher biological activity in more acidic soils. Two soil types are typical of urban soils: deeply cultivated garden soils with high nutrient content and high water-holding capacity, and soils formed from rubble (e.g., from buildings demolished during war). The latter are alkaline, dry, and well-aerated in the beginning, but develop with humus accumulation (Sukopp et al., 1979b). They are usually low in nitrogen, but have moderate to high contents of phosphorus, calcium, potassium, and other nutrients (Bradshaw and Chadwick, 1980; Blume, 1993). Concentrations of metals, in particular of lead, are often considerably higher in urban areas than those in agricultural soils (Thornton, 1991). Especially in areas of past mining and smelting, metal concentrations may be very high. Although plant toxicity can be observed in some instances, there is little quantitative informa-

tion available on the impacts of metal contaminants in urban soils on plants and animals (Thornton, 1991). The results of all these human impacts on the soil biota in urban areas are poorly known. Generally there seems to be a reduction both in species numbers and biomass of soil organisms. Where soil disturbance and stressed conditions are present at the same time, there may be no permanent residents of the soil at all. In some cases the absence of decomposer organisms will lead to accumulation of litter and poor habitat quality for vegetation (Harris, 1991). Groundwater Groundwater is affected by human activities both quantitatively and qualitatively. Anthropogenic heat sources in buildings, sewage canals, etc., lead to increased temperatures in groundwater near the surface in cities (Balke, 1974; H¨otzl and Makurat, 1981). How the habitat and its communities are affected, however,

DISTURBANCE IN URBAN ECOSYSTEMS

403

Table 16.2 Tree growth of oaks (Quercus robur) under changing groundwater conditions in a city park in Berlin 1 Time

Disturbance

1773–1831

Natural groundwater level

2.2 mm

1832–1901

Groundwater level lowered by 1m

1.1 mm

1902–1945

further lowering, tree cutting

1.4 mm

1946–1978

high groundwater level, additional watering, soil cultivation

3.2 mm

1

Tree-ring width

Sukopp et al. (1979a).

has been little studied. Ecological effects of quantitative changes of the groundwater are better known. In urban areas, there is generally more variation in groundwater level and ﬂow direction than in rural areas (Leuchs and R¨omermann, 1991). Beginning in medieval times, the construction of locks and dams for water mills caused higher groundwater levels, and led to bog formation in the upper reaches of the rivers in Central Europe (Brande, 1986). Later, low-lying areas and bogs were drained to provide new agricultural land. Water usage by growing populations, and the reduction in groundwater regeneration as a result of the sealing of surfaces, began to deplete groundwater levels, a process which accelerated after the Industrial Revolution. Among the consequences of drainage for the vegetation were damage and growth reduction of individual plants, most notably trees (Table 16.2), decline of plant species dependent on groundwater (phreatophytes: Londo, 1976) and changes in the nutrient status of the soils affected. Sandy soils can become more acidic owing to the loss of calcium in the groundwater. In bogs, on the other hand, increased aeration following the lowering of groundwater leads to decomposition and nitriﬁcation of the peat, and higher nutrient availability (Sukopp, 1981). In an environmental impact analysis of a planned motorway through the area of an inner city park, Sukopp et al. (1979a) estimated that 450 trees would die because of the lowering of the groundwater level during the construction phase. Surface water Surface water bodies in urban areas suffer from a

variety of man-induced disturbances that result in artiﬁcial, straightened, and sealed beds of rivers and lakes, an alteration of the hydrodynamics of rivers, interruption of the ﬂow continuum by impoundments, or the complete destruction of individual water bodies by ﬁlling up or by converting natural streams into underground canals. This impedes longitudinal and vertical migrations of animals, and riparian vegetation is often directly destroyed (Schuhmacher, 1991). Main inﬂuences on the water itself are eutrophication and pollution, as well as an increase in wave action due to recreational and commercial trafﬁc on the water. This has led to a sharp decline in reed beds in many European countries (den Hartog et al., 1989).

URBAN BIOTA AND ITS REACTION TO DISTURBANCE

Composition of urban ﬂoras Some disturbance types cause direct damage to plants. The use of salt for clearing snow and ice can have a detrimental effect on trees along urban streets. Auhagen and Sukopp (1980) found that close to 50 000 street trees (~20% of all street trees) in Berlin were threatened by salt; 90% of street tree mortality in Berlin was due to salt (Ruge, 1978). Sugar maple (Acer saccharum), widely planted in the U.S.A. as a roadside tree, is particularly sensitive to salt, and trees within 10 m of salted roads can be killed (Bradshaw and Chadwick, 1980). Shrubs and herbaceous vegetation are also inﬂuenced by salt. High salinities result in open swards and sometimes strips of bare ground, often called “salt burn”, immediately adjacent to streets (Gilbert, 1991). Of the species present in these salt-burned sites, some are known to be salt-tolerant, and in several European countries the invasion of maritime species into upland areas has been reported. Puccinellia distans, originally restricted to coastal salt marshes, is now the most widespread maritime species along roads in Britain (Gilbert, 1991). In Poznan, Poland, the species is dominant along streets, but also colonizes nitrogen-rich sites (Jackowiak, 1982, 1996). Exhaust fumes of cars contain substances (e.g., lead, cadmium) which are found in high concentrations along streets. Wu and Antonovics (1976) found concentrations of lead next to a busy street up to 50 times higher than in a control site. These high concentrations of lead resulted in selection for lead tolerance in populations of

404

Herbert SUKOPP and Uwe STARFINGER

Table 16.3 Examples of threatened species in Berlin and Germany 1 Species

Germany

Berlin Species number

Threatened (%)

Species number

Threatened (%)

Ferns and ﬂowering plants (indigenous and archeophytes)

993

49.2

2728

32.0

Mosses and liverworts

405

75.8

1000

13.9

Mammals

53

54.7

87

49.4

Amphibians

14

78.6

19

57.9

1014

58.5

1300

41.1

Butterﬂies 1

Auhagen (1991).

Plantago lanceolata. High concentrations of lead and cadmium were reported by Blume et al. (1977) from Berlin, where they were restricted to the uppermost layers of the soil (Fig. 16.1). As pH values also increased signiﬁcantly adjacent to the street, mobility of the heavy metals was low and no damaging effects on plants were found (Blume et al., 1977). On account of the intensity of human inﬂuence on urban ecosystems, many species are in danger of becoming extinct or are already extinct. The proportion of threatened species in various groups of organisms is often considerably higher than in areas with lower anthropogenic inﬂuence. In the area of Berlin, more than half of the species of several plant and animal groups are threatened, whereas nationwide this ﬁgure is much lower (Table 16.3). The process of species loss over the past 100 years was described by Drayton and Primack (1996) for a park in metropolitan Boston (U.S.A.). They calculated that the rate of decline was 0.36% yr−1 for the native plant species present in 1894, resulting in a total loss of 155 out of the original 422 species in this area of 400 ha. The changes in abiotic factors in cities described above result in changes in the species composition of plants and animals. Together with the formation of new types of habitat – open disturbed grounds as opposed to the closed forests prevailing in the natural landscape – the ﬁrst set of new plant species arrived. As the native ﬂora had evolved under forested conditions these newcomers, transported deliberately or by chance, proved to be competitive under the conditions now widely present. The overall result of the introduction and naturalization of plants was an increase in the numbers of spontaneously occurring species [J¨ager (1988), for central Europe] which is

especially noticeable in cities. Klotz (1990) found a highly signiﬁcant regression between plant species number (after logarithmic transformation) and human population in 13 European cities. This was conﬁrmed by Pyˇsek (1993) with a larger data set (77 cities and villages in Central Europe). The latter study showed that this increase is steeper in smaller settlements and levels off at ~1500 species in the largest cites. As population and area are correlated, this is partly an area effect, but also a result of the high habitat diversity in larger cities (Pyˇsek, 1993). Typically, there are just over 500 spontaneously growing species of ferns and ﬂowering plants in small towns in Central Europe, but 1300 or more in big cities (Table 16.4). On the basis of their ﬂoristic composition, cities can be divided into concentric zones: 1, densely built-up central areas; 2, partly built-up central areas; 3, inner suburbs; and 4, outer suburbs (Kunick, 1974, 1981). These zones are, of course, not strictly concentric; their distribution and share of the total area varies with peculiarities in the history of individual cities. The ﬂoristic composition of these zones reﬂects the abiotic factors inﬂuencing plant growth. The sealing Table 16.4 Approximate numbers of plant species (ferns and ﬂowering plants) growing naturally in Central European towns 1 Town size

Species number

Small and medium towns

530–560

Cities with 100 000–200 000 inhabitants

650–730

Cities with 250 000–400 000 inhabitants

900–1000

Cities with more than 1 million inhabitants >1300 1

Sukopp and Trepl (1993).

DISTURBANCE IN URBAN ECOSYSTEMS

of surfaces with asphalt, concrete, or buildings, for example, varies clinally from 70%–100% in zone 1 to 0%–15% in zone 4. Also, the overheating of cities (see above) is more pronounced in zones 1 and 2 than in 3 and 4. Total vegetation cover, and the number of species recorded as rare in Central Europe, decrease from the urban fringe to the city center. The total number of spontaneously growing plant species, the number of non-native species, and the number of annuals are all higher in the more central areas (Table 16.5). The latter trend becomes more apparent by comparing urban ﬂoras to those of adjacent non-urban areas. There are twice as many non-native plant species in the whole city of Berlin (41%) as in surrounding districts (20–25%: Klemm, 1975). Falinski (1971) compared ﬂoras of human settlements in Poland and found an increasing percentage of non-native plant species from small forest settlements (20–30% non-natives) to big cities (50–70% non-natives). A similar reaction of the ﬂora to land-use type in an urban area was found in Japan. In a city of the Chiba prefecture within the metropolitan area of Tokyo, Numata (1977) reported different percentages of nonnative naturalized plant species: 49% in residential areas, 32% in upland ﬁelds, 14% in rice ﬁelds, 13% on a riverside, and 4% on a forest ﬂoor. Non-native naturalized species can be assigned to two groups according to the time of introduction to a region: those introduced in prehistoric times, termed archeophytes (from Greek archaios = old), and those introduced in historic times, the neophytes (Greek neos = new). In central Europe, archeophytes are commonly considered to be species introduced before 1500 AD (Schroeder, 1969). These categories are ecologically signiﬁcant because archeophytes are adapted to habitat types created early in history, including pastures, ﬁelds, and their edges; many of the neophytes, on the other hand, occur predominantly in ruderal, industrial, and urban habitats. This was shown in the above example (Kowarik, 1990): the proportion of neophytes is much higher in the ﬂora of Berlin than in adjacent non-urban areas, but no difference was found for archeophytes. The high proportion of non-native species in cities is partly due to the fact that cities are centers of spread, because new species arrive at railway stations or ports and may be cultivated for the ﬁrst time in (botanical) gardens. On the other hand, anthropogenic changes of growing conditions facilitate their spread; many nonnatives originate from warmer regions and depend on

405 Table 16.5 Characteristics of ﬂoristic city zones in Berlin 1 Zone 2 Total vegetation cover (%) Vascular plant species 3

1

2

3

4

32

55

75

95

380

424

415

357

Species rare in Germany 3

17

23

35

58

Non-native plant species (%)

49.8

46.9

43.4

28.5

Annuals (%)

33.6

30.0

33.4

18.9

1

Sukopp et al. (1979b). 1, densely built-up inner city; 2, partly built-up inner city; 3, inner suburbs; 4, outer suburbs. 3 Number of species in a square kilometer. 2

the higher temperatures in cities. Ailanthus altissima (Kowarik and B¨ocker, 1984) and Chenopodium botrys (Sukopp, 1971), for example, are both well established in European cities and originate in warmer areas. Similar urban conditions, as well as the effects of transport of organisms (the latter at least in the cases of plants and birds), lead to relatively uniform species composition in the centers of various cities in the European lowlands. Of 321 non-native ferns and ﬂowering plants found in Braunschweig (Niedersachsen), more than 80% were also found in Berlin, Vienna, and London (Brandes, 1987). Despite considerable differences in climate, one can speak of a common stock of non-native naturalized species in western and central European cities. The main factor limiting vegetation in streets is the sealing of the surface with asphalt or paving stones. In other areas the vegetation is disturbed by human trampling or by vehicles, leading to soil compaction. Nevertheless, the ﬂora of roadsides can be quite rich; in Berlin, in a study of 61 streets, a total of 375 ﬂowering plant species were found growing naturally, a fourth of the total ﬂora of the city (Langer, 1994). Most of these showed low frequency in the sampling areas. One group of 13 species was present in all areas. Most of these were annuals with some resistance to trampling. Most of the widespread plant communities belong to the phytosociological class Plantaginetea, a group which contains plant communities of heavily trodden swards. Succession in urban vegetation Several authors have described the general vegetation succession on anthropogenic sites from pioneer stages to woodland. Eliaˇs (1996) describes ﬁve stages for Slovakia (Table 16.6). Gilbert (1991) has given detailed

406

Herbert SUKOPP and Uwe STARFINGER

Table 16.7 Vegetation succession on four types of urban waste-land in Berlin from early (top) to late (bottom) stages 1 Railway ballast

Sand

Brick rubble

Humus rich piled-up topsoil

Conyza canadensis stage

Bromus–Corispermum community

Chenopodium botrys community

Chenopodium strictum

Arrhenatherum-pioneer stage

Berteroa incana community

Oenothera stage

Lactuca–Sisymbrium altissimum community

Dry grassland (Arrhenatherum or Festuca)

Festuca trachyphylla dry grassland

Poa–Tussilago community

Artemisia vulgaris community

Mixed deciduous woodland (e.g. Betula)

Robinia woodland

Chelidonium–Robinia community

Sambucus nigra community

Oak woodland

?

Maple woodland

Maple woodland

1

Modiﬁed from Sukopp (1990). Table 16.8 Species richness of spontaneous ferns and ﬂowering plants in urban waste land in Berlin 1

Table 16.6 Successional stages on anthropogenic habitats 1 Stage

Vegetation type

Dominant species Duration (yr)

1st

pioneer

summer and overwintering annuals

1–2

biennials

1–2

Derelict housing estate

5

Site type

ha 2

#2

Neo. 2 Ann. 2

Construction site

0.6

172

26

Brick rubble site

0.5

158

20

19

15

325

36

23

23

2nd

biennial

3rd

forbs

perennials

Derelict railway land

17

332

36

20

4th

grassland

perennial grasses 5–10

Derelict railway goods station

63

417

40

26

5th

woodland

shrubs, trees

Derelict railway goods station

73

395

34

20

1

>10

Modiﬁed from Eliaˇs (1996).

1

Kowarik (1986). Abbreviations: ha, area in ha; #, total number of species; Neo., neophytes (% of total); Ann., annuals (% of total)

2

accounts of four similar stages, which he called the Oxford ragwort stage, the tall-herb stage, the grassland stage, and scrub woodland, and stressed the importance of the substratum and the role of chance in determining the succession. Ash (1991) estimated the time needed for woodland to develop on different urban soil types to be between 30 and 40 years for pulverized fuel ash on the one hand, and 100 years for acidic materials such as colliery shale on the other. Examples for urban sites with a relatively long history (several decades) of more or less undisturbed succession have been described in England (Gilbert, 1991) and Germany (e.g., Sukopp, 1990). A successional scheme for waste-lands in Berlin is shown in Table 16.7. An important feature of these inner-city waste-land sites is their richness in neophytes. A major constituent of woodland stages is the North American black locust (Robinia pseudoacacia). It differs from native tree species such as sand birch (Betula pendula) which it may replace, not only in the

fact that it has no specialized herbivores, but also in its symbiosis with nitrogen-ﬁxing Rhizobium bacteria, and in its clonal growth. Due to the accumulation of nitrogen, succession of sites with Robinia is markedly different from sites that contain native species only (Kowarik, 1992). Even if succession can take place on urban wastelands, small-scale disturbances commonly exist in the form of unorganized land uses, such as trampling, “unregulated” garbage disposal, cutting of vegetation for pet food, etc. These inﬂuences lead to a small mosaic of successional stages with the possibility of high species numbers, especially in large railway areas (Table 16.8). Vegetation succession on landﬁlls usually starts directly after the deposition ceases. In the ﬁrst year, plant stands develop from diaspores that were present in the garbage (e.g., Cucumis spp., Cucurbita spp., Solanum spp., etc.). Further succession leads to stands

DISTURBANCE IN URBAN ECOSYSTEMS

407

Table 16.9 Species richness of different species groups in the succession on landﬁlls in Berlin 1 Age (yr)

1

1

1

3

3

4

4

10

10

10

10

10

20

20

20

20

Plot No.

1

2

3

1

2

1

2

4

5

6

7

8

9

10

11

12

Vegetation cover (%)

60

75

40

40

75

80

95

100

95

50

100

95

95

100

100

100

No. of species

33

50

23

19

32

27

25

14

9

12

8

6

21

15

18

26

“Feral” crops and ornamentals

12

14

6

3

5

2

1

Annual and biennial ruderals

4

5

1

4

8

7

5

Field and garden weeds

9

12

7

4

2

6

Grassland species

1

4

3

1

8

1

7

5

1

2

1

1

1

Perennial weeds of settlements

3

6

3

6

6

6

8

7

7

3

3

4

4

1

2

5

4

Nitrophilous forest-edge species

1

4

1

Forest species

1

2

1

Mosses 1

1

1

2

1

1

5

3

3

2

2

3

3

3

2

8

4

9

13

1

1

2

2

1

1

2

2

Kunick and Sukopp (1975).

of short-lived ruderal plants, and eventually species of shrubs and forests invade (Table 16.9). These successions can be used to indicate speciﬁc disturbances (Pyˇsek and Hajek, 1996). The presence of ionized substances is indicated by salt-tolerating plants (e.g., Chenopodium ﬁcifolium, C. glaucum, C. rubrum, and Puccinellia distans). Oil derivatives are detected by the decrease of “petroleophobe” species, (Arrhenatherum elatius, Artemisia vulgaris) and the increase of “petroleotolerants” (e.g., Cirsium arvense, Urtica dioica). Heavy-metal contamination can be indicated by damaged tissue, and the presence of speciﬁc types of necroses and chloroses. Animal communities Animals are affected by urban activities in a multitude of ways. Noise disturbs many birds more in the form of a singular noise event than as a continuous phenomenon. The ﬁreworks of New Years’ Eve celebrated in German cities were shown to alter considerably the ﬂight movements of crows (Corvus sp.) (J¨adicke and Storck, 1979). The activity of free-ranging cats in a Japanese city was lowest during periods when most people with their dogs were present in the streets (Obara, 1995). Gepp (1977) discussed circumstances causing losses of free-living animals in towns. Street, rail, and air trafﬁc result in the killing of individual animals, especially mammals, birds, amphibians, and insects. Even the structure of urban buildings kills animals

(e.g., when birds ﬂy against windows, or when insects become trapped in rooms). Some species or groups of species are thus eliminated from central areas in cities [e.g., toads (Bufo bufo) and hedgehogs (Erinaceus europaeus)]. On a transect from suburb to center in Graz, Austria, the number of inviduals of ﬂying insects did not vary much, but diversity and total biomass decreased sharply (Gepp, 1977). Restoration of buildings and open space in a densely built-up residential area in Berlin led to a sharp decline in abundance of bird species (H.-G. Braun, pers. comm.). In the metropolitan area of Tokyo, the increase of disturbance during the urbanization process was closely related to the history of species extinctions. Several mammals and large insects disappeared from central parts much earlier than from suburban regions (Obara, 1995). Landﬁlls can be rich in animals, especially species of open ground, such as thermophilous spiders, beetles, and locusts. Birds were studied by Steiof (1987) on six landﬁlls in Berlin with a total area of 165 ha. Of the 44 species found breeding on the landﬁlls, 12 were on the Berlin red data list of endangered species. In the course of succession toward closed stands, the rare species decreased, and ubiquitous animal species became more numerous. Animals of cities have adapted to humans to varying degrees. Some species occur regularly in cities so they can be called urban species [feral pigeon (Columba livia f. domestica), jackdaw (Corvus monedula)]. Some

408

habitat types and their animals disappear as cities develop (e.g., those of nutrient-poor soils). In other situations urban habitat features can take the place of certain natural features; bird species from rocks and cliffs can breed on ledges and walls of buildings, species from caverns can live in cellars or other rooms (Gilbert, 1991). Total numbers of animal species in urban habitats show similar trends to those of plants. At least in many groups of invertebrates, birds, and mammals (except large carnivores) species numbers in cities are higher than in adjacent non-urban areas. The highest species numbers are found in areas of the outer and inner suburbs; in city centers and new housing developments they are considerably lower (Gilbert, 1991; Klausnitzer, 1993). Many of the non-native species of animals and fungi which prefer urban habitats originate from warmer regions (Pisarski and Trojan, 1976; Erkkil¨a and Niemel¨a, 1986).

EVALUATION

The composition of urban ecosystems as opposed to those of rural or near-natural landscapes is the result of a multitude of disturbances which collectively form the human impact. There are different general approaches to the question of how to compare the intensities of disturbance and the changes urban disturbances cause in ecosystems. Chronic disturbances applied to natural communities can produce changes that are in some respect the reverse of succession, changes termed “retrogression” by Whittaker and Woodwell (1973). Retrogression may be quantiﬁed using indicators such as weighted averages of “decreasers, increasers, and invaders”, reduction of species diversity from that of undisturbed situations, or community coefﬁcients for disturbed and undisturbed samples. In Europe there is a long tradition of attempts to classify the human impact on the vegetation using different degrees of naturalness. In a review of the literature, Kowarik (1990, 1991) pointed out that there are two principal ways to do this: the concepts of Westhoff (1949), Ellenberg (1963) and others express naturalness as a distance from a pristine, undisturbed ecological condition which existed before human impact began. Jalas (1955), T¨uxen (1956), and Sukopp (1972), on the other hand, tried to evaluate the human

Herbert SUKOPP and Uwe STARFINGER Table 16.10 The hemeroby system (examples ordered with increasing degree of hemeroby) 1 Human inﬂuence

Ecosystems

Not present

high mountains

Emissions, minor biomass removal

little-disturbed forests, growing bogs

Tree-cutting, mowing of grass

more intensively managed forests, dry grassland

Plowing, draining, fertilizing

forest plantations of alien species, intensively managed grassland

Deep plowing, intensive fertilization, biocides

ﬁelds, gardens, vineyards

Total destruction of vegetation and soil

new landﬁlls, partially paved urban areas

Poisoned or completely sealed surfaces

no vegetation of vascular plants

1

Sukopp (1978); Kowarik (1990).

impact on the vegetation by relating site conditions and vegetation to a potential (or future) undisturbed situation. The difference between the two reference points which Kowarik terms “nature I” and “nature II” becomes apparent in examples where strongly altered sites are left relatively undisturbed for a period of time as in the example of urban waste land described above (p. 406, Table 16.7). Jalas’ concept of hemeroby (Greek: hemeros = tame; bios = life), reﬁned by Sukopp (1972) and Kowarik (1988), is widely applied in Central Europe. It classiﬁes the overall effects of human impact on vegetation in 7 grades (Table 16.10). Kowarik (1988) produced hemeroby spectra for the vascular plant species of Berlin by evaluating their presence in plant communities at different levels of hemeroby. He found the highest overall species richness in situations of intermediate hemeroby, conﬁrming Connell’s “intermediate disturbance hypothesis”. Nonnative species, however, react more positively to human impact: the highest number of them was found in more strongly inﬂuenced vegetation of a higher hemeroby level (Fig. 16.3). The strong preference of certain species for certain levels of hemeroby allows their use as indicators of particular disturbance intensities (Kowarik, 1991). CONCLUSIONS

Climate, soil, air, and water in cities suffer from human

DISTURBANCE IN URBAN ECOSYSTEMS

409

Fig. 16.3. Species richness in relation to human impact (increasing from hemeroby level 1 to 9) in the ﬂora of the city of Berlin (from Kowarik, 1991, modiﬁed. Levels are not the same as in Table 16.10). Fat solid line, all species; solid line, natives; dotted line, non-natives; dashed line, neophytes; dash-dotted line, archeophytes.

impacts that are usually much stronger than in rural areas. In spite of this multitude of disturbances, cities are quite rich in habitats as well as in plant and animal species. Typically, disturbance of urban ecosystems leads to a decrease in the number of species native to the region and an increase of introduced non-native species. This process is at least partly reversed in the course of succession. Disturbance and the recovery from it during succession are key factors determining the habitat mosaic in urban ecosystems. As an ever-increasing percentage of the world’s population lives in cities, urban biota are under increasing pressure. At the same time, the conservation of urban biota is of great importance, because cities are the primary location where people have contact with nature (Starﬁnger and Sukopp, 1994). In many parts of the world, knowledge about urban ecosystems and their disturbance regimes is still insufﬁcient.

ACKNOWLEDGMENTS

Lawrence Walker and two anonymous reviewers have commented on and greatly improved the manuscript. Bettina Matzdorf helped with the ﬁgures.

REFERENCES Aey, W., 1990. Historical approaches to urban ecology. Methods and ﬁrst results from a case study (L¨ubeck, West-Germany). In: H. Sukopp, I. Kowarik and S. Hejny (Editors), Urban Ecology. SPB, Academic Publishing, Den Haag, pp. 113–129. Ash, H.J., 1991. Soils and vegetation in urban areas. In: P. Bullock and P.J. Gregory (Editors), Soils in the Urban Environment. Blackwell, Oxford, pp. 153–170. Auhagen, A., 1991. Rote Listen der gef¨ahrdeten Pﬂanzen und Tiere in Berlin 1991 – Zusammenfassung und Empfehlungen. Landschaftsentwicklung und Umweltforschung, S 6: 5–11. Auhagen, A. and Sukopp, H., 1980. Das Streusalz. Umweltschutzforum Berlin, 47/48: 23–42.

410 Balke, K.D., 1974. Der thermische Einﬂuß besiedelter Gebiete auf das Grundwasser, dargestellt am Beispiel der Stadt K¨oln. gwf-Wasser/ Grundwasser, 115: 117–124. Blume, H.P., 1993. B¨oden. In: H. Sukopp and R. Wittig (Editors), Stadt¨okologie. G. Fischer, Stuttgart, pp. 154–171. Blume, H.P., Chinnow, D., Hartnick-K¨ummel, C., Hellriegel, T. and Stoll, E., 1977. Bau- und nutzungsbedingte Ver¨anderungen an ¨ Straßenrand-Okosystemen. Z. TU Berlin, 9: 278–322. Blume, H.P., Bornkamm, R. and Sukopp, H., 1979. Vegetationssch¨aden und Bodenver¨anderungen in der Umgebung einer M¨ulldeponie. Z. Kulturtech. Flurbereinigung, 20: 65–79. B¨ocker, R., 1985. Bodenversiegelung – Verlust vegetationsbedeckter Fl¨achen in Ballungsr¨aumen (Surface sealing – loss of vegetation in metropolitan areas. Example Berlin (West) German w. English summary.) Landschaft Stadt, 17: 57–61. Bradshaw, A.D. and Chadwick, M.J., 1980. The Restoration of Land. Blackwell, Oxford, 317 pp. Brande, A., 1986. Mittelalterliche Siedlungsvorg¨ange in Berliner Pollendiagrammen. Cour. Forschungsinst. Senckenberg, 86: 409–414. Brandes, D., 1987. Verzeichnis der im Stadtgebiet von Braunschweig wildwachsenden und verwilderten Gef¨aßpﬂanzen. Universit¨atsbibliothek der Technischen Universit¨at Braunschweig, 44 pp. den Hartog, C., Kvˆet, J. and Sukopp, H., 1989. Reed. A common species in decline. Aquat. Bot., 35: 1–4. Drayton, B. and Primack, R.B., 1996. Plant species lost in an isolated conservation area in Metropolitan Boston from 1894 to 1993. Conservation Biol., 10: 30–39. Duvigneaud, P. and Denayer-de Smet, S., 1977. L’ecosyst`eme urbain bruxellois. In: P. Duvigneaud and P. Kestemont (Editors), Productivit´e Biologique en Belgique. SCOPE. Travaux de la Section belge du Programme Biologique International Gembloux, pp. 581–599. Eliaˇs, P., 1996. Vegetation dynamics of anthropogenic habitats in ¨ settlements. Verh. Ges. Okol., 25: 219–224. Ellenberg, H., 1963. Vegetation Mitteleuropas mit den Alpen. Ulmer, Stuttgart. Erkkil¨a, R. and Niemel¨a, T., 1986. Polypores in the parks and forests of the city of Helsinki. Karstenia, 26: 1–40. Falinski, J.B., 1971. Synanthropisation of plant cover. II. Synanthropic ﬂora and vegetation of towns connected with their natural conditions, history and function. Mater. Zakłado Fitosocjologie Stosowanej U.W., 27: 1–317. Gepp, J., 1977. Technogene und strukturbedingte Dezimierungsfak¨ toren der Stadttierwelt – ein Uberblick. In: J. Gepp (Editor), Stadt¨okologie. Verlag f¨ur die Technische Universit¨at, Graz, pp. 99– 127. Gilbert, O.L., 1991. The Ecology of Urban Habitats. Chapman and Hall, London, 369 pp. Grenzius, R., 1984. Starke Versauerung der Waldb¨oden Berlins. Forstwissenschaftliches Centralblatt, 103: 131–139. Grime, J.P., 1979. Plant Strategies and Vegetation Processes. John Wiley, New York, 222 pp. Harris, J.A., 1991. The biology of soils in urban areas. In: P. Bullock and P.J. Gregory (Editors), Soils in the Urban Environment. Blackwell, Oxford, pp. 139–152. Horbert, M., Kirchgeorg, A. and von St¨ulpnagel, A., 1983. Ergebnisse stadtklimatischer Untersuchungen als Beitrag zur Freiraumplanung. Umweltbundesamt, Berlin, 187 pp.

Herbert SUKOPP and Uwe STARFINGER H¨otzl, H. and Makurat, A., 1981. Ver¨anderungen der Grundwassertemperaturen unter dicht bebauten Fl¨achen am Beispiel der Stadt Karlsruhe. Z. Dtsch. Geol. Ges., 132: 767–777. Jackowiak, B., 1982. Occurrence of Puccinellia distans in the city of Poznan. Badania ﬁziograﬁczne nad Polska Zachodnia, 33: 129–142. Jackowiak, B., 1996. Chorological-ecological model of the spread of Puccinellia distans (Poaceae) in central Europe. Fragm. Florist. Geobot., 41: 551–561. J¨adicke, B. and Storck, H.J., 1979. Großr¨aumige Umlenkung der Schlafplatzﬂ¨uge von Kr¨ahen durch Silvesterl¨arm. J. Ornithol., 120: 326–327. J¨ager, E., 1988. M¨oglichkeiten der Prognose synanthroper Pﬂanzenausbreitungen. Flora, 180: 101–131. Jalas, J., 1955. Hemerobe und hemerochore Pﬂanzenarten: ein terminologischer Reformversuch. Acta Soc. Fauna Flora Fennica, 72: 1–15. ¨ Klausnitzer, B., 1993. Okologie der Großstadtfauna. G. Fischer, Stuttgart, 454 pp. Klemm, G., 1975. Pﬂanzengeographischer Vergleich neuerer brandenburgischer Lokalﬂoren. Gleditschia, 3: 35–52 (cited by Kowarik, 1990). Klotz, S., 1990. Species/area and species/inhabitants relations in European cities. In: H. Sukopp, I. Kowarik and S. Hejny (Editors), Urban Ecology. SPB, Academic Publishing, Den Haag, pp. 99– 103. Kowarik, I., 1986. Vegetationsentwicklung auf innerst¨adtischen Brachﬂ¨achen – Beispiele aus Berlin (West). Tuexenia, 6: 75–98. Kowarik, I., 1988. Zum menschlichen Einﬂuß auf Flora und Vegetation. Landschaftsentwicklung und Umweltforschung, 56: 1–280. Kowarik, I., 1990. Some responses of ﬂora and vegetation to urbanization in Central Europe. In: H. Sukopp, I. Kowarik and S. Hejny (Editors), Urban Ecology. SPB, Academic Publishing, Den Haag, pp. 45–74. Kowarik, I., 1991. The adaptation of urban ﬂora to man-made perturbations. In: O. Ravera (Editor), Terrestrial and Aquatic Ecosystems: Perturbation and Recovery. Ellis Horwood, London, pp. 176–184. Kowarik, I., 1992. Einf¨uhrung und Ausbreitung nichteinheimischer Geh¨olzarten in Berlin und Brandenburg und ihre Folgen f¨ur Flora und Vegetation. Verh. Bot. Ver. Berlin Brandenburg Beiheft 3: 1–188. Kowarik, I. and B¨ocker, R., 1984. Zur Verbeitung, Vergesellschaftung und Einb¨urgerung des G¨otterbaumes (Ailanthus altissima (Mill.) Swingle) in Mitteleuropa. Tuexenia, 4: 9–29. Kratzer, A., 1937. Das Stadtklima. Vieweg, Braunschweig, 184 pp. Kunick, W., 1974. Ver¨anderungen von Flora und Vegetation einer Großstadt, dargestellt am Beispiel von Berlin (West). Ph.D. Dissertation, Technische Universit¨at Berlin, 472 pp. Kunick, W., 1981. Comparison of the ﬂora of some cities of the Central European Lowlands. In: R. Bornkamm, J.A. Lee and M.R.D. Seaward (Editors), Urban Ecology, 2nd European Ecological Symposium, Berlin, 1980. Blackwell, Oxford, pp. 13– 22. Kunick, W. and Sukopp, H., 1975. Vegetationsentwicklung auf M¨ulldeponien Berlins. Berl. Naturschutzbl¨atter, 56: 141–145. Kuttler, W., 1993. Stadtklima. In: H. Sukopp and R. Wittig (Editors), Stadt¨okologie. G. Fischer, Stuttgart, pp. 113–153.

DISTURBANCE IN URBAN ECOSYSTEMS Landsberg, E., 1981. The Urban Climate, International Geophysics Series, Vol. 28. Academic Press, New York, 275 pp. Langer, A., 1994. Flora und Vegetation st¨adtischer Straßen am Beispiel Berlins. Landschaftsentwicklung und Umweltforschung, S 10: 1–199. Leuchs, W. and R¨omermann, H., 1991. Auswirkungen stadt¨okologischer Gestaltungsmaßnahmen auf die Grundwassersituation. In: H. Schuhmacher and B. Thiesmeyer (Editors), Urbane Gew¨asser. Westarp-Wissenschaften, Essen, pp. 427–445. Londo, G., 1976. Over de Nederlandse lijst van hydro-, freato- en afreatofyten. Gorteria, 8: 25–29. McDonnell, M.J. and Pickett, S.T.A., 1990. Ecosystem structure and function along urban-rural gradients: an unexploited opportunity for ecology. Ecology, 71: 1232–1237. Numata, M., 1977. The impact of urbanization on vegetation in Japan. In: A. Miyawaki and R. T¨uxen (Editors), Vegetation Science and Environmental Protection. Maruzen, Tokyo, pp. 161–171. Obara, H., 1995. Animals and man in the process of urbanization. In: H. Sukopp, M. Numata and A. Huber (Editors), Urban Ecology as the Basis of Urban Planning. SPB, Academic Publishing, Den Haag, pp. 191–201. Oke, T.R., 1987. Boundary Layer Climates. Methuen, London, 435 pp. Pisarski, B. and Trojan, P., 1976. Zoocenozy obszarow zurbanizowanych (Animal associations in urbanized areas). Wiadomosci Ecologiczne, 22: 338–344. Pyˇsek, A. and Hajek, M., 1996. Die Ruderalvegetation der Ablagerungspl¨atze und ihre praktische Ausnutzung zur ¨ Kontaminationsentdeckung. Verh. Ges. Okol., 25: 215–217. Pyˇsek, P., 1993. Factors affecting the diversity of ﬂora and vegetation in central European settlements. Vegetatio, 106: 89–100. Rebele, F., 1994. Urban ecology and special features of urban ecosystems. Global Ecol. Biogeogr. Lett., 4: 173–187. Rowntree, R.A. (Editor), 1984. Ecology of the urban forest. Part I: structure and composition. Urban Ecol., 8: 1–178. Rowntree, R.A. (Editor), 1986. Ecology of the urban forest. Part II: function. Urban Ecol., 9: 227–437. Ruge, U., 1978. Physiologische Sch¨aden durch Umweltfaktoren. In: F.H. Meyer (Editor), B¨aume in der Stadt. Ulmer, Stuttgart, 134– 198. Schlenther, L., Marschner, B., Hoffmann, C. and Renger, M., 1996. Ursachen mangelnder Anwuchserfolge bei der Aufforstung der Rieselfelder in Berlin-Buch – bodenkundliche Aspekte. Verh. Ges. ¨ Okol., 25: 349–359. Schroeder, F.-G., 1969. Zur Klassiﬁzierung der Anthropochoren. Vegetatio, 16: 225–238. Schuhmacher, H., 1991. Limnologische Vorgaben und Bewertungskriterien zur o¨ kologischen Verbesserung urbaner Fließgew¨asser. In: H. Schuhmacher and B. Thiesmeyer (Editors), Urbane Gew¨asser. Westarp-Wissenschaften, Essen, pp. 16–36. Sharpe, D.M., Stearns, F., Leitner, L.A. and Dorney, J.R., 1986. Fate of natural vegetation during urban development of rural landscapes in southeastern Wisconsin. Urban Ecol., 9: 267–287. Sousa, W.P., 1984. The role of disturbance in natural communities. Ann. Rev. Ecol. Syst., 15: 353–391. Starﬁnger, U. and Sukopp, H., 1994. Assessment of urban biotopes for nature conservation. In: E.A. Cook and H.N. van Lier (Editors), Landscape Planning and Ecological Networks. Elsevier, Amsterdam, pp. 89–115.

411 Steiof, K., 1987. Brutv¨ogel und Deponien-Rekultivierung – ein Beitrag zur Landschaftsbewertung und -planung am Beispiel Berlin. Landschaftsentwicklung und Umweltforschung, 47: 1–107. ¨ Sukopp, H., 1971. Beitr¨age zur Okologie von Chenopodium botrys L. 1. Verbreitung und Vergesellschaftung. Verh. Bot. Ver. Provinz Brandenburg, 108: 3–25. Sukopp, H., 1972. Wandel von Flora und Vegetation in Mitteleuropa unter dem Einﬂuß des Menschen. Ber. Landwirtsch., 50: 112–130. Sukopp, H., 1978. An approach to ecosystem degradation: opening remarks by session chairman. In: M.W. Holdgate and M.J. Woodman (Editors), The Breakdown and Restoration of Ecosystems. Plenum, New York, pp. 123–127. Sukopp, H., 1981. Grundwasserabsenkungen – Ursachen und Auswirkungen auf Natur und Landschaft Berlins. Wasser – Berlin, Bd. 1. Die technisch-wissenschaftlichen Vortr¨age auf dem Kongress Wasser 1981, Berlin, pp. 239–272. Sukopp, H. (Editor), 1990. Stadt¨okologie. Das Beispiel Berlin. Dietrich Reimer, Berlin, 455 pp. Sukopp, H. and Trepl, L., 1993. Stadt¨okologie und Naturschutz in der Großstadt. Biol. Schule, 42: 187–197. Sukopp, H. and Wittig, R. (Editors), 1993. Stadt¨okologie. G. Fischer, Stuttgart, 402 pp. Sukopp, H., Anders, K., Bierbach, H., Brande, A., Blume, H.P., Elvers, H., Horbert, M., Horn, R., Kirchgeorg, A., L¨uhrte, A. v., Riecke, F., Stratil, H., Trepl, L. and Weigmann, G., 1979a. ¨ Okologisches Gutachten u¨ ber die Auswirkungen von Bau und Betrieb der BAB Berlin (West) auf den Großen Tiergarten. Senator f¨ur Bau- und Wonhungswesen, Berlin, 105 pp. Sukopp, H., Blume, H.P. and Kunick, W., 1979b. The soil, ﬂora, and vegetation of Berlin’s waste lands. In: I.C. Laurie (Editor), Nature in Cities. John Wiley, Chichester, pp. 115–132. Sukopp, H., Hejny, S. and Kowarik, I. (Editors), 1990. Urban Ecology. SPB, Academic Publishing, Den Haag, 282 pp. Sukopp, H., Numata, M. and Huber, A. (Editors), 1995. Urban Ecology as the Basis of Urban Planning. SPB, Academic Publishing, Den Haag, 218 pp. Thornton, I., 1991. Metal contamination of soils in urban areas. In: P. Bullock and P.J. Gregory (Editors), Soils in the Urban Environment. Blackwell, Oxford, pp. 47–75. ¨ Trepl, L., 1992. Stadt-Natur – Okologie, Hermeneutik und Politik. In: Bayerische Akademie der Wissenschaften (Editor), Stadt¨okologie, ¨ Rundgespr¨ache der Kommission f¨ur Okologie, Band 3. Pfeil, M¨unchen, pp. 53–58. T¨uxen, R., 1956. Die heutige potentielle nat¨urliche Vegetation als Gegenstand der Vegetationskartierung. Angew. Pﬂanzensoziol., 13: 5–42. Ullmann, I. and Heindl, B., 1989. Geographical and ecological differentiation of roadside vegetation in temperate Europa. Bot. Acta, 4: 261–269. von St¨ulpnagel, A., Horbert, M. and Sukopp, H., 1990. The importance of vegetation for the urban climate. In: H. Sukopp, S. Hejny and I. Kowarik (Editors), Urban Ecology. SPB, Academic Publishing, Den Haag, pp. 175–194. Westhoff, V., 1949. Schaakspiel met de natuur. Natuur Landschap, 3. White, C.S. and McDonnell, M.J., 1988. Nitrogen cycling processes and soil characteristics in an urban versus rural forest. Biogeochemistry, 5: 243–262.

412 White, P.S. and Pickett, S.T.A., 1985. Natural disturbance and patch dynamics: an introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, New York, pp. 3–13. Whittaker, R.H. and Woodwell, G.M., 1973. Retrogression and Coenocline Distance. Handbook of Vegetation Science 5. The Hague, pp. 53–73.

Herbert SUKOPP and Uwe STARFINGER Wittig, R. and Sukopp, H., 1993. Was ist Stadt¨okologie? In: H. Sukopp and R. Wittig (Editors), Stadt¨okologie. G. Fischer, Stuttgart, pp. 1–9. Wu, L. and Antonovics, J., 1976. Experimental ecological genetics in Plantago. II. Lead tolerance in Plantago lanceolata and Cynodon dactylon from a roadside. Ecology, 57: 205–208.

Chapter 17

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

INTRODUCTION

The expansion of the human population and its associated agriculture and industry over the last several centuries has brought with it major alterations to landscapes, disturbance regimes, and dispersal of species. While it has become dogma that disturbance of natural systems promotes invasion by non-indigenous species, the causes of invasion and the links between disturbance and biological invasions have yet to be adequately understood. Perturbations to natural or semi-natural systems have in the past been broadly lumped under the heading of disturbance, and several review papers have discussed the relationship between disturbance and biological invasion (Hobbs, 1989; Rejm´anek, 1989; Hobbs and Huenneke, 1992). While each of these reviews provides valuable information regarding the actual or potential role of habitat disruption in promoting invasion, we believe that many important questions about the relationship between disturbance and invasion remain unanswered. For example, although it is clear that some catastrophic disturbances (e.g., logging, clearing for agriculture) are followed by rapid expansion of populations of nonindigenous species, it is not known if invaders become a self-sustaining and long-term part of the “recovering” ecosystem or whether they are successional to native species. Also, the site-speciﬁc nature of “natural” disturbances in promoting invasion is unknown, as is the relative importance of disturbance versus propagule availability in promoting species change within communities. All ecosystems experience disturbance on some spatial and temporal scale, yet disturbance is difﬁcult to quantify. Indeed, few studies measure the impact of the disturbing agent in terms of area, spatial

structure, intensity, and temporal distribution of the impacts. We believe that some of the reasons for the lack of clarity surrounding the linkage between disturbance and invasions include non-rigorous use of terminology surrounding the word “disturbance”, difﬁculty in quantifying or characterizing disturbance regimes in most sites, and a failure to distinguish between anthropogenic and “natural” disturbances and between irregular catastrophic disturbances and recurrent smaller-scale disturbances. Finally, current disturbance patterns may not reﬂect conditions that promoted the initial invasions, and often information on pattern and process during the invasion process is lacking. There are numerous anecdotal studies where causative mechanisms are invoked but are largely obscure, and where multiple causation is likely. Disturbance is often not independent of associated impacts such as proximity to human habitation or density of nearby populations of invaders, yet these colonization-related processes are generally ignored. Also, many invading species alter the disturbance regime as they become well established, and by so doing can favor further invasion of the site (e.g., D’Antonio and Vitousek, 1992). If this happens rapidly, then the role of disturbance in promoting the initial invasion may be difﬁcult to separate from the effect of the invader on the disturbance regime. In this chapter we review accepted deﬁnitions of disturbance, selecting a somewhat conservative one for the purposes of providing as concise and unambiguous a review as possible. We focus on major disturbance types for which we found the best information, and review studies assessing the conditions under which these disturbances promote or inhibit invasion. For terrestrial invasions, we categorize disturbances by

413

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Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

spatial scale and by type (obviously anthropogenic versus more or less natural), conﬁning our review mainly to plant invasions. We also discuss a subset of aquatic plant invasions that occur at the interface between fully terrestrial and fully aquatic systems. These include invasions into riparian corridors, stream margins, and wetlands. We then reverse the issue and examine how invading species themselves alter the disturbance regime. Finally, we assess how altered disturbance regimes caused by introduced species may promote or inhibit further invasions. We use the terms non-native, introduced, and non-indigenous species interchangeably throughout. What is an invasion A daunting problem associated with this review is the lack of a standard scale for deﬁning when an invasion has occurred. Introduced species can be present on a site but may not maintain themselves over several generations or may do so only at a very low density. Are these successful invasions? Most studies use presence/absence data or subjective assessments to conclude whether or not an invasion has occurred. We found numerous examples where introduced species were reported from relatively undisturbed sites, and where we could not judge whether these invasions represented self-replacing populations or were similar in their magnitude or persistence to populations in nearby disturbed habitat. In these cases we considered these invasions to be successful unless otherwise indicated by the authors. Deﬁning the disturbance regime Perhaps the most commonly used deﬁnition of disturbance is that of White and Pickett (1985), who stated that a disturbance is “any relatively discrete event in time that disrupts ecosystem, community, or population structure and changes resources, substrate availability, or the physical environment”. However, we believe that this deﬁnition is too vague and includes a wide range of changes to populations and ecosystems that might best be described as perturbations or modiﬁcations rather than disturbances. For example, their deﬁnition can be interpreted to include an increase in the availability of resources without the direct killing of individuals. This could include atmospheric deposition of nitrogen, which several investigators have termed a “disturbance” to natural systems (e.g., Hobbs and Huenneke,

1992). We do not deny that such modiﬁcations may promote changes in species composition including invasions; however, we do not consider an increase in resources alone to be a disturbance without a discrete killing event or an event that signiﬁcantly reduces standing biomass. Accordingly, we prefer to use the deﬁnition provided by Sousa (1984): “A disturbance is a discrete, punctuated killing, displacement, or damaging of one or more individuals (or colonies) that directly or indirectly creates an opportunity for new individuals (or colonies) to become established” or for individuals adjacent to the disturbance to garner further resources (our addition). To Sousa’s deﬁnition we also add chronic grazing by livestock or wildlife, which may be viewed as a continuous series of damaging events. Our deﬁnition excludes: (1) gradual changes in groundwater due to humanrelated alterations of hydrology; (2) gradual increases in aquatic pollutants that might stress individuals and gradually lead to the decline of populations through decreased reproduction; (3) atmospheric pollutants that may weaken plants and make them more susceptible to other stresses or agents of disturbance; and (4) nitrogen deposition, which might change plant performance and the strength of competitive interactions over several years without discrete mortality. We also do not consider the phenomenon of invasion by itself to be a disturbance, although speciﬁc invaders may cause death of native organisms and ultimately contribute to altered disturbance regimes. Every site can be characterized as having a natural disturbance regime, where killing events occur with given intensities, sizes, and average return intervals (see Sousa, 1984; White and Pickett, 1985). Characteristics of a disturbance regime are determined by physical site characteristics (e.g., topography, soil structure, slope and soil stability, and altitude) and factors extrinsic to the system, such as the local or regional weather (e.g., frequency of lightning, level of fuel moisture). In addition, the community of species present on a site may inﬂuence any aspect of that disturbance regime and the responsiveness of the system to the disturbance. For example, inherent ﬂammability characteristics of individual stems and leaves can inﬂuence ﬁre intensity and ﬁre return interval. Size or stature of trees in a forest can inﬂuence susceptibility to windthrow. Presumably, the disturbance regime of a particular area changes as the climate changes, and has therefore been in continuous ﬂux through geological time. However, we will conﬁne our discussion to

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

disturbances or disturbance regimes that are presumed to be representative of the recent history of an area. Anthropogenic disturbances frequently are very different from anything experienced by the natural community occupying a site. Here we will examine both how anthropogenic disturbances inﬂuence invasion, and whether or not elements of a “natural” disturbance regime promote or even suppress invasion. Suppression of a natural disturbance regime has been an important part of land management in the 20th century. We include cases where the natural disturbance regime has been altered, and examine how these changes inﬂuence invasion.

DISTURBANCE AND INVASION: THE EVIDENCE

For the purposes of this review we will divide disturbances into overt anthropogenic disturbances and disturbances that are natural or near-natural. The former group includes disturbance corridors (roads, trails, pipelines, power-line rights of way), livestock grazing, and clearing for agriculture, timber harvesting or other directed human purposes. These are largely disturbances of a quality or magnitude that do not resemble natural disturbances which might have occurred in that area. Near-natural and natural disturbances include ﬂoods, small-scale soil disturbance associated with animal activity, insect outbreaks, grazing by native animals, short-term drought, ﬁre, hurricanes and other storms, and ﬂoods. Because the origin of ﬁres is often not reported and previous ﬁre frequencies are often unknown, we chose to treat all ﬁres not clearly associated with logging or other overt forms of land clearing as natural or near-natural disturbances. Terrestrial ecosystems: anthropogenic disturbances Disturbance corridors Numerous articles document the occurrence of introduced plant species along roads and disturbance corridors (Table 17.1). Here we also include a few studies in which the nature of the disturbance at a forest edge was unspeciﬁed and could have been catastrophic (agriculture), but where the study focused on the disturbed edge as a source for weed invasion (Brothers and Spingarn, 1992; Brandt and Rickard, 1994). The construction of disturbance corridors creates a new environment for colonization; in addition to elimination

415

of the pre-existing community, construction of a corridor often disrupts the soil proﬁle and structure, eliminates the seed bank, and alters site hydrology and chemistry. The end result is that roadsides and trail edges are novel habitats, and tend to be populated by “weedy” introduced species. However, at least a portion of this occurrence is the result of increased frequency of contact with vehicles or humans, which might inadvertently be spreading propagules along corridors (Chaloupka and Domm, 1986; Lonsdale and Lane, 1994). The importance of propagule arrival through human trafﬁc is rarely compared to or differentiated from the role of disturbance in promoting invasions. It is not surprising to ﬁnd that introduced species dominate road verges and other disturbance corridors (Table 17.1). What is more interesting is the possibility that these corridors act as entry points for nonindigenous species into natural communities. We found a total of 11 studies, involving 14 community types, that provided data on the occurrence of introduced species along roadsides, and their subsequent penetration into the surrounding relatively undisturbed vegetation. In seven cases, introduced species were largely concentrated near disturbed edges and rarely penetrated into surrounding vegetation (Forcella and Harvey, 1983; Hobbs and Atkins, 1991; Brothers and Spingarn, 1992; Wein et al., 1992; Parker et al., 1993; Brandt and Rickard, 1994). In the other seven cases, corridors were acting to promote invasion of adjacent undisturbed habitat [Tyser and Key, 1988; Burgess et al., 1991; MacDonald et al., 1991; Luken and Goessling, 1995; Zink et al., 1995 (3 cases)]. In the former group of studies, no penetration of undisturbed vegetation was observed even where natural disturbances occurred in the surrounding vegetation. For example, Parker et al. (1993) found that two species of Eurasian weeds were abundant along an abandoned roadside in Wisconsin (U.S.A.) but did not penetrate the intact prairie nearby, in spite of the occurrence there of rodent-caused soil disturbances. Likewise, Brothers and Spingarn (1992) found that introduced species common on disturbed edges of intact forests in Indiana (U.S.A.) did not penetrate the forest remnants, in spite of the occurrence of tree-fall gaps in those habitats. By contrast, Luken and Goessling (1995) in Kentucky (U.S.A.) and MacDonald et al. (1991) in the island R´eunion found that shade-tolerant introduced species along disturbed forest edges could penetrate the understory of adjacent undisturbed forests. Zink et al. (1995) in southern California (U.S.A.) found

416

Table 17.1 Studies examining the occurrence of introduced species on disturbance corridors (roads, trails, pipelines, etc.) Location 1

Invader 2

Findings

Study type 3

Reference

California (coastal habitats)

Eur. grasses, forbs

introduced species most abundant on roadsides

Ane.

Knops et al. (1995)

California (coastal sage scrub, oak woodland, native grassland, chaparral)

Eur. annuals

Eur. annuals dominate pipeline coridor, spread into 3/4 adjacent communities

Obs.

Zink et al. (1995)

North America

Salsola iberica

dense on roadsides, does poorly in adjacent desert

Obs.

Brandt and Rickard (1994)

Centaurea maculosa

dense roadside populations support adjacent grassland populations

Obs.

Tyser and Key (1988)

Montana (alpine/subalpine)

introduced weeds

abundant on roadsides, no penetration of adjacent sites

Obs.

Forcella and Harvey (1983)

Colorado (desert)

annual grasses, forbs

increase with vehicle “tracking”

Obs.

Shaw and Diersing (1990)

New Mexico (montane)

Dipsacus sp.

spreading along roads

Obs.

Huenneke and Thomson (1995)

Arizona (Sonoran Desert)

Eur. annuals

abundant on all disturbance corridors, spreads into adjacent desert

Obs.

Burgess et al. (1991)

Wisconsin (native prairie)

Eur. forbs

conﬁned to roads, not spreading into prairie

Obs.

Parker et al. (1993)

Kentucky (temperate deciduous forest)

Lonicera maackii

grows best along disturbed edges, can penetrate forest

Obs., Exp.

Luken and Goessling (1995)

Indiana (temperate deciduous forest)

many species

abundant along edges, do not penetrate undisturbed forest

Obs.

Brothers and Spingarn (1992)

N.W. Territories: Canada (boreal forest)

Eur. ruderals

Bromus inermis slowly spreading from roads, other species not spreading

Obs.

Wein et al. (1992)

W. Australia (dry forest, shrubland)

Eur. annuals

occur on roads, do not penetrate adjacent vegetation

Obs., Exp.

Hobbs and Atkins (1991); Hester and Hobbs (1992)

Australia, Kakadu (monsoonal forest)

many species

transported by buses, some abundant along roads

Obs.

Lonsdale and Lane (1994)

South Africa (shrubland)

many species

very abundant along roads

Ane., Obs.

Brown and Gubb (1986)

La R´eunion (tropical forest)

many species

abundant along roads, spreading into adjacent forest

Ane.

MacDonald et al. (1991)

Other continents

1 2 3

Ecosystem type in parentheses following location. Where species names or life forms were not given, invaders are presented in general terms; Eur., European origin. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted.

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

Washington (Great Basin Desert) Montana (short-grass prairie)

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

that three of four undisturbed, previously uninvaded communities (native grassland, oak woodland, and coastal sage scrub) bordering a pipeline corridor were being invaded by the introduced species abundant on the corridor. They concluded that the corridor was responsible for giving these species a foothold from which to invade the adjacent communities. The one uninvaded community was dense chaparral. Although we did not attempt an exhaustive review of the literature on the use of corridors by introduced animals, the potential for and impacts of such use could be signiﬁcant. May and Norton (1996), for example, suggested that roads provide entry corridors for feral cats (Felis catus), dingo (Canis familiaris dingo) and foxes (Vulpes vulpes) into native Australian rainforest, where they are a threat to native animal populations. Livestock grazing Grazing by introduced ungulates is a common component of ecosystems throughout the world. In spite of its prevalence, and documented cases of rangeland degradation and changes in species composition, information on the precise causes of species change is limited. Milchunas et al. (1988) concluded that livestock grazing is not always a disturbance in rangeland ecosystems; this depends on the evolutionary history of the species present on a site and the frequency and intensity of grazing both by wildlife and by livestock. They argued that grazing by native or introduced ungulates may not be a disturbance if the species in a system have evolved with grazing, and can fully compensate for immediate removal of plant material. However, we believe that the removal of biomass results in a change in resources for the remaining or new individuals in most instances, even if this release or change in the resources is rarely measured. Further, we feel that livestock grazing is probably always at a severity or frequency different from those of wildlife grazing and thus represents an alteration in the natural disturbance regime. There have been dramatic changes in rangeland composition in the arid and semi-arid western United States and Australia over the past 150 years (Heady, 1977; Walker et al., 1981; Vavra et al., 1994). In many cases, rangelands have become dominated by introduced species (Heady, 1977; Sparks et al., 1990; Whisenant and Wagstaff, 1991; Miller et al., 1994). In spite of the great spatial extent of this conversion, the causes of most of the large-scale changes remain obscure because they occurred prior to quantitative measurements of the

417

vegetation. For example, in California, native perennial grasses were replaced almost completely by introduced Mediterranean species in the 1800s during a period of drought and over-grazing (Heady, 1977). Yet it is not clear what the relative roles of competition, drought, ﬁre, and grazing were in this conversion (Bartolome and Gemmill, 1981). In the intermountain west in North America, large areas of sagebrush (Artemisia) steppe have been converted from native shrub/perennial grass communities to low-diversity stands of the Eurasian annual grass Bromus tectorum (cheatgrass) during a long period of intense livestock grazing. Recent studies of invasion by B. tectorum into remnants of the Great Basin never grazed by livestock suggest that signiﬁcant invasion can occur without grazing, and that it is ﬁre rather than grazing that has promoted the conversion to nearly monospeciﬁc stands of cheatgrass (Whisenant, 1990a; Svejcar and Tausch, 1991). Several investigators admit that it is difﬁcult to dissect out the roles of ﬁre and grazing in causing changes in arid-zone rangelands (e.g., Sparks et al., 1990). Overall we found only 14 case studies and two reviews that explicitly correlate recent grazing disturbance with biological invasions (Table 17.2). Ten of these provided evidence suggesting at least some increase in the occurrence of introduced species with grazing. Three studies showed that invasion occurred without grazing, and could not be clearly tied to it. One study (Milchunas et al., 1990) showed that introduced species decreased in grazed compared to ungrazed rangeland. Although there is great interest among managers in understanding how the timing of livestock grazing can alter range composition including the abundance of non-native species, we found few quantitative studies in refereed journals that report on this issue. Whisenant and Wagstaff (1991) in Utah found that light and heavy spring grazing promoted the occurrence of introduced annuals in a desert saltbush community, while fall grazing (light or heavy) had no impact on the occurrence of introduced species, and fall-grazed plots were similar to ungrazed control areas. The removal of livestock after decades of livestock grazing often does not result in recovery of native species (e.g., Bartolome and Gemmill, 1981), and exclosure experiments after long-term grazing are difﬁcult to interpret in terms of assessing what the initial grazing impacts were. Despite this, long-term impacts of the cessation of grazing are variable

418

Table 17.2 Relationship between livestock grazing and occurrence of non-native species Location 1

Invader

Findings

Study type 2

Reference

California (annual grassland)

Erodium sp.

promoted by sheep grazing

Exp., Obs.

Rice (1987)

Nevada, Idaho (Great Basin Desert)

Bromus tectorum, Bromus rubens

spreads without grazing

Obs.

Whisenant (1990a); Svejcar and Tausch (1991)

Utah (Great Basin Desert)

Bromus tectorum, other annuals

grazing plus ﬁre promoted invaders

Multi-site, Cor.

Sparks et al. (1990)

Utah (Great Basin Desert)

Bromus tectorum, Halogeton sp.

increased with spring grazing, fall grazing – no effect

Obs., Exp.

Whisenant and Wagstaff (1991)

Colorado (short grass prairie)

Eurasian species

heavy grazing reduced invaders compared to ungrazed

Obs.

Milchunas et al. (1990)

Arizona (desert grassland)

Eragrostis lehmanniana

spreads with or without grazing

Obs.

Anable et al. (1992); McClaran and Anable (1992)

British Columbia (grazing land)

Centaurea diffusa

increases with light or heavy grazing, grazing poorly quantiﬁed

Cor.

Myers and Berube (1983)

Intermountain west, U.S.A. (desert)

Eurasian annuals

increase in dominance with heavy grazing

Rev.

Miller et al. (1994)

Mexico (Sonoran Desert)

Pennisetum ciliare

invasion occurs with or without grazing

Obs.

B´urquez and Quintana (1994)

Argentina (subhumid grassland)

exotics

grazing increases exotics

Gra.

Sala et al. (1986)

Argentina (montane)

exotics

grazing caused slight increase in exotics

Chr.

D´ıaz et al. (1994)

Australia (temperate grassland)

exotics

heavy grazing promotes exotics

Cor.

McIntyre and Lavorel (1994)

Australia (temperate grassland)

exotic annuals

grazing promotes annual exotics over native perennials

Rev.

Tr´emont and McIntyre (1994)

Australia (temperate woodlands)

annual grasses and forbs

increased grazing leads to increased exotics

Multi-site, Cor.

Prober and Thiele (1995)

Many areas of world

pines

grazing herbaceous layer promotes pine invasion

Rev.

Richardson and Bond (1991)

North America

1

Ecosystem type in parentheses following location. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review. 2

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

Other continents

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

and site-speciﬁc (Hatch et al., 1992; Bartolome and McClaran, 1992). Because of the increase in biomass of introduced species after the cessation of long-term livestock grazing, several investigators have suggested that carefully managed grazing may be useful in controlling introduced annual plants and promoting native perennial species. Overall, studies of grazing impacts in western North America and Australia suggest (and it has become dogma) that there tends to be a decrease in native perennial grasses and an increase in dominance by introduced annual species with heavy grazing [see Miller et al. (1994) for a review of work in North America]. However, there is a general lack of information on the actual mechanisms of conversion. An enormous amount of effort has gone into trying to understand the causes of vegetation change in rangelands and it is clear that grazing has played an important role (Walker et al., 1981; Vavra et al., 1994); but the relative roles of grazing, climatic ﬂuctuation, and ﬁre remain elusive (see Oesterheld et al., Chapter 11, this volume). Catastrophic large-scale disturbances We reviewed 12 studies that detailed the occurrence of introduced species on lands cleared for agriculture, grazing, or forestry (Table 17.3). Of the disturbances considered, land-clearing associated with agriculture is likely to be the most extreme, since it involves disruption of the soil proﬁle as well as total removal of vegetation. We address here only the invasion or persistence of weeds on such sites after abandonment. All studies reported that introduced species were abundant on severely altered landscapes for the ﬁrst several years after abandonment. This conclusion applies equally to clear-cuts and former agricultural lands, whether in deserts, temperate montane habitats, or tropical forests. Only seven studies provided information on species composition more than seven years after human activities ceased. In three of these studies (Anderson and Marlette, 1986; Hunter, 1991; Brandt and Rickard, 1994) introduced species (grasses) continued to dominate the sites for up to 50 years. All three of these were desert ecosystems that had experienced severe soil disturbance. In two studies (DeFerrari and Naiman, 1994; D´ıaz et al., 1994), one in the Paciﬁc Northwest rainforests and the other in montane grasslands, introduced species were abundant in recently disturbed areas, but declined after more than seven years and were replaced by native species. Two other studies, both in humid habitats in the southeastern

419

United States, found that dominance by persistent nonnative trees increased over time since abandonment (Doren and Whiteaker, 1990a,b; Bruce et al., 1995). Both studies were in humid coastal prairie ecosystems where soil hydrology and structure had been severely altered during farming. Three studies in tropical-island ecosystems provided no information on length of time since disturbance, but reported widespread domination of clear-cut forest patches by introduced woody species (Gade, 1985; Parnell et al., 1989; Savage, 1992). Terrestrial ecosystems: natural/near-natural disturbances Small-scale opening of patches Table 17.4 compares 17 studies that address the relationship between small-scale disturbance and invasion. Two studies examined the role of insect-induced shrub mortality in promoting introduced species, and one study examined the role of death of native bunch grasses in providing canopy gaps for establishment. The other fourteen studies examined colonization by native and introduced species on areas subject to animal disturbance, or mechanical disturbance that simulated small-scale animal disturbance. Surprisingly, we found almost no studies of colonization of non-indigenous species in tree-fall gaps in forested ecosystems. We suspect that there are more studies that are appropriate here, but before the last ten years many investigators did not distinguish between introduced and native colonizers of small gaps or areas of disturbed soil. In the case of canopy gaps not associated with soil disturbance, one study in coastal California found that insect-induced mortality of the native leguminous shrub Lupinus arboreus led to formation of nitrogenrich soil patches which were rapidly colonized by introduced annual grasses and thistles rather than native prairie species (Maron and Connors, 1996). In a second study of insect-induced mortality of native plants in a shortgrass prairie ecosystem in Colorado, grub outbreaks did not affect the richness or abundance of introduced species (Milchunas et al., 1990). In a Californian study (Peart, 1989), mortality of native bunchgrasses increased recruitment of introduced perennial grasses. Our 14 studies of small-scale soil disturbance included a total of 21 different habitats. Disturbance promoted invasion into 16 of these, had no effect on invasion in three others and decreased the success of invaders in only two sites. The most common

420

Table 17.3 Studies examining the occurrence of invasive non-native species in anthropogenic, catastrophic disturbances of terrestrial environments Location 1

Disturbance

Effect on site

Impact on invaders

Study type 2

Reference

Nevada (Mojave Desert)

bombs

total destruction

persistent stands of introduced annual grasses

Obs.

Hunter (1991)

Washington (Great Basin Desert)

agriculture

total destruction

Bromus tectorum domination, even after 47 yr

Obs.

Brandt and Rickard (1994)

Washington (temperate rain forest and montane)

logging/clear cuts

canopy removal

increase within clear-cuts for 3–7 yr, decline thereafter

Multi-site Cor.

DeFerrari and Naiman (1994)

Utah (Great Basin Desert)

agriculture

total destruction

Agropyron desertorum persistent – 50 yr

Obs.

Anderson and Marlette (1986)

Texas (humid prairie)

agriculture

total destruction

Sapium sebiferum increases with time since Multi-site agriculture Obs.

Bruce et al. (1995)

Montana (montane coniferous forest)

logging/clear cuts

canopy removal

up to 60% cover even 7 yr later

Multi-site Obs.

Forcella and Harvey (1983)

Florida (ﬂooded prairie)

agriculture

vegetation destruction, soil alteration

persistent invasion by Schinus terebinthifolius

Obs.

Doren and Whiteaker (1990a,b)

Argentina (montane grassland)

agriculture

total destruction

Eur. annuals common in recently fallow ﬁelds, rare in 25 yr old site

Obs.

D´ıaz et al. (1994)

Argentina (inland pampas)

agriculture

total destruction

Eur. weeds dominated up to 5 yr after abandonment

Obs.

D’Angela et al. (1986, 1988)

Mauritius (insular tropical forest)

logging, agriculture

partial and total destruction

exotics dominate all cut forests

Obs.

Parnell et al. (1989)

Rodrigues (insular tropical forest)

logging, agriculture

partial and total destruction

exotics dominate all cut forests

Ane.

Gade (1985)

W. Samoa (insular tropical forest)

logging, agriculture

partial and total destuction

vine-tangles abundant in 2º forest

Obs.

Savage (1992)

North America

1

Ecosystem type in parentheses following location. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review. 2

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

Other continents

Location 1

Invader

Disturbance

Effects of disturbance

Study type 2

Reference

North America California (coastal prairie)

Eur. annuals

exotics dominate dead shrubs patches

Obs., Exp.

Maron and Connors (1996)

California (coastal grassland)

Eur. grasses

insect-induced shrub mortality canopy death

Exp.

Peart (1989)

California (coastal grassland) California (coastal grassland, dune, scrub) California (serpentine grassland)

Eur. grasses Carpobrotus edulis Bromus mollis

gopher mounds gopher mounds + mechanical gopher mounds

increases establishment of perennial invaders enhanced establishment of invaders promotes invasion into grassland; no effect on dune and coastal scrub promotes invasion of native grassland

Exp. Obs., Exp.

Peart (1989) D’Antonio (1993)

Obs., exp.

California (valley grassland) California (coastal shrubland, Sierran foothill) Oregon (coastal prairie)

Erodium sp. Cytisus scoparius

Washington (coastal prairie) Washington (coastal prairie) Nevada (Great Basin Desert) Montana (short grass prairie)

Cytisus scoparius Senecio vulgaris Bromus rubens Centaurea maculosa Eur. species

gopher mounds gopher mounds + mechanical mechanical (simulated rodent) mechanical mechanical gophers ground squirrels, other mammals insect-induced death of natives bison wallows

Hobbs et al. (1988); Hobbs and Mooney (1991) Rice (1987) Bossard (1991)

reduced invasion promotes invasion enhanced invasion increased spread of invader into native prairie no effect on invasion

Exp. Exp. Obs. Obs.

Parker (1996) Bergelson et al. (1993) Hunter (1991) Tyser and Key (1988)

Obs.

Milchunas et al. (1990)

colonized by 3 ruderal species

Obs.

Wein et al. (1992)

mechanical

enhanced invasion

Exp.

mechanical

enhanced invasion

Exp.

Hobbs (1989); Hobbs and Atkins (1988) Burke and Grime (1996)

mechanical

no effect

Exp., Obs.

Belsky (1986)

Colorado (shortgrass prairie) N.W. Territories: Canada (boreal forest) Other continents W. Australia (Eucalyptus woodland, heathland) United Kingdom (grassland) Tanzania (savanna)

Senecio jacobaea

Eur. ruderals

Avena fatua, Ursinia sp. herbaceous exotics herbaceous exotics

Obs., Exp. promotes invasion reduced invasion in coastal site, enhanced it Exp. in Sierra site promotes invasion of prairie Exp.

McEvoy and Rudd (1993)

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Table 17.4 Studies reporting relationship between small-scale disturbance and presence or abundance of invasive non-native species

1

Ecosystem type in parentheses following location. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review. 2

421

422

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

invaders to be promoted were European annuals or biennials. In one of the studies that showed no impact of disturbance on invasion, D’Antonio (1993) found in coastal California that herbivory on seedlings of an aggressive non-indigenous species was so great in two of her three study habitats that it overrode any effect of disturbance in promoting invasion. The two studies which showed a negative effect of disturbance on invasion both involved invasion of the shrub Scotch broom (Cytisus scoparius) into the western United States. Bossard (1991) found that disturbance reduced establishment of broom in one site in California because seed predators and herbivores were attracted to disturbed microsites and removed most of the broom seeds from them. Parker (1996) found that localized disruption of the cryptogamic crust in a Washington coastal-prairie site decreased rather than increased establishment of Scotch broom. Gap dynamics are known to be important in promoting community- or landscape-scale coexistence in a wide range of communities from prairies to tropical rainforests, and it is likely that small-scale disturbances are important for the persistence within the landscape of short-lived, well-dispersed non-indigenous species. However, as in the case of anthropogenic-disturbance corridors (roads, trails), the role that small-scale disturbance plays in invasion of adjacent undisturbed habitat by introduced species is probably dependent on habitat and species. Most of the studies we found suggested that the introduced species studied could invade away from soil disturbances into “undisturbed” adjacent habitat. Hunter (1991) provided evidence that disturbance by gophers (Thomomys bottae) increases abundance of red brome (Bromus rubens) in ungrazed sites in the Nevada desert, but this species also appears to be able to invade undisturbed habitat where gopher disturbances are lacking (Beatley, 1966; Svejcar and Tausch, 1991). Peart (1989) reported that, although establishment of the three introduced grass species he studied was enhanced by gopher disturbances, two of the species were able to colonize undisturbed prairie, particularly areas dominated by annual grasses. Reproduction of individuals on gopher mounds was high, and may be an important source of seeds for colonization of surrounding areas. D’Antonio (1993) found that seedlings of the introduced succulent Carpobrotus edulis required rodent disturbance to establish in a California coastal grassland; once established, however, this long-lived species could grow out over and suppress herbaceous individuals within the

adjacent grassland. By contrast, Hobbs and Mooney (1991) and Hobbs et al. (1988) found that, although gopher disturbances promoted invasion of California serpentine grasslands by the Eurasian grass Bromus mollis, this species did not spread into undisturbed portions of these sites. Indeed, further disturbance of an initial gopher disturbance led to the disappearance of B. mollis from the local patch. Likewise, McEvoy and Rudd (1993) found that small-scale soil disturbance led to localized outbreaks of the Eurasian biennial Senecio jacobaea in an Oregon coastal prairie, but that it could not invade the dense adjacent undisturbed vegetation. Any habitat with burrowing organisms is likely to have a “natural” regime for occurrence of local-scale disturbances. The importance of size, frequency, and pattern of small-scale disturbances in relationship to invasion rates and persistence of invaders has rarely been examined. Bergelson et al. (1993) demonstrated experimentally that the abundance of an introduced annual forb was greater when disturbances were large (30 cm diameter vs. 5 or 15 cm) and randomly arrayed over the landscape. Peart (1989), however, found no effect of gap size on colonization success by introduced grasses in a California prairie, whereas McConnaughay and Bazzaz (1987) suggested that seed size of the colonizer determines the importance of gap size. In addition, there are no examples to show how alteration of “natural” local-scale disturbance regimes have interacted with invasion. Fires Numerous studies have examined the inﬂuence of ﬁre on the susceptibility of arid and semi-arid ecosystems to invasion (Table 17.5). Most do not quantify ﬁre frequency, intensity, or size, but simply compare invasion in burned versus unburned sites, or before versus after ﬁre. In 18 out of 24 studies involving either controlled burns, multi-site comparisons, or long-term observations after wildﬁre, ﬁre was found to promote or enhance invasion by non-native species. In at least three studies (Hobbs and Atkins, 1990; Trabaud, 1990; Wein et al., 1992), dominance by introduced species was short-lived after ﬁre, and native species became dominant within a few years. We identiﬁed 12 studies that reported invasion occurring without ﬁre or any apparent change in ﬁre regime, but where ﬁres promoted the spread or increase in density of the species in question. The ecosystem types most readily invaded without ﬁre are desert habitats such as the Great Basin of western

Location 1

Invader

Impacts of ﬁre

Study type 2

Reference

California (Mojave Desert)

Bromus rubens, Schismus barbatus

invades without ﬁre; enhanced by ﬁre

Multi-site, Cor.

Brown and Minnich (1986)

California (Sierran foothills)

Eur. annual grass and forbs

grasses reduced by ﬁre, forbs enhanced

Controlled burn

Parsons and Stohlgren (1989)

California (maritime chaparral)

Carpobrotus edulis

germination reduced by ﬁre but invasion promoted by ﬁre

Obs., Exp.

Zedler and Scheid (1988); Hickson (1988); D’Antonio et al. (1993)

California (Sierran foothills)

Cytisus scoparius

promoted by ﬁre

Exp.

Bossard (1991)

Oregon (Great Basin Desert)

Bromus tectorum

maybe invasion without ﬁre; enhanced by ﬁre

Obs., Ane.

Klemmedson and Smith (1955)

Idaho (Great Basin Desert)

Bromus tectorum + other annuals

invades small amount without ﬁre; enhanced by ﬁre

Multi-site, Cor.

Whisenant (1990a)

Utah (Great Basin Desert)

Bromus tectorum + other annuals

maybe enhanced by ﬁre (sites also heavily grazed)

Multi-site, Cor.

Sparks et al. (1990)

Nevada (Great Basin Desert)

Bromus rubens

invades without ﬁre; promoted by ﬁre

Cor.

Beatley (1966); Hunter (1991)

Nevada (Great Basin Desert, island in lake)

Bromus tectorum, Bromus rubens

invades without ﬁre

Obs.

Svejcar and Tausch (1991)

Intermoutain west U.S.A. (Great Basin Desert)

Taeniatherum asperum

invades without ﬁre some sites; other sites it needs ﬁre to invade

Ane.

Young and Evans (1971)

Arizona (Sonoran Desert grassland)

Eragrostis lehmanniana

germination enhanced by ﬁre; spread maybe enhanced by ﬁre; can spread without ﬁre

Exp., Controlled burn

Ruyle et al. (1988); Cable (1971); Anable et al. (1992)

Sonora, Mexico (Sonoran Desert)

Pennisetum ciliare

invades slowly without ﬁre; enhanced by ﬁre

Ane.

B´urquez and Quintana (1994)

South Dakota (shortgrass prairie)

Bromus japonicus

invades without ﬁre; suppressed by ﬁre

Exp., Controlled burn

Whisenant (1990b); Whisenant and Uresk (1990)

Hawaii (seasonal submontane woodland)

perennial C4 grasses

some invade without ﬁre, all promoted by ﬁre

Multi-site, Obs.

Hughes et al. (1991); Smith and Tunison (1992)

N.W. Territories, Canada (boreal forest)

Eur. ruderals

large natural ﬁre promoted temporary invasion by 3 species

Obs.

Wein et al. (1992)

North America

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Table 17.5 Studies examining the relationship between ﬁre and invasions by non-native species in terrestrial ecosystems (excluding riparian)

Other continents Hyparrhenia rufa

invades without ﬁre; promoted by ﬁre

Obs.

Bilbao (1995)

annual grasses

ﬁre promotes short-lived ﬂush of intr. annuals

Obs.

Trabaud (1990)

W. Australia (shrublands, heathland)

Eur. annual grasses

decreased invasion shrubland; no effect in heathland

Exp.

Hobbs and Atkins (1991) continued on next page

423

Venezuela (llanos) France

424

Table 17.5, continued Location 1

Invader

Impacts of ﬁre

Study type 2

Reference

W. Australia (Eucalyptus woodland)

ann. forbs, grasses

enhanced by ﬁre

Obs.

Bridgewater and Backshall (1981)

W. Australia (Banksia woodland)

Eur. annuals

increased by ﬁre but for short time only

Controlled burn

Hobbs and Atkins (1990)

W. Australia (Eucalyptus woodland)

Eur. annuals

decreased by ﬁre

Controlled burn

Hester and Hobbs (1992)

E. Australia (grassland)

Eur. annuals

enhanced by ﬁre (site heavily grazed by livestock)

Obs.

Lunt (1990)

South Africa (fynbos)

Pinus radiata

invasion very slow without ﬁre; enhanced by ﬁre

Obs.

Richardson and Brown (1986)

South Africa (fynbos)

Pinus halepensis

invasion very slow without ﬁre; enhanced by ﬁre

Obs.

Richardson (1988)

South Africa (fynbos)

Banksia

enhanced by ﬁre

Obs.

Richardson et al. (1990)

1

Ecosystem type in parentheses following location. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review. 2

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

North America and the Sonoran Desert of Mexico and Arizona. Whisenant’s (1990a) study of invasion by cheatgrass of the Snake River plains in Idaho is an excellent example of invasion in the absence of ﬁre (and livestock grazing), but clear enhancement of dominance by recurrent ﬁre. In addition to deserts, we found examples of ﬁre promoting invasion of Mediterranean ecosystems such as the fynbos of South Africa and maritime chaparral of California. Invasion by introduced species occurs only exceptionally after chaparral ﬁres in California and has been reported only for maritime chaparral (Hickson, 1988; Zedler and Scheid, 1988; D’Antonio et al., 1993). Invasion here does not appear to be due to an alteration in natural disturbance regimes, but rather to the more open form of the maritime chaparral with a less diverse assemblage of native post-ﬁre species (Tyler, 1994). Numerous investigators have now shown that natural ﬁre in the South African fynbos promotes invasion of these sites by introduced pines and Australian species of Acacia and Hakea (Table 17.5). Indeed, Richardson and Bond (1991) reported numerous cases from throughout the world where range extension by pines is promoted by ﬁre. Because of the longevity of pines, these post-ﬁre invasions are likely to represent long-term successional changes in natural communities. This is also true for invasion of maritime chaparral in California by Carpobrotus edulis, which is an aggressive competitor against native species (D’Antonio and Mahall, 1991); its invasion likely represents a long-term successional change at these sites. In only four studies was ﬁre found to cause a decrease in abundance of introduced species. Hester and Hobbs (1992) reported that ﬁre and manipulation of ﬁre regimes may be a useful management tool to control non-native species in Australia: most of the native species at their sites responded positively to ﬁre, whereas the introduced species present on the sites were negatively impacted. By contrast, in a management-oriented study in California (Parsons and Stohlgren, 1989), ﬁre caused a decrease in abundance of introduced annual grasses, but an increase in abundance of introduced forbs, with no net positive effect on native species. The season of burning was found to be an additional factor inﬂuencing the likelihood of invasions by non-indigenous species (Parsons and Stohlgren, 1989; Hobbs and Atkins, 1990; Whisenant, 1990b). Thus, where the life forms of the available suite

425

of introduced species are diverse, their responses to ﬁre are likely to be variable. Overall, we found few studies to support the role of ﬁre in promoting invasion into most ecosystems which have a long history of ﬁre, such as the mixed conifer forests of western North America, California chaparral, prairies of the midwestern United States, heathland and open Eucalyptus forest in Western Australia, and African savannas. Indeed, to generalize about the importance of ﬁre would require an extensive review of all studies documenting changes in species composition in response to ﬁre, and most such studies do not even mention whether the responding species are introduced or native. Virtually all of the studies we report on here suggesting that ﬁre events or recurrent ﬁre promoted long-term presence of invaders were in deserts, or ecosystems such as oceanic islands (see Hughes et al., 1991) where the role of ﬁre as a selective force in plant evolution appears to have been minimal. The mountain fynbos of South Africa was the only ecosystem that is being invaded on a very large scale by ﬁre-promoted species, and where ﬁre has been a recurrent phenomenon prior to invasion (we do not consider the invasion by C. edulis into maritime chaparral to be a large-scale invasion, because this form of chaparral is restricted). Although ﬁre suppression has been an important part of land management in many regions of the world, there are few published reports of how a decrease in ﬁre frequency might promote invasion by nonindigenous species. Bruce et al. (1995) suggested that ﬁre suppression in Texas coastal prairies might be allowing invasion by Chinese tallow tree, Sapium sebiferum. Their study sites, however, were all severely disturbed by agriculture and hydrological modiﬁcation prior to invasion. Although ﬁre suppression in Californian chaparral has caused variation in ﬁre size and intensity, and in germination frequencies of various native species (Moreno and Oechel, 1991), there is no evidence that it has favored invasion by non-indigenous species. However, more studies are needed in this area. Storm damage and other large-scale disturbances We found almost no quantitative studies examining the relationship between canopy damage caused by storms and invasion of otherwise intact natural communities by non-indigenous species. Williams (1993) found that closed-canopy forests in Virginia in the eastern United States were invaded by the shadeintolerant Asian tree Paulownia tomentosa only after

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Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

these forests had been severely damaged in August of 1969 by Hurricane Camille. Likewise, native forests on Tahiti have been almost completely transformed to monospeciﬁc stands of the New World tree Miconia calvescens; accounts suggest that the rise to dominance by Miconia followed severe hurricane damage to the native forest canopy (J. Schwartz, pers. commun.). Most accounts of invasion following hurricane damage are similarly anecdotal. We found no information on the role of landslides in promoting or limiting invasion, although we believe that this phenomenon is important. For example, in coastal California, landslides are rapidly colonized by Pampas grass, Cortaderia jubata, an invader from South America (D’Antonio, pers. observ.). Summary of terrestrial plant invasions and disturbance In spite of the common belief that disturbance is a precursor to invasion in plant communities, we found numerous examples where invasion could occur without disturbance. This was particularly true in desert habitats. Crawley (1987) found for the British ﬂora that invading species were more common in plant communities where the average cover was low – which is certainly true of desert habitats compared to grasslands or forest. In addition to the studies outlined above, there are several other reports of plant species that can invade without any apparent disturbance. A detailed list has been provided by Rejm´anek (1989). Several of the reports of species invading forested sites without any obvious disturbance are on islands (e.g., Lorence and Sussman, 1986; Huenneke and Vitousek, 1990; Rejm´anek, 1996). Are these also “open” or “undersaturated” systems? While this is possible, we recommend interpreting such studies with caution. In many of the reports of invasions into undisturbed island ecosystems, sites adjacent to the invaded forests have undergone catastrophic disturbance, are full of nonnative species, and serve as reservoirs for dispersal into undisturbed sites. Also, islands have a long history of habitat modiﬁcation, including elimination of native avifauna (e.g., Olson and James, 1982; Steadman and Kirch, 1990; James, 1995) and alteration of native invertebrate communities, which may have reduced the resistance of the native biota to invasions (e.g., Lake and O’Dowd, 1991; D’Antonio and Dudley,

1995). Overt disturbance may not be necessary to promote invasions on islands today, but present-day communities may lack mechanisms that previously might have made them more resistant to invasion. Few studies have actually attempted to determine the mechanisms through which disturbance might inﬂuence invasion. An understanding of these mechanisms would require elucidation of factors within undisturbed communities that make them more or less resistant to invasion. Disturbance might inﬂuence invasion by: (1) eliminating predators or herbivores that would otherwise reduce or eliminate the invaders; (2) reducing competitive pressure from pre-established species; (3) stimulating germination of seeds of invaders; and/or (4) altering resources to levels that favor the invaders rather than pre-existing individuals. Where mechanisms have been investigated they have been found to be complex and site-speciﬁc. For example, D’Antonio et al. (1993) investigated the mechanisms through which ﬁre appears to promote invasion of maritime chaparral in California by the South African succulent Carpobrotus edulis. They found that high temperatures killed C. edulis seeds, but that invasion occurred anyway because of the existence of microsites where soil temperatures were not elevated during ﬁre, and because seeds of this ﬂeshy-fruited species are readily dispersed into burned sites by abundant native frugivores. Thus, ﬁre intensity and fuel distribution should have a signiﬁcant inﬂuence on whether or not invasion will occur. In addition, ﬁre altered soil conditions and favored invasion by promoting C. edulis seedling growth, so that herbivores could no longer crop back new invading plants. In a grassland site, D’Antonio (1993) found that soil disturbance by rodents promoted invasion by C. edulis because it stimulated seed germination and decreased seedling competition between C. edulis and Eurasian grasses. Rodent-caused soil disturbance had no overall net effect on invasion by C. edulis in a dune and a coastal-scrub site where herbivory by rabbits and deer consistently eliminated emerging seedlings: disturbances were not large enough to decrease the density of herbivores and thereby enhance invasion. Predicting when and why disturbance promotes invasion will thus require many more such mechanistic studies because currently we have very little understanding of which mechanisms are more likely to be responsible for community resistance to invasion and how disturbance regime interacts with these.

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Riparian and wetland invasions The relationship between disturbance and invasion by non-indigenous species into aquatic ecosystems appears to be substantially different from that in many terrestrial ecosystems, primarily owing to the pervasive role that disturbance has in structuring community composition. Hydrological dynamics are of fundamental importance to aquatic systems, and may constitute direct-disturbance impacts, as in destructive ﬂooding following spates or snowmelt (likewise, disruption of ﬂows, such as the physical damming of channels by landslides, beaver dams, or human structures, alters the relationship between hydrology and biota). While the degree to which surface hydrology modiﬁes habitat and community structure is the basic difference between aquatic and terrestrial systems, many other disturbances or perturbations are shared with upland systems, including ﬁre, forest blowdown, or grazing. In deﬁning disturbance in riparian and wetland habitats, we again restrict our discussion mainly to physical processes which remove tissue and open space for colonization. Others have broadened the deﬁnition to include natural physical stresses, such as sustained drought (Stanley et al., 1994) and chronic changes (usually increases) in nutrient supplies, sediment deposition, and chemical pollution. For the purpose of this review, we consider most of these to be non-episodic perturbations which constitute changed conditions for organisms rather than punctuated mechanisms structuring these systems. Riparian ecosystems Riparian ecosystems are considered disturbanceprone; nearly every unregulated watercourse experiences substantial, often catastrophic, variation in discharge, which can remove established biota from in-stream and ﬂoodplain environments. Indeed, most stream ecologists consider that native aquatic and riparian organisms depend upon periodic ﬂooding for creation and maintenance of suitable habitat, removal of excess deposited material, and reduction in density of competitors (Fisher et al., 1982; Resh et al., 1988; Poff, 1992). Riparian vegetation alongside unregulated or regularly disturbed rivers is early-successional in character, generally composed of fast-growing, shortlived, or disturbance-tolerant species which can reestablish populations before the next scouring event (Hupp and Osterkamp, 1985). These are traits frequently shared with invasive, non-indigenous plants in terrestrial systems. In at least 18 of the 38 studies we reviewed which examined the status of introduced

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plants in riparian corridors, natural ﬂood regimes in some manner promoted the growth of invasive species (Table 17.6). Another six studies found that exotic riparian species were favored on bare soils if space was opened by physical disturbance other than the ﬂood regime, and in most of the remaining studies the role of substrate scouring was not considered. Studies in which natural disturbance favored invasion are divided between cases of non-indigenous species tending to remain within the most frequently disturbed near-stream environment, and studies which document some expansion into less disturbed assemblages following initial establishment in the riparian zone. In the Ardour River in France, Tabacchi (1995) showed that introduced species comprised 20% of the active streambank vegetation, but were poorly adapted to tolerate drought away from the channel. Older arms of the river included fewer non-native species, and these were more sensitive to local loss from ﬂooding. Likewise, the Mediterranean tree Fraxinus ornus is dispersed downstream by water ﬂow and grows abundantly (including regrowth from damaged boles) in the high-energy channel habitat; but 65 years after the introduction of this tree into new channels, it had not expanded into adjacent riparian vegetation (Th´ebaud and Debussche, 1991). DeFerrari and Naiman (1994) and Planty-Tabacchi et al. (1996) reported that nonindigenous species were abundant in riparian zones and in particular on gravel bars within stream beds in Washington and Oregon (and also in France), but that only infrequently did these species penetrate the intact mesic forest away from the stream beds. Microstegium vimineum is an introduced, shade-tolerant annual grass, which was reduced by 50% during ﬂooding in a North Carolina ﬂoodplain, then doubled its previous abundance the following year through high seed set. Still, it has not invaded the native riparian woodland nearby (Barden, 1987). In Great Britain, Japanese knotweed (Fallopia japonica = Reynoutria japonica), found naturally on sparsely vegetated lava ﬂows, is a riparian ruderal species that appears to be replaced by other vegetation within 50 years (Palmer, 1994). In most of these cases, the invaders tend not to alter the character of the riparian ecosystem substantially, nor to displace native species. Other non-indigenous streamside plants, or even the same species in a different environment, may be less benign. Numerous studies show that, while initial establishment of plants has occurred in regularly disturbed riparian areas, invasion often proceeds

428

Table 17.6 Non-indigenous plants associated with riparian ecosystems, and relationships with disturbance and ﬂow regimes Location 1

Setting

Invader

Findings

Study type 2

Reference

New York (temperate forest)

woodland/urban streams and canals

Lythrum salicaria

establishes in disturbed soil on dikes/ditches, spreads into streams/marshes

Obs.

Stuckey (1980)

Pennsylvania (temperate forest)

woodland marsh and stream

Lythrum salicaria

ﬂuctuating water levels favor germination on open soils; spreads into marsh

Obs.

Thompson (1991)

Virginia (temperate forest)

woodland ﬂoodplain

Paulownia tomentosa

hurricane opens space for germination, but declines with succession

Obs.

Williams (1993)

North Carolina (temperate forest)

woodland ﬂoodplain

Microstegium vimineum

ﬂood and ﬁre open space for germination; poor colonization in established vegetation

Obs.

Barden (1987)

Texas (semi-arid scrub)

newly formed river delta at reservoir

Tamarix sp.

explosive germination on stabilized soils

Obs.

Robinson (1965)

Ohio (temperate forest)

restored riparian marsh, Typha spp. agricultural

Rocky Mts. (arid montane)

springs and streams in rangelands

Eleagnus angustifolia established on stabilized levees, subsequently invaded riparian areas

Obs.

Knopf and Olson (1984)

Colorado (arid montane)

high plains riparian areas, grazing, and some agriculture

Salix spp. (Eurasian)

Exp., Obs.

Shafroth et al. (1994)

Utah (Great Basin Desert)

many former Populus riparian zones

Tamarix sp., natural high salinity tolerated better by exotics; Eleagnus angustifolia physical disturbance may not be required for establishment

Cor.

Carman and Brotherson (1982)

Southwest U.S.A.

desert riparian areas

Tamarix sp.

ﬂooding promotes germination, but establishment requires sustained moisture

Obs.

Horton et al. (1960)

Arizona (Mojave Desert)

Populus riparian, water diverted for agriculture and urban use

Tamarix sp.

declining water level favors Tamarix, which can use unsaturated soil water unlike native Populus and Salix

Exp.

Busch et al. (1992)

Southwest U.S.A.

desert riparian, Tamarix sp. agricultural, rangelands

water diversion and impoundment reduces ﬂood frequencies, and is associated with expansion of Tamarix

Obs.

Everitt (1980); numerous others

Arizona (Sonoran Desert)

Populus/Platanus riparian, grazed

Cynodon dactylon

Bermuda grass stabilizes substrate, provides ﬂood refuge for macrophytes

Cor.

Dudley and Grimm (1994)

Arizona (Sonoran Desert)

mixed riparian forest, nature reserve

Tamarix sp., Paspalum sp., Cynodon dactylon

ﬂood caused greater mortality to Tamarix than native Populus and Salix; Paspalum and Cynodon increased rapidly

Obs.

Stromberg et al. (1993)

North America

invasive Typha declines as planted and natural riparian Obs. vegetation close space

continued on next page

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

cuttings and seedlings survived wide range of hydrological conditions; expands by vegetative growth in moist sites

Niswander and Mitsch (1995)

Location 1

Setting

Invader

Findings

Study type 2

Reference

Arizona (Mojave Desert)

mesquite riparian area

Tamarix sp.

ﬁre reduces Tamarix more so than Prosopis, but regenerates rapidly

Cor.

Busch (1995)

Southwest U.S.A.

desert riparian areas; diversions for agriculture, etc.

Tamarix sp. and dams eliminated ﬂooding; Tamarix germinates more Eleagnus angustifolia successfully than Populus, Eleagnus highly shade-tolerant so establishes under canopy

Obs.

Howe and Knopf (1991)

Rocky Mts. U.S.A. (Great Basin/montane)

riparian

Eleagnus angustifolia Eleagnus phenology more ﬂexible than Populus; tolerates drought and shading; takes advantage of ecological variation

Obs.

Shafroth et al. (1995)

Colorado (semi-arid montane)

cottonwood riparian, dams for agriculture

Tamarix sp.

stabilized ﬂow associated with Tamarix increase and Populus decline, channel narrows and deepens

Obs.

Snyder and Miller (1992)

Paciﬁc Northwest (temperate forested montane coniferous forest) riparian areas, minor logging

many species

24–30% of riparian species are aliens, associated with natural ﬂooding; some invasion of mature vegetation, but most species in “young” communities

Cor.

Planty-Tabacchi et al. (1996)

California (Great Basin)

groundwater pumped for export

annual grasses

lowered water associated with decline in native riparian plants; increase in exotic annuals

Cor.

Manning (1992)

California (sage scrub chaparral)

ephemeral channel, agriculture and urban development

Schinus molle

road and agriculture run-off maintains continuous seepage; allows slow establishment of pepper tree

Obs.

Howard and Minnich (1989)

California (coastal mesic forest)

permanent river in nature preserve

Rana catesbiana, macrophytes

reduced ﬂow and scouring during drought favor plant Cor. build-up, then favoring bullfrog increases; both decline with return to natural ﬂooding

Northern Territory, Australia

arid watercourses

Tamarix aphylla

rare ﬂood/wet year dispersed seed and favored establishment, may displace native river red gum (Eucalyptus camaldulensis)

Obs.

Grifﬁn et al. (1989)

South Australia

woodland riparian

Pittosporum undulatum

native invasive species; increases during drought, although native Tristaniopsis resists ﬂooding better

Obs.

Melick and Ashton (1991)

South Australia

woodland riparian

many spp.

ﬁre, clearing, and fragmentation all area associated with increase in exotics

Obs.

Recher et al. (1993)

Western Australia

seasonal tropical riparian/marshlands

Mimosa pigra

colonizes “undisturbed” marsh, invades then into riparian, and ultimately uplands, alters vegetation and associated wildlife

Cor.

Braithwaite et al. (1989)

Great Britain

ﬂoodplains, farmland

Fallopia japonica

establishes on soils opened by ﬂood scour, replaced by Obs. other vegetation in <50 yrs

Kupferberg (1994)

Other continents

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Table 17.6, continued

Palmer (1994)

429

continued on next page

430

Table 17.6, continued Setting

Invader

Findings

Study type 2

Reference

France (also compared with Oregon and Washington, USA)

wooded ﬂoodplain, fairly natural but developed uplands

many spp.

“ruderal” species favored in open and dry streamside sites, cleared by ﬂoods

Cor.

Tabacchi (1995); Planty-Tabacchi et al. (1996)

Southern France

Mediterranean riparian woodlands

Fraxinus ornus

variable ﬂows open space for recruitment, and mature stems resprout after ﬂoods

Obs.

Th´ebaud and Debussche (1991)

Portugal

Mediterranean shrubland, regulated river, agricultural

Paspalum paspalodes, frequencies increased with agricultural disturbance, bank alteration Myriophyllum aquaticum, others

Cor.

Ferreira and Moreira (1995)

Czech Republic

temperate woodlands, agriculture and urban

Heracleum mantegazzianum, Impatiens glandulifera, Reynoutria japonica, R. sachalinensis

on open soils (canals, roadways), but spread into less disturbed lands; species vary in dependence upon disturbance level and ability to invade less disturbed sites

Cor.

Pyˇsek and Prach (1994)

Czech Republic

woodlands, agriculture and urban

many spp.

most species invaded faster on open soil from ﬂoods, roads

Cor.

Pyˇsek and Prach (1993)

Poland

riparian

many spp.

17 of 33 non-native species present were associated with open banks resulting from natural ﬂooding

Cor.

Kornas (1990)

Great Britain – Wales

5 rivers, diverse land uses

Reynoutria japonica

most frequent on disturbed/open ground, natural or managed vegetation resisted invasion

Cor.

Beerling (1991)

South Africa

desert riparian woodlands in national park

abundances of all four were associated with soil Ricinus communis, disturbance, road cuts, and artiﬁcial (stable) water Nicotiana glauca, Prosopis spp., Datura supply; perennials have most impact on native species innoxia

Obs.

Boyer (1989)

South Africa

semi-arid ﬂoodplain, plantation forests

Pinus patula, Acacia mearnsii

exotic trees decrease ﬂow, increase diel ﬂow ﬂuctuation

Exp./Obs.

Dye and Poulter (1995)

Cape Province, South Africa

riparian, fynbos woodland

Pittosporum undulatum

water use efﬁciency similar to native Cunonia capensis, may invade without disturbance

Cor.

Smith (1990)

1

Habitat types in parentheses after location. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review. 2

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

Location 1

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

into other, relatively undisturbed areas. For instance, Beerling’s (1991) analysis of Fallopia japonica in Wales indicates that it regularly occurs outside of the riparian environment, but at lower densities than within the riparian corridor. In semi-natural Polish habitats, Kornas (1990) found similarly that naturally disturbed riparian zones held 17 of the 33 non-indigenous plants; many of these species passed a “ruderal phase” in such sites before subsequent spread. In Australia, the highly destructive Mimosa pigra ﬁrst invades sedgeland and riparian areas, then spreads broadly into many habitats, from fully aquatic locations to paperbark (Melaleuca spp.) and monsoonal forests (Braithwaite et al., 1989). In the United States, purple loosestrife (Lythrum salicaria), while not necessarily associated with ﬂood-denuded streambanks, germinates well on open soil and appears to establish initially along canals, dikes, and other “disturbed” areas with frequent ﬂuctuation of water levels; thereafter, it expands into less disturbed wetland environments, where it displaces native species (Stuckey, 1980; Thompson, 1991). Heracleum mantegazzianum and other invasive species in Europe similarly escaped from gardens to increase exponentially in the absence of competition in frequently ﬂooded riparian areas (and other disturbed corridors). They have now moved out into vegetated upland habitats (Pyˇsek and Prach, 1993, 1994). In the Rocky Mountain states in the western United States, Russian olive (Eleagnus angustifolia) colonizes levees, disturbed riparian zones, ditches, and mined sites (and is planted for wildlife habitat) (Shafroth et al., 1995); it then expands from disturbed sites into springs, riparian woodland, and upland areas, forming thick forests (Knopf and Olson, 1984). It invades mature groves of cottonwood (Populus deltoides), potentially replacing them, as a consequence of tolerating drier and shadier conditions, and having a more ﬂexible phenology which allows it to tolerate a wide range of ecological stresses (Howe and Knopf, 1991; Shafroth et al., 1995). A single study in which disturbance appeared simply to inhibit establishment of an exotic species involved Pittosporum undulatum in rainforest in Australia. There a native riparian tree, Tristaniopsis laurina, dominates riparian habitat because of its greater resistance to ﬂooding, whereas Pittosporum, native to other parts of south-eastern Australia, survives better away from water (but is eliminated by ﬁre) (Melick and Ashton, 1991). The relationship between ﬂood disturbance and the establishment of non-indigenous riparian species is

431

more complex than suggested thus far, as illustrated by the genus Tamarix (tamarisk, saltcedar). Species of Tamarix are invaders into riparian areas throughout the western United States (Robinson, 1965) and elsewhere. They are particularly common in arid and semi-arid river channels, where infrequent large ﬂoods disperse the abundant seeds downstream to open moist sites that are ideal for germination (Horton et al., 1960; Robinson, 1965; Grifﬁn et al., 1989). Given its ability to tolerate high salinities (Carman and Brotherson, 1982), exploit unsaturated soils (unlike native phreatophytes; Busch et al., 1992), and recover from ﬁre (Busch, 1995; Horton, 1977), tamarisk is widely assumed to displace native trees such as Populus spp. and Salix spp. However, tamarisk is most abundant in channels where river ﬂows have been regulated by upstream dams and diversions, and numerous researchers have suggested that the absence of natural ﬂoods has promoted its dominance, as well as the decline of native riparian trees, under otherwise “undisturbed” conditions (Everitt, 1980; Howe and Knopf, 1991; Snyder and Miller, 1992). Furthermore, two studies indicate that occasional high ﬂows (recurrence interval 10 years) actually cause greater mortality to Tamarix than to co-occurring native trees (Stromberg et al., 1993, T. Dudley, unpublished). The resolution to this paradox (Tamarix beneﬁting from disturbance, but most abundant where disturbance is attenuated) involves differentiating the multiple roles that ﬂood disturbance plays. Flooding distributes seeds and creates suitable germination sites, but seedlings and young plants appear sensitive to physical disturbance; however, once established, trees are less prone to mortality and are unlikely to be removed except by extremely large events. We found that in an unregulated stream in the Sonoran Desert which experiences relatively frequent destructive ﬂoods (about one every three years), Tamarix ramosissima is present but extremely rare (and native riparian vegetation is well established: Dudley and D’Antonio, 1993). On the other hand, our study site in the Colorado Desert ﬂoods less frequently – approximately once every ten years – allowing Tamarix seedlings to mature and become resistant to subsequent ﬂooding (Dudley, unpublished). Similarly, invasion of Tamarix aphylla in the Finke River in Australia may be facilitated by a low ﬂood frequency; a storm event of sufﬁcient magnitude to promote the 1974 tamarisk establishment occurs with a return interval of 10–50 years (Grifﬁn et al., 1989). Since ﬂooding is not required for seed dispersal

432

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

and germination, the absence of ﬂooding in regulated watercourses allows eventual choking of the channel. It remains unclear whether Tamarix, once established in such a system, is a better competitor than native vegetation, or whether river regulation has resulted in conditions that are no longer suitable for native trees and Tamarix simply takes over as native trees decline (Howe and Knopf, 1991). In essence, for many systems it is not disturbance per se that fosters invasions by non-indigenous species, but the alteration of natural disturbance regimes to which most riparian-dependent native species are adapted. This relationship is more clearly demonstrated by aquatic fauna, particularly in streams in the western United States which naturally experience large seasonal variation in ﬂow regimes, most of which have been attenuated by water diversion and ﬂood control. Peter Moyle and his colleagues (Moyle and Williams, 1990; Brown and Moyle, 1991; Baltz and Moyle, 1993) have pointed out that the decline or loss of most native California stream ﬁshes has resulted, in part, from water developments which have altered natural hydrologies. These ﬁsh have been replaced by introduced centrarchids (sunﬁsh, freshwater bass), ictalurids (catﬁsh) and other habitat generalists from the eastern part of the continent and elsewhere (e.g., cyprinodontids – carp from Asia) which are preadapted to streams less liable to ﬂash ﬂoods, with greater sediment accumulation, higher turbidity, higher temperatures, and a more lentic nature. Possibly as important as attenuated high ﬂows in reducing the “natural” character of hydrological regimes are changes during periods of low ﬂow, especially postponement of drought and reduction in its severity, as a result of scheduled releases from reservoirs, again favoring ﬁsh associated with more stable hydrological regimes. The same types of ﬁsh have become common and are replacing native vertebrates elsewhere in regulated rivers around the world, including Africa (Gore et al., 1991), Australia (Sheldon and Walker, 1993), and the southwestern United States (Marsh and Minckley, 1982). Reduced ﬂooding has also promoted establishment of numerous species of crayﬁsh (Astacidae, Cambaridae) (Hobbs et al., 1989; Smith et al., 1996), possibly aquatic insects (Kido et al., 1993), freshwater mollusks (Corbicula ﬂuminea: Hornbach, 1992), and frogs [Rana catesbiana: Hayes and Jennings (1986); Kupferberg (1994); R. berlandieri: Clarkson and Rorabaugh (1989)], although it is rarely possible unambiguously to differentiate the importance of interacting

factors (habitat modiﬁcation, pollution, introduction of non-natives, human harvesting, etc.) in determining causes for declines in native biodiversity (Hayes and Jennings, 1986). Zebra mussels (Dreissenia spp.) are extremely costly invaders which may thrive in the less disturbance-prone, and often regulated, waterways of the eastern United States, as they are known to be intolerant of scouring in their native European waters (Strayer, 1991). Numerous cases of invasive plants increasing as a result of moderated ﬂow regimes certainly exist – for example, Peruvian peppertrees (Schinus molle) establishing where residential/agricultural runoff provides perennial moisture in an ephemeral stream channel (Howard and Minnich, 1989). An altered disturbance regime may not be necessary to promote invasions into riparian and riverine ecosystems, but we found few studies [e.g., colonization by beavers (Castor canadensis) of streams in Tierra del Fuego, Argentina (Lizarralde, 1993)] implying that disturbance is not involved. Bermuda grass (Cynodon dactylon) has occupied stream channels in the Sonoran Desert of the United States; in Arizona, it appears that livestock grazing of riparian vegetation in channels with natural ﬂooding regimes has been a disturbance factor (and possible dispersal factor) promoting its expansion (Dudley and Grimm, 1994). Naturally stable streams may be more sensitive to invasion without disturbance alteration; a spring stream in Texas was infested by both aquatic plants (Eichhornia crassipes, water hyacinth) and animals (e.g., Marisa cornuarietis, ramshorn snail) most typically associated with lentic waters (Bowles and Arsufﬁ, 1993). Wetland and lentic plant invasions In wetland and lentic habitats the role of disturbance in promoting or inhibiting invasions is more difﬁcult to assess, as the mechanisms facilitating invasions are rarely known or tested. Furthermore, the long history of land-use modiﬁcation and nutrient augmentation in surrounding catchments (Leach, 1995) makes it extremely difﬁcult to isolate and evaluate the inﬂuence that disturbance, as we have deﬁned it, may have on species invasions. Attached and ﬂoating plants have colonized temperate and tropical lakes, reservoirs, and waterways around the world (Pieterse and Murphy, 1990); these invasions have been amply discussed elsewhere, yet there is still little information on factors promoting invasion (Ashton and Mitchell, 1989). In wetlands, several studies have shown that anthropogenic disturbance promoted establishment of nonnative species in environments that do not experience

Location 1

Setting

Invader

Findings/disturbance

Study type 2

Reference

Pennsylvania (temperate forest)

marshes

Lythrum salicaria

establishes on open sites and ditches, expands into natural marsh; hydrologic alteration promotes exotics

Obs.

Wilcox (1989, 1995)

Florida (sub-tropical marsh/woodland)

many habitats – wet marsh to dry uplands, protected reserve subject to groundwater pumping

Melaleuca quinquenervia

invades “undisturbed” sawgrass prairie and cypress swamp and disturbed soils, but hydrological alteration may be involved

Obs.

Bodle (1994)

Florida

same

Melaleuca quinquenervia

seeds released and germinate in response to soil disruption and ﬁre; not required

Obs.

Myers (1983)

Britain

marsh/reservoir

Crassula helmsii

increases with ﬂuctuating water level and high alkalinity

Obs.

Dawson (1994)

Germany

temperate forest pond, coal mining

Crassula helmsii

establishes on disturbed soil of landﬁll pond

Obs.

Kuepper et al. (1996)

Eastern U.S.A. (temperate deciduous forest)

seasonal wetlands

many exotic species

increase with nutrient and metal inputs, hydrological changes less important

Cor.

Ehrenfeld and Schneider (1991)

Texas (hill country)

artesian spring pools

Eichhornia crassipes, Marisa cornuarietis, others

stable spring ﬂow (lacking disturbance) allows exotic plants and invertebrates to increase, compete with natives

Obs.

Bowles and Arsufﬁ (1993)

New Zealand

lakeshore, agriculture and urban

Ceratophyllum demersum, Egeria

both increase on lake margins, wave-swept rocks, associated with eutrophication

Obs.

Wells and Clayton (1991)

New South Wales, Australia

intermittent wetlands, agricultural

many species

fewest exotics in agricultural wetlands, moderate in disturbed roadsides, least in natural swamps, which also supported highest native species richness, resist invasion

Cor.

McIntyre et al. (1988)

1

Habitat types in parentheses after location. Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review. 2

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Table 17.7 Non-indigenous plants associated with lentic/wetland ecosystems, and relationships with disturbance regimes

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Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

substantial natural disturbance (Table 17.7). Decline in water quality was more critical than alteration in the water table in allowing invasion into Chamaecyparis wetlands (Ehrenfeld and Schneider, 1991). Two Australian studies showed that culturally disturbed wetlands (agricultural, cleared, and grazed) supported a higher proportion of introduced plant species than lessdisturbed nearby wetlands (35–39% vs. 19%) (McIntyre et al., 1988; Recher et al., 1993). In a restored wetland, once the native herbaceous vegetation had developed a clear zonation after two years, invasion by cat’s-tails, including Typha angustifolia, was reversed. Crassula helmsii from Oceania grows in moist sites and as an emergent plant in Great Britain (Dawson, 1994) and frog bit (Hydrocharis morsusranae) from Europe grows as a ﬂoating plant in Canada (Catling et al., 1988); both tolerate ﬂuctuating water levels better than local plants in natural and managed ponds and waterways. In these cases and in several others described previously, well-developed native vegetation appeared to provide “biotic resistance” to invasion, but this has not been the case with invasion by Melaleuca quinquenervia into Florida wetlands (Myers, 1983). Moist, rarely ﬂooded soils were optimal for Melaleuca germination, even under undisturbed conditions, but survival under less optimal conditions, and exceptional ability to resist ﬁre mortality and reproduce following ﬁre, along with other traits, allowed this extremely invasive tree to establish in many “natural” environments, even within the Everglades National Park. We mentioned earlier that purple loosestrife (Lythrum salicaria) occasionally invades vegetated wetland habitats, generally from more modiﬁed waterways but without obvious dependence upon disturbance (Thompson, 1991).

Summary of riparian and wetland invasions and disturbance It is clear that ﬂood scouring in riparian areas tends to favor the establishment of non-indigenous plant species, and that aquatic ecosystems throughout the world harbor large numbers of such species. In at least half of the studies we examined, riparian corridors also served as footholds from which non-indigenous species could enter adjacent, relatively undisturbed habitats.

While ﬂood disturbance promotes colonization by invasive species in riparian areas, examples from both plant and animal studies highlight the point that natural disturbances often favor native, rather than introduced, species. Thus, it is the disruption of natural disturbance regimes, rather than simply disturbance itself, which results in successful (and problematic) invasions in riparian ecosystems. In several of the cases we reviewed, native species exhibited greater population stability than did nonindigenous species, as would be expected if natural or relatively unmodiﬁed disturbance regimes promote the survival of native aquatic and riparian species adapted to regional environmental variability and associated disturbance regimes. Conversely, the further the environment is perturbed or altered from such regimes, the more likely that native species cannot be sustained, and species adapted to other environments will be able to establish abundant (but often ﬂuctuating) populations. Sometimes such population expansion will be detrimental to native biodiversity. Yet there are remarkably few instances in which experimental or other rigorous data allow one to differentiate between the hypothesis that non-indigenous species displace native species and the alternative hypothesis that they simply tolerate the altered conditions they encounter and utilize resources no longer suitable for indigenous species. This is, for example, the dilemma posed by the large-scale invasions of the ﬂoodplains of the western United States by Tamarix (saltcedar). While the studies covered nearly all suggest that some relationship exists between disturbance and invasion by non-indigenous species, it should be clear that this relationship is not simple. Some introduced plants (and animals) beneﬁt from reduced physical disturbance, others from the presence of regular disturbance to keep habitat space open for colonization. Also, some of these species may cause detrimental impacts to native biodiversity and ecosystem function, while others are simply ruderal species with no signiﬁcance to the protection of ecosystems. They are all of research interest and potential management concern, but to use the presence of exotic species as a de facto indicator of ecosystem degradation (cf. Keynhans, 1996) may not be justiﬁable without ranking the potential for ecological impact by individual species (Dudley and Collins, 1995).

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS SPECIES FEEDBACK TO ECOSYSTEM PROCESSES AND DISTURBANCE REGIMES

Background While many introduced species have little effect on the communities they enter (Simberloff, 1981), a particular subset of species alter ecosystem processes, including the disturbance regime. Here we deﬁne ecosystem processes as: whole-system ﬂuxes of energy, the amount or pathway of inputs, outputs, and the cycling of materials, and the ways that these processes vary in time (Vitousek, 1990; D’Antonio and Vitousek, 1992). Disturbance regimes are among the primary controls over the temporal variation of these processes, and their direct effects are well recognized; events such as hurricanes, landslides, and ﬁres cause discrete and punctuated changes in whole-ecosystem respiration and photosynthesis, energy and water balance, nutrient cycling, or nutrient outputs (Swanson et al., 1988; Lodge and McDowell, 1991; Ojima et al., 1994). In addition to opening up resource space for new or adjacent organisms, disturbance events can alter factors that control ecosystem processes; ﬂuxes of energy and matter may be changed, so that more of a resource is biologically available than before. For example, Steudler et al. (1991), in a tropical forest disturbed by Hurricane Hugo in 1989, found that the net rate of nitrogen mineralization in soil was 2.8 times the rate in undisturbed control plots. They ascribed this difference to a pulse of litter deposition and changes in microclimate. Chapin et al. (1996) suggested that ecosystem-level changes will be most likely when the invader possesses a trait that is discretely different than that of native species. When an invader differs from natives only in continuously distributed traits such as litter chemistry, growth rate, or size [termed “quantitative traits” by Chapin et al. (1996)] ecosystem processes will be less likely to be changed as a result of compensatory responses by functionally similar species. Alterations of disturbance regimes are perhaps the most dramatic examples where species traits can propagate through ecosystems to alter the growth and interactions of organisms. In this section, we will address two general questions about the effects of invaders on disturbance regimes: 1) What are the mechanisms through which invaders change disturbance regimes? and 2) Do invader-driven changes in disturbance regimes feed back to enhance or depress the success of the invader?

435

Many of the studies presented in the literature are anecdotal, but we include these to report the range of possible effects of invaders. We will focus more speciﬁcally on cases that are both widespread and that have some empirical basis. We also focus on situations where invaders have affected natural disturbance regimes. Although invaders have been shown to inﬂuence human-generated disturbances such as land use (Meyer, 1992), causality is much less clear. Survey summary Our survey uncovered 58 studies where invaders were connected to changes in disturbance regime by either correlation or observation (see Tables 17.8, 17.9, and 17.10). We separated invaders into those that differed qualitatively from natives in traits that affected disturbance regimes, and those that differed quantitatively (sensu Chapin et al., 1996). Invaders were determined to be qualitatively different if they had no functional analog in the known history of the invaded ecosystem. Functional analogy was determined as sharing a suite of similar traits that affected disturbance processes. For example, the feeding activities of feral pigs (Sus scrofa) have altered soil disturbance regimes in grasslands and forests in the Hawaiian islands (Spatz and Mueller-Dombois, 1975; Jacobi, 1981). There are no historical precedents for disturbances caused by large mammals in the Hawaiian islands. Although there were burrowing, ground-nesting birds (James, 1995) and burrowing land crabs (H. James, pers. commun.) that may have caused soil disturbance, neither of these now-extinct organisms was functionally analogous to pigs, and their distribution was likely more narrow. It is the combination of the foraging style, hooves, and body size of the pigs that create a type of soil disturbance unprecedented in the pre-human evolutionary history of Hawaiian ecosystems. We concluded that in 31 of the 58 cases surveyed, invaders differed qualitatively from natives in traits that affected disturbance processes (see Tables 17.8, 17.9 and 17.10). Of these 31 cases, 24 introduced a new disturbance regime that was without historical precedence. The remaining 7 cases resulted in the suppression or exacerbation of an indigenous disturbance regime. We designated invaders as differing quantitatively from natives when they shared a suite of similar traits, and differed only in the magnitude of the effect on disturbance processes. For example, Le Maitre et al. (1996) observed higher ﬁre frequencies in the South

436

Table 17.8 Effects of plant, pathogen, and invertebrate invaders on disturbance Invader

Growth form

Effect on disturbance regime 2

Invasion type 3 Study type 4

Reference

Hawaii (shrubland)

Heteropsylla cubana

psyllid

+ stand dieback

n.d.

Obs.

MacDonald and Cooper (1995)

Philippines (forest)

Ips interstitialis

beetle

+ stand dieback

Q,H

Obs.

Goldammer and Penaﬁel (1990)

New Zealand (woodland)

Leycesteria formosa, Rubus fruticosus

cane

− frost kill

Q,H

Exp.

Bannister (1990)

New Zealand (pasture)

Lumbricidae

earthworms

+ soil turnover

Q,H

Exp.

Syers et al. (1979)

Australia (dunes)

Ammophila arenaria, Cakile sp. Elymus farctus, Euphorbia paralias

grass

+ foredune stabilization

Q,N

Obs.

Heyligers (1988)

Australia (forest)

Sirex noctilio

wasp

+ stand dieback

n.d.

Obs.

Navaratnam and Catley (1986)

R´eunion Island (forest)

Lantana camara

shrub

+ bole break of trees in a L. camara thicket after cyclone

D,N

Obs.

MacDonald et al. (1991)

R´eunion Island (forest)

Rubus alceifolius

cane

+ tree fall of trees with R. alceifolius in canopy after cyclone

D,N

Obs.

MacDonald et al. (1991)

R´eunion Island (forest)

Casuarina equisetifolia, Cryptomeria japonica, Solanum mauritianum

tree

+ tree fall after cyclone

D,N

Obs.

MacDonald et al. (1991)

R´eunion Island (shrubland)

Psidium cattleianum

tree

+ stand defoliation

D,N

Obs.

MacDonald et al. (1991)

South Africa (fynbos)

Acacia cyclops, A. longifolia, A. saligna, Leptospermum laevigatum, Pinus pinaster

tree

− overland ﬂow

Q,H

Sim.

Van Wilgen and Richardson (1985)

South Africa (fynbos)

Acacia mearnsii

tree

+ ﬂood erosion of stream-banks

D,H

Ane.

MacDonald and Richardson (1986)

Central Africa (forest)

Maesopsis eminii

tree

+ soil erosion

n.d.

Ane.

Binggeli and Hamilton (1993)

Sub-saharan Africa (dunes)

Aristida pungens, Rhanterium suaveolens

herb

+ fore-dune stabilization

Q,N

Obs.

Bendali et al. (1990)

Tunisia (woodland)

Pinus halepensis

tree

− soil erosion

n.d.

Ane.

Wojterski (1990)

West coast U.S.A. (mud ﬂats) Carcinus maenas

crab

+ sediment mixing

n.d.

Ane.

Cohen et al. (1995)

Western U.S.A. (forest)

Cronartium ribicola

fungus

+ tree-gap formation

Q,H

Obs.

Byler et al., unpublished data

Central U.S.A. (prairie)

Lumbricidae

earthworms

+ soil turnover

Q,H

Exp.

James and Seastedt (1986); James 1991 continued on next page

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

Location 1

Location 1

Invader

Growth form

Eastern U.S.A. (forest)

Ceratocystis ulmi

fungus

+ tree-gap formation

Q,H

Obs

Manion (1981)

Eastern U.S.A. (forest)

Cryphoonectria [Endothiella] parasitica

fungus

+ tree-gap formation

Q,H

Sim.

Shugart and West (1977)

North America (prairie)

Bromus sp.

grass

+ soil erosion

Q,H

Ane.

Cheater (1992)

North America (forest)

Porthetria dispar

moth

+ stand defoliation

Q,H

Cor.

McManus and McIntyre (1981)

1

Effect on disturbance regime 2

Invasion type 3 Study type 4

Reference

Ecosystem type in parentheses following location. +, invader has a positive effect on some aspect of the disturbance regime; −, invader has a negative or a suppressant effect. 3 Invasion types: D, invaders that differ discreetly or qualitatively from natives; Q, invaders that differ continuously or quantitatively from natives; N, invaders introduce a new type of disturbance; H, invaders increase or decrease a disturbance that has been historically present in the system; n.d., not enough data to determine invasion type. 4 Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review; Sim = simulation. 2

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Table 17.8, continued

437

438

Table 17.9 The effects of vertebrate invaders on disturbance regimes Location 1

Invader 2

Effect on disturbance regime 3

Invasion type 4

Study type 5

Hawaii (rainforest)

pigs

+ litter and soil mixing and tree-fern decrease in native understory species, death increase in invaders

D,N

Gra.

Aplet et al. (1991); Vtorov (1993)

Hawaii (subalpine grassland) pigs

+ litter and soil mixing and tussock breaking

invasion of exotic grasses

D,N

Gra.

Spatz and Mueller-Dombois (1975); Jacobi (1981)

Hawaii (mesic forest)

pigs

+ litter and soil mixing

decrease in native understory species, increase in invaders

D,N

Gra.

Higashino and Stone, unpublished data

California (oak woodland)

pigs

+ litter and soil mixing

increased exotic component, shift from shrubs to grasses

D,N

Gra.

Peart and Patten (1992)

California (coastal prairie)

pigs

+ litter and soil mixing and plant burial

no change

D,N

Obs.

Kotanen (1995)

Northeastern U.S.A. (hardwood forest)

pigs

+ soil mixing and removal of vernal ﬂora

loss of understory species

Q,N

Obs.

Bratton (1974, 1975); Singer et al. (1984)

Southern U.S.A. (savanna)

pigs

+ litter and soil mixing, tussock breakage

no change

Q,H

Obs.

Baron (1982)

Sub-Antarctic island (shrubland)

pigs

+ litter and soil mixing, tussock breakage

loss of understory species

D,N

Obs.

Challies (1975)

Australia (wetlands)

pigs

+ litter and soil mixing, tussock breakage, stream-bank erosion

invasion of exotic grasses

D,N

Obs.

Bowman and Panton (1991)

Australia (rainforest)

pigs

+ litter and soil mixing, tussock breakage, stream-bank erosion

no change

D,N

Obs.

Russell-Smith and Bowman (1992)

Eastern U.S.A. (salt marsh)

horses

selective grazing

change in competitive dominance of grasses

n.d.

Obs.

Furbish and Albano (1994)

Northern Australia (rainforest)

water buffaloes

+ soil and root compaction, destruction of plants, erosion of stream banks

ecosystem degradation and loss of native species

D,N

Obs.

Russell-Smith and Bowman (1992)

Northern Australia (rainforest)

cattle

+ soil and root compaction, plant destruction

ecosystem degradation and loss of native species

D,N

Obs.

Russell-Smith and Bowman (1992)

Hawaii (shrubland)

goats

+ soil and root compaction, plant destruction

ecosystem degradation and loss of native species

D,N

Ane.

Stone (1985)

New Zealand (forest)

possums

+ forest defoliation

increased streamﬂow, siltation, and stand dieback

D,N

Obs.

Batcheler (1984)

Australia (woodland)

rabbits

+ grazing pressure

invasion of exotic herbs

Q,H

Gra.

Cochrane and McDonald (1966)

Australia (woodland)

rabbits

+ grazing pressure

overgrazing

Q,H

Ane.

Myers (1986)

Reference

Ecosystem type in parentheses following location. Pigs, Sus scrofa; horse, Equus caballus; water buffalo, Bubalus bubalis; cattle, Bos taurus; goat, Capra hircus; possum, Trichosurus vulpecula; rabbit, Oryctolagus cuniculus. 3 +, invader has a positive effect on some aspect of the disturbance regime; −, invader has a negative or a suppressant effect. 4 Invasion types: D, invaders that differ discreetly or qualitatively from natives; Q, invaders that differ continuously or quantitatively from natives; N, invaders introduce a new type of disturbance; H, invaders increase or decrease a disturbance that has been historically present in the system; n.d., not enough data to determine invasion type. 5 Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review; Sim., simulation. 2

Carla M. D’ANTONIO, Tom L. DUDLEY and Michelle MACK

1

Effect on community structure

Location 1

Invasive species

Growth form

Effect on disturbance regime

Invasion type 2

Study type 3

Reference

Australia (riparian)

Cenchrus ciliaris

grass

+ ﬁre frequency, areal extent

D,H

Obs.

Latz (1991)

Australia (woodland)

Ehrharta calycina

grass

+ ﬁre frequency

D,H

Obs.

Baird (1977)

Australia (woodland)

Eragrostis curvula, Ehrharta calycina

grass

+ ﬁre frequency

D,H

Exp.

Milberg and Lamont (1995)

Australia (mesic forest)

Melinis minutiﬂora

grass

+ ﬁre frequency

D,H

Obs.

Gill et al. (1981)

Australia (mesic forest)

Mimosa pigra

shrub

− ﬁre frequency

D,N

Obs.

Gill et al. (1990); Lonsdale and Miller (1993)

Australia (mesic forest)

Pennisetum polystachyon

grass

+ ﬁre frequency and intensity

D,H

Obs.

Gill et al. (1990)

California (chaparral)

Lolium perenne

grass

+ ﬁre frequency

D,H

Cor.

Zedler (1983)

Florida (prairies)

Schinus terebinthifolius

shrub

− ﬁre intensity

D,N

Exp.

Doren and Whiteaker (1990a,b)

Hawaii (subalpine woodland)

Anthoxanthum odoratum, Holcus lanatus grass

+ ﬁre intensity

Q,N

Cor.

Smith and Tunison (1992)

Hawaii (shrubland)

Andropogon virginicus, Cenchrus ciliaris, grass Hyparrhenia rufa, Pennisetum setaceum

+ ﬁre frequency

D,N

Cor.

Smith (1985); Hughes et al. (1991)

Hawaii (woodland)

Melinis minutiﬂora

grass

+ ﬁre frequency, areal extent

D,N

Cor.

Smith and Tunison (1992)

Hawaii (woodland)

Schizachyrium condensatum

grass

+ ﬁre size and frequency

D,N

Cor.

Smith (1985); Hughes et al. (1991)

Western U.S.A. (sagebrush steppe)

Bromus tectorum

grass

+ ﬁre size and frequency

D,N

Cor.

Stewart and Hull (1949); Whisenant (1990a)

Western U.S.A. (sagebrush steppe)

Taeniatherum caput-medusae

grass

+ ﬁre size and frequency

D,N

Obs.

Menke (1989); Young (1992)

grass, forbs

+ ﬁre size and frequency

D,N

Obs.

Brown and Minnich (1986)

Western U.S.A. (Mojave Desert) Brassica tournefortii, Bromus rubens, Schismus barbatus South Africa (fynbos)

Acacia saligna, Hakea sericea, Pinus spp. tree, shrub

+ ﬁre intensity

Q,H

Sim.

Van Wilgen and Richardson (1985)

South Africa (fynbos)

Hakea sericea, Pinus spp.

+ ﬁre intensity

Q,H

Obs.

Van Wilgen and Richardson (1985)

South Africa (fynbos)

Acacia cyclops, A. longifolia, A. saligna, tree, shrub Leptospermum laevigatum, Pinus pinaster

+ ﬁre frequency

Q,H

Obs.

Le Maitre et al. (1996)

South America (savanna)

Brachiaria spp., Hyparrhenia rufa, Melinis minutiﬂora, Panicum maximum.

+ ﬁre frequency

Q,H

Obs.

Blydenstein (1967); Medina (1987)

1

tree, shrub

grass

Ecosystem type in parentheses following location. Invasion types: D, invaders that differ discreetly or qualitatively from natives; Q, invaders that differ continuously or quantitatively from natives; N, invaders introduce a new type of disturbance; H, invaders increase or decrease a disturbance that has been historically present in the system; n.d., not enough data to determine invasion type. 3 Abbreviations: Ane., anecdotal (no data presented); Obs., observational (with data); Exp., experimental manipulation of vegetation conducted; Chr., chronosequence; Cor., correlational data; Gra., grazing exclosures; Rev., review; Sim., simulation. 2

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

Table 17.10 Invaders that enhance or depress ﬁre disturbance

439

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African fynbos associated with invasions by pines and acacias. Since the non-native trees differed from natives primarily in their ﬂammability, we designated them as being only quantitatively different from the natives. Out of the 58 cases that we surveyed, invaders differed quantitatively from natives in 21 cases. Only 3 of these 21 cases resulted in the introduction of a new type of disturbance; the remaining 18 of the 21 cases resulted in an exarcerbation or suppression of the indigenous disturbance regime. Finally, 6 of the total of 58 cases did not provide enough information for us to classify the invasion. In summary, we found that invaders that differed qualitatively from natives in a suite of traits were more likely to introduce a new disturbance regime to an invaded ecosystem. Although invaders that only differed quantitatively from natives did exhibit effects on disturbance regimes [contrary to the predictions of Chapin et al. (1996)], we found that these invaders were most likely to suppress or exacerbate an indigenous disturbance regime, resulting in more subtle changes in ecosystem processes and community structure than invaders which introduced qualitatively new disturbances. Mechanisms of alteration of disturbance regime Invasive species change disturbance regimes by affecting either the physical or the biological aspects of the disturbance. Invaders can introduce, enhance, or suppress events such as ﬁre and erosion, or may themselves be disturbance agents. Feeding activities may disturb soil or sediment, or may damage or kill and consume other members of the community. For the purposes of this review, we will not consider individual acts of predation a disturbance. Introduced species also change disturbance regimes by changing the aggregate biotic response of a community to a disturbance. Sousa (1984) characterized disturbance as a dynamic interaction between a damaging force and a biotic response. If invaders change the aggregate response or range of responses to a physical or biological force, the consequences of disturbance may change. For example, a 1989 cyclone on the island of R´eunion caused wide-spread uprooting and bole-breakage of non-native tree species, while native tree species remained relatively intact; hence the abundance of introduced trees changed the response and susceptibility of the community to disturbance (MacDonald et al., 1991). The only native trees to

show heavy damage either were within thickets of the introduced shrub Lantana camara, or had canes of the introduced blackberry Rubus alceifolius intertwined in their canopies. Invaders that affect physical forces of disturbance Invaders that enhance ﬁre: D’Antonio and Vitousek (1992) suggested that grass invasions cause potentially irreversible changes to ecosystems because of alterations in the ﬁre regime. The high surface-area/volume ratio of grass leaves and the typical accumulation of dead biomass increase the probability of ﬁre in invaded ecosystems. These grass/ﬁre feedbacks have been well documented in many parts of the world, particularly in the western United States (see Tables 17.5 and 17.10). They are responsible in western North America for the large-scale conversion of mixed shrub-steppe perennial vegetation to stands of introduced annual grasses (e.g., Klemmedson and Smith, 1955; Whisenant, 1990a). In Idaho rangeland, for example, ﬁre frequencies after invasion have been reported to have increased from once every 60–110 years to once every 3–5 years (Whisenant, 1990a). In Hawaii Volcanoes National Park, both ﬁre frequency and extent have increased since grass invasion (Smith and Tunison, 1992), and in the seasonal submontane zone ﬁres promote the spread and thickening of introduced grasses (Hughes et al., 1991; Tunison, unpublished data). These grasses create a more homogeneous canopy than in native forest, promoting higher wind speeds and greater rates of ﬁre spread (Freifelder et al., 1998). Positive effects of invaders on ﬁre regimes have also been suspected in some ecosystems where woody invaders are present but feedbacks are complex. For example, in South African fynbos, the Australian shrub Hakea sericea increases fuel loads and decreases average moisture content of live fuel, yet simulated rates of ﬁre spread under moderate weather conditions were lower than in uninvaded shrubland, owing to differences in fuel-packing density (Van Wilgen and Richardson, 1985). Sites invaded by Acacia saligna also had higher fuel loads but had higher fuel moisture than uninvaded sites. Model predictions under moderate weather were similar to predictions for fynbos invaded by H. sericea. However, under extreme weather conditions, ﬁre intensity in all invaded sites was much higher than in pristine fynbos (Van Wilgen and Richardson, 1985). The impacts of these intense ﬁres on further invasion have not been assessed, although

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

ﬁre is thought to stimulate germination and possibly facilitate the establishment of new stands of A. saligna (Holmes et al., 1987). Invaders that suppress ﬁre: We found only two studies that correlated presence of invaders with suppression of existing ﬁre regimes. The pepper tree Schinus terebinthifolius invades ﬁre-maintained coastal and mesic prairies and abandoned farmland in Florida. According to Doren and Whiteaker (1990b), mature stands can have densities up to 11 355 stems ha−1 with almost complete suppression of understory grasses. They found that the reduction in ﬁne fuel beneath S. terebinthifolius canopies depressed ﬁre intensity, area, and rate of spread. These effects could beneﬁt recruitment and survival of the invader, because young seedlings are susceptible to ground ﬁres and adults are killed by crown ﬁres (Doren and Whiteaker, 1990a). In Australia, Mimosa pigra, a low growing woody shrub from tropical America, invades mesic forests and ﬂood plains and depresses the growth of understory species (Braithwaite et al., 1989; Gill et al. 1990). Braithwaite et al. (1989) found that stands of M. pigra reduced light penetration to the understory by 50%. Gill et al. (1990) observed that M. pigra understories had lower stocks of ﬁne fuel compared to uninvaded areas. Mimosa pigra and other invasive shrubs probably decrease the frequency and intensity of ﬁres early in the dry season when these systems have historically burned (Gill et al., 1990). Wildﬁres have been observed to go out as they enter thickets of M. pigra (Lonsdale and Miller, 1993). Since M. pigra is ﬁre-tolerant [Lonsdale and Miller (1993) reported abundant regrowth following ﬁre], ﬁre suppression most likely does not directly beneﬁt the invader. Invaders that increase geomorphological disturbance: The magnitude of geomorphological disturbances can be increased by invaders. Acacia mearnsii, an Australian invader of South African ecosystems, increases stream-bank erosion as a result of being uprooted during periods of high ﬂow (MacDonald and Richardson, 1986). MacDonald and Cooper (1995) attributed this to the facts that A. mearnsii has a lower root-to-shoot ratio and is more shallowly rooted than the dominant native species, Prionium serratum. Following ﬂoods, river banks are rapidly colonized by introduced species (MacDonald and Richardson, 1986). Maesopsis eminii, a canopy tree that invades tropical mesic forests in Central Africa, has a thinner litter layer

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and increased rates of soil erosion beneath its canopy (Binggeli and Hamilton, 1993). The introduction of the Australian possum, Trichosurus vulpecula, to New Zealand led to stand-level dieback of hardwood forests and increases in stream ﬂow and sedimentation in lowland farmland (Batcheler, 1984). It is not clear in any of these cases that invader-induced changes in disturbance regime result in feedbacks that enhance the growth and reproduction of the invader. Invaders that decrease geomorphological disturbance: There are many examples of plants that bind and stabilize disturbed substrates. Grasses with extensive shallow root systems are used to stabilize landslides and eroded hillslopes, mining and construction scars, soil waste heaps, abandoned agricultural land, and sand dunes (van Kraagenoord and Hathaway, 1986). Clonal species that ramify by stolons or rhizomes are particularly successful at colonizing and stabilizing disturbed substrates, because they can grow from many points and keep up with rapid substrate movement and deposition. Some species introduced for sand or soil stabilization have become invaders of natural ecosystems. Ammophila arenaria, European beach grass or marram, alters dune formation patterns where it has invaded in North America (Dolan et al., 1973; Wiedemann and Pickart, 1996), New Zealand, and Australia (Heyligers, 1988). In southern Australia, A. arenaria is more efﬁcient at trapping sand than native beach grasses, and results in new foredune formation and larger dune size than in dune systems with native grasses (Heyligers, 1988). Substrate stabilization by invaders can lead to accelerated rates of succession. Cynodon dactylon (Bermuda grass) invades stream courses in Arizona and appears to affect community development by increasing substrate stability during ﬂoods (Dudley and Grimm, 1994). During ﬂoods, sites heavily dominated by C. dactylon retained more substrate, including basal fragments of native aquatic macrophytes. Post-ﬂood development of the aquatic macrophyte community proceeded more quickly in these sites than in those lacking C. dactylon. Overgrazed and eroded hillslopes in Mediterranean woodlands in Tunisia are invaded by Pinus halepensis, which stabilizes slopes and enhances establishment of native shrubs and trees (Wojterski, 1990). Invasion of South African fynbos by a suite of woody invaders has resulted in a 3- to 10-fold increase in aboveground biomass (Van Wilgen and Richardson, 1985). Increased transpiration and evaporation of intercepted rainfall

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has led to a 50% decrease in ﬂow levels of some stream catchments, which presumably leads to reduced sedimentation and erosion (Le Maitre et al., 1996). Invaders that affect the consequences of disturbance Some invaders introduce a qualitatively new response to a disturbance, often concurrent with the introduction of a new disturbance regime. This appears to be particularly true where introduced grasses have created a grass/ﬁre cycle. For example, Schizachyrium condensatum, a Central American grass that has invaded seasonal submontane Hawaiian woodlands, not only increases the frequency of ﬁres and the area burnt, but also regenerates from basal meristems following ﬁre (Smith and Tunison, 1992). Few of the dominant native species in these sites are able to do so, and sites rapidly become dominated by introduced grasses. Introduced species in grazing-land ecosystems also frequently introduce a qualitatively different suite of traits which interact with the biological disturbance of grazing. There are many examples where invaders with thorns, prickles, or chemical defenses have reduced the impact of grazers on a system and led to the abandonment of pastures by grazers and herdspeople (Auld and Tisdell, 1986; Jayanth and Ganga-Visalakshy, 1996; D. Glusenkamp, pers. commun.). Subsequent reductions in grazing pressure could eventually lead to increased representation of less grazing-tolerant species and/or woody species. Invaders that differ from natives only quantitatively can change community response to disturbance as well. Caldwell et al. (1981) in Utah compared the response to grazing of a native grass Agropyron spicatum and its rather similar introduced congener A. desertorum, and found that the canopy of the non-native species recovered more quickly from grazing and had a higher photosynthetic capacity in the new tissue than did the native species. Under heavy livestock grazing, this and other native grasses are being replaced over much of their range by more grazing-tolerant, introduced species (Caldwell et al., 1981). In South African fynbos, Australian species of Banksia and introduced species of Pinus respond to ﬁre by proliﬁc seedling establishment, leading to replacement of native shrubs (Richardson et al., 1990). Native shrubs in this area are also adapted to ﬁre and seedling establishment increases after ﬁre, but the introduced species are more effective in producing offspring. Invaders that are biotic disturbance agents Pigs: The European pig (Sus scrofa) is currently

found on all continents except Antarctica, and on many oceanic islands. Characteristic of most pig-invaded areas are altered soil-disturbance regimes. In Californian coastal prairies, pigs disturb 7.4% of the surface area each year; whereas native soil-disturbers (e.g., gophers) only accounted for disturbance of less than 1% of meadow areas (Kotanen, 1995). By grubbing for roots, underground stems and macroinvertebrates, pigs can dig up large areas, resulting in plant death (Bratton, 1974; Challies, 1975; Kotanen, 1995), root death (Singer et al., 1984), mixing of soil horizons (Singer et al., 1984; Vtorov, 1993; Kotanen, 1995), increased rates of nutrient mineralization (Singer et al., 1984; Vitousek, 1986) and decreased rates of nitrogen retention (Singer et al., 1984). Pig-grubbing may be associated with a depression of a number of soil microand macro-arthropods (Singer et al., 1984; Vtorov, 1993), but may also be associated with increased numbers of macroinvertebrates, particularly annelids (Anderson-Wong, 1994). Grubbing has been implicated in the spread of Phytophthora cinnamomi, a fungal root pathogen (Auld and Tisdell, 1986). The introduction of pigs into forested communities often results in the suppression, removal, or replacement of the herbaceous understory (Bratton, 1974, 1975; Challies, 1975; Aplet et al., 1991; Russell-Smith and Bowman, 1992). Removal of this functional group may alter nitrogen retention in the ecosystem (Singer et al., 1984; Aplet et al., 1991). In forests of gray beech (Fagus grandifolia) in North America, pig-grubbing can disturb up to 80% of the soil surface (Bratton, 1974) and leads to more rapid decomposition of soil organic matter (Singer et al., 1984). Singer et al. (1984) compared soil characteristics before pig invasion and 10 years later, and found a 65% decrease in soil organic matter, with signiﬁcant leaching of base cations and nitrogen from the soil proﬁle. They also found elevated nitrogen in soil water and stream water from catchments where pigs had rooted. Removal of preferred food species can feed back to affect pig populations. On the sub-Antarctic Auckland Island, pig populations increased explosively soon after introduction (~1807), then declined to a few individuals; by 1940 there was a stable population of less than 45 (Challies, 1975). Challies (1975) suggested that the population decline may have been associated with the near-extirpation of the preferred food – a group of endemic subantarctic large-leafed species. In Hawaiian montane rainforests, areas with pigs have higher densities of introduced, ﬂeshy-fruited species

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such as Passiﬂora mollisima, the banana poka, and Psidium cattleianum, the strawberry guava (Stone, 1985; Huenneke and Vitousek, 1990). These fruits are preferred by pigs (Smith, 1985; Stone and Loope, 1985). Aplet et al. (1991) in mesic Hawaiian forests found Myrica faya, an introduced nitrogen-ﬁxing tree, to be associated with pig disturbance. Myrica faya is in turn associated with increased densities of non-native earthworms, a food favored by pigs (Aplet, 1990). The introduction of pigs to grassland ecosystems has less obvious consequences. Although pig disturbance may lead to a change in species composition, it does not necessarily alter community structure (Spatz and Mueller-Dombois, 1975; Jacobi, 1981; Baron, 1982; Russell-Smith and Bowman, 1992). Pig rooting may lead to establishment of invaders, but these invaders are likely similar to displaced native species (Spatz and Mueller-Dombois, 1975; Jacobi, 1981). In Hawaiian montane grasslands, the Eurasian perennial grass Holcus lanatus established on pig disturbances more quickly than native grasses, leading to decreased densities of native perennial grasses (Spatz and MuellerDombois, 1975). In coastal Californian grassland, on the other hand, European annual grasses did not establish more readily on pig disturbances than natives (Kotanen, 1995). Although pigs have severely damaged lowland swamp and sedgeland communities in Northern Australia, communities remain dominated by native graminoid species (Russell-Smith and Bowman, 1992). Baron (1982) found that, although pig activity in dry grassland on Horn Island (Mississippi, U.S.A.) initially reduced perennial grass cover to 8%, 6 months later cover by native perennial grasses had increased by 163%. Most grassland systems have historically experienced soil disturbance regimes (Brown, 1989) which might explain why these communities appear to be more resilient to pig-generated disturbance. Other large feral animals: It is widely known that feral goats and sheep have severely altered island ecosystems (Coblentz, 1978; Van Vuren and Coblentz, 1987), where they are suspected of contributing to soil loss and soil disturbance. However, most studies of these animals have focused on their impacts on endangered native island species of plants, and not on their role in altering disturbance cycles or promoting further invasions. Mueller-Dombois and Spatz (1975) found that the removal of feral goats led to a decline in introduced annual grasses and an increase in native woody perennials and introduced perennial

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grasses. These particular introduced perennial grasses are known to be ﬁre-promoting (Smith and Tunison, 1992). Thus, elimination of feral goats may lead to increases in ﬁre frequency and further invasion by introduced grasses. In the Northern Territory of Australia, feral Eurasian water buffalo (Bubalus bubalis) have been found to increase invasion by annual plants because of their soil-disturbing activities (Cowie and Werner, 1993). Russell-Smith and Bowman (1992) reported that soil damage caused by water buffalo was present in all but one of 16 different vegetation types in this area, and 20% of over 1000 sites they surveyed were impacted by water buffalo. However, the longterm ecosystem consequences of disturbance by water buffalo and the plant species they appear to promote are unknown. Feral banteng (Bos javanicus) in Australia also reach very high densities, but do not appear to have deleterious effects on native forest ecosystems (Bowman and Panton, 1991). Small herbivores and detritivores: Small herbivores and detritivores have been implicated in changed disturbance regimes, including increased or decreased substrate mixing and stand-level dieback. In New Zealand pastures, European earthworms (Lumbricidae) signiﬁcantly increased decomposition and plant production (Syers et al., 1979). In contrast, European earthworms in the North American tallgrass prairie reduced soil turnover and nutrient mineralization compared to the native earthworm species (Megascolecidae), which they are replacing (James, 1991). In Australia, European rabbits (Oryctolagus cuniculus) have been implicated in severe range degradation and surface soil erosion (Myers, 1986; Fanning, 1994). These changes are unlikely to have positive feedback on rabbit populations. In many systems, introduced herbivorous insects can initiate a “boom-and-bust” cycle of stand-level dieback. North American bark beetles (Ips spp.), have caused increased tree death following ﬁres in tropical pine forests (Goldammer and Penaﬁel, 1990). Dieback of many North American hardwoods has been associated with outbreaks of the European gypsy moth Lymantria dispar (Houston, 1981). In both of these examples, stand regeneration occurs following collapse of pest populations. In Hawaii, Leucaena leucocephala, an invasive shrub on many oceanic islands, underwent severe dieback when the psyllid bug Heteropsylla cubana was introduced from Middle America (MacDonald and

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Cooper, 1995). Ten years after introduction of the psyllid bug, however, L. leucocephala seems to be recovering; many dense monospeciﬁc stands still exist (M. Mack, pers. observ.). Parasites and pathogens of biotic disturbance agents: Biotic disturbance agents themselves may be impacted or exterminated by introduced parasites or pathogens. In Africa, the rinderpest virus decimated natural populations of herbivores across the continent, presumably decreasing natural grazing disturbance (Dobson, 1995). There are many examples in the agricultural literature where outbreaks of herbivorous insects have been decimated by parasites or pathogens introduced as part of a biological control program (Allen and Rada, 1984). In natural systems, effects are less clear, but several examples of clear causation exist: Epizootics of Entomophaga maimaiga, a fungus from Japan, have been observed in populations of the highly destructive gypsy moth in North America (United States Department of Agriculture, 1995). In Canada, outbreaks of the forest-defoliating winter moth Operophtera brumata have been controlled by a suite of parasites (Reardon, 1981). In all of these cases, an introduced pathogen caused reduction in the population of the biotic disturbance agent, thereby decreasing its effect on the system. It is clear that the effect of the pathogen does not feed back to increase its own success – at least, not immediately.

CONCLUSIONS

Introduced species are common in disturbance corridors whether these are natural, as in the case of riparian corridors, or anthropogenic (e.g., roads, trails, pipelines). These corridors frequently (in approximately 50% of cases) act as foci from which nonindigenous species enter adjacent undisturbed habitats. Whether or not invasion of adjacent habitats occurs is likely a function of the habitat type (Zink et al., 1995), the biology of the invader, and the length of time since invaders entered the area. In two studies claiming that invaders generally did not move away from disturbance corridors, the authors (Wein et al., 1992; Parker et al., 1993) admitted that at least one of the species studied is beginning to appear away from disturbed edges. Numerous authors have suggested that lag periods between the ﬁrst appearance of a species (along a disturbance corridor, for instance),

and widespread invasion are common (e.g., Hobbs and Humphries, 1995). Thus, managers should not assume that introduced species will stay conﬁned to trails and disturbed corridors and edges, but should minimize production of corridors and reduce the population size of invaders along those corridors. Catastrophic disturbances, whether natural (severe ﬂoods, hurricanes) or anthropogenic (agricultural transformation, logging), tend to favor the establishment of entire communities of introduced species. In many of the studies we surveyed, non-indigenous species were eventually successional to native species, particularly in the case of logging disturbance. However, in cases of severe soil disruption (e.g., abandoned agricultural ﬁelds in the desert), introduced species appear to retain their dominance even after many years. In terrestrial ecosystems, as long-term studies accumulate, we predict that sites where the soil proﬁle has been severely disturbed, as in agriculture, will be those where introduced species persist, and that less disturbed sites where the soil proﬁle has not been greatly altered, such as logged forests, will recover with native species more quickly. Rates of recovery will likely depend on habitat productivity. Indeed, our review suggests that abandoned arid-land agricultural sites are dominated by introduced species for a very long time. Introduced species can invade many habitats without apparent disturbance. This is particularly true in deserts, islands, and coastal habitats. Managers of reserves in such areas should be particularly sensitive to the presence of disturbance corridors and other avenues through which propagules of introduced species might arrive in their ecosystems. In many terrestrial systems, the intensity of initial invasions might be low in the absence of disturbance but they are then enhanced by later disturbances. This is particularly true for grasses that are favored by ﬁre. Natural disturbances are often instrumental in promoting invasion. Small-scale natural disturbances in terrestrial ecosystems, and ﬁre, were found to promote invasion in more than 75% of the cases examined. The persistence of invaders was habitat-dependent. Natural disturbances in riparian systems virtually always promoted invasion if the propagules of invaders were available. Because of the importance of riparian corridors for native plant and animal diversity, understanding the impacts and persistence of invaders in these habitats is crucial. In the case of some of the largest-scale invasions that have been documented, the relative roles of particular

DISTURBANCE AND BIOLOGICAL INVASIONS: DIRECT EFFECTS AND FEEDBACKS

types of disturbance and stress are unclear, and multiple causation is likely. For example, in the almost complete transformation of California grasslands to domination by Eurasian species, drought, ﬁre, and livestock grazing probably acted together to promote invasion and transformation of the landscape. In the almost complete transformation of ﬂoodplain forests in the Colorado River basin in the United States over the past 50 years, it is the combination of decreased water table, increased soil salinity, and reduced vigor of native species as a result of alterations in natural disturbance regimes, that has led to massive invasion by Tamarix. Introduced species themselves often cause a change in the disturbance regime. We found that non-indigenous species differing from the native species in the invaded sites in a suite of traits, at least some which are qualitative, tend to introduce novel disturbances into communities. Rather than introducing novel disturbances, introduced species that differ mainly quantitatively from natives in the invaded sites, tend only to modify the historical disturbance regime for these sites. Do invaders cause disturbance feedbacks that promote more invasion? We were not able to assess how many invaders show this pattern, because few people gather quantitative information on the subject. However, in the case of ﬁre-promoting grasses, it is clear that introduced species can create feedbacks that will perpetuate their own kind. Feral animals that cause soil disturbance often promote introduced plants, but there are few data to show how this in turn affects the dynamics of the feral animals. Introduced species can affect not only the size, frequency, intensity, or type of disturbance by being agents of disturbance themselves or fueling disturbances, but they can alter the response of the community to disturbance. While studies documenting this phenomenon are few, we believe that it will become an increasingly important form of feedback as introduced species continue their expansion into more or less natural ecosystems. The search for generalities and predictive power in the area of biological invasions is frustrated by a lack of quantitative studies, particularly ones focusing on the mechanisms through which communities might resist invasion (but see Hobbs and Atkins 1988, 1990; Hobbs et al., 1988; Bossard, 1991, 1993; D’Antonio, 1993; D’Antonio et al., 1993; Burke and Grime, 1996). Without knowledge of organizing forces within communities, and thus of why communities are or

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are not resistant to invasion, it is difﬁcult to predict the impacts of changes in the disturbance regime on invasion. ACKNOWLEDGMENTS

Heather Schwartz typed the bibliography to this paper, a monumental effort almost deserving of co-authorship. Financial Support was provided by NSF BSR 9119618 and WRC (State of California) W-867. We would also like to thank Eric Berlow, Ingrid Parker, Lawrence Walker, Steve Whisenant and Richard Hobbs for comments on an earlier draft and the D’Antonio lab group for feedback on ideas and references. REFERENCES Allen, G. and Rada, A., 1984. The role of biological control in pest management. In: G. Allen and A. Rada (Editors), The Role of Biological Control in Pest Management. Conference Proceedings, Santiago, Chile, pp. 1–8. Anable, M.E., McClaran, M.P. and Ruyle, G.B., 1992. Spread of introduced Lehmann lovegrass Eragrostis lehmanniana Nees. in Southern Arizona, USA. Biol. Conserv., 61: 181–188. Anderson, J.E. and Marlette, G.M., 1986. Probabilities of seedling recruitment and the stability of crested wheatgrass stands. In: K.L. Johnson (Editor), Crested Wheatgrass: Its values, problems, and myths. Conference Proceedings, Utah State University, pp. 97– 105. Anderson-Wong, P., 1994. A relationship between earthworms and feral pig activity in a Hawaiian rain forest. In: Hawaii Conservation Conference. Conference Proceedings, Honolulu, Hawaii, p. 6. Aplet, G.H., 1990. Alteration of earthworm community biomass by the alien Myrica faya in Hawai’i. Oecologia, 82: 414–416. Aplet, G.H., Anderson, S.J. and Stone, C.P., 1991. Association between feral pig disturbance and the composition of some alien plant assemblages in Hawaii Volcanoes National Park. Vegetatio, 95: 55–62. Ashton, P.J. and Mitchell, P.S., 1989. Aquatic plants: patterns and modes of invasion, attributes of invading species and assessment of control programs. In: J.A. Drake, H.A. Mooney, F. di Castri, R.H. Groves, F.J. Kruger, M. Rejm´anek and M. Williamson (Editors), Biological Invasions: A Global Perspective. Wiley, Chichester, pp. 111–147. Auld, B.A. and Tisdell, C.A., 1986. Impact assessment of biological invasions. In: R.H. Groves and J.J. Burdon (Editors), Ecology of Biological Invasions. Cambridge University Press, Cambridge, pp. 79–88. Baird, A.M., 1977. Regeneration after ﬁre in King’s Park, Western Australia. J. R. Soc. West. Aust., 60: 1–22. Baltz, D.M. and Moyle, P.B., 1993. Invasion resistance to introduced species by a native assemblage of California stream ﬁshes. Ecol. Appl., 3: 246–255. Bannister, P., 1990. Frost resistance of leaves of some plants growing in Dunedin, New Zealand, in winter 1987 and late autumn 1989. N. Z. J. Bot., 28: 359–362.

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Melick, D.R. and Ashton, D.H., 1991. The effect of natural disturbance in warm temperate rainforest in southeast Australia. Aust. J. Bot., 39: 1–30. Menke, J.W., 1989. Management controls on productivity. In: L.F. Huenneke and H.A. Mooney (Editors), Grassland Structure and Function: California Annual Grassland. Kluwer, Dordrecht, pp. 173–199. Meyer, W.B., 1992. Human population growth and global land-use cover change. Annu. Rev. Ecol. Syst., 23: 63–87. Milberg, P. and Lamont, B.B., 1995. Fire enhances weed invasion of roadside vegetation in southwestern Australia. Biol. Conserv., 73: 45–49. Milchunas, D.G., Sala, O.E. and Lauenroth, W.K., 1988. A generalized model of the effects of grazing by large herbivores on grassland community structure. Am. Nat., 132: 87–106. Milchunas, D.G., Lauenroth, W.K., Chapman, P.L. and Kazempour, M.K., 1990. Community attributes along a perturbation gradient in a shortgrass steppe. J. Vegetation Sci., 1: 375–384. Miller, R.F., Svejcar, T. and West, N.E., 1994. Implications of livestock grazing in the intermountain sagebrush region: plant composition. In: M. Vavra, W.A. Laycock and R.D. Pieper (Editors), Ecological Implications of Livestock Herbivory in the West. Society for Range Management, Denver, Colorado, pp. 101– 146. Moreno, J.M. and Oechel, W.C., 1991. Fire intensity effects on germination of shrubs and herbs in southern California chaparral. Ecology, 72: 1993–2004. Moyle, P.B. and Williams, J.E., 1990. Biodiversity loss in the temperate zone: decline of the native ﬁsh fauna of California. Conserv. Biol., 4: 275–284. Mueller-Dombois, D. and Spatz, G., 1975. The inﬂuence of feral goats on the lowland vegetation of Hawaii Volcanoes National Park. Phytocoenologia, 3: 1–29. Myers, J.H. and Berube, D.E., 1983. Diffuse knapweed invasion into rangeland in the dry interior of British Columbia. Can. J. Plant Sci., 63: 981–987. Myers, K., 1986. Introduced vertebrates in Australia with emphasis on the mammals. In: R.H. Groves and J.J. Burdon (Editors), Ecology of Biological Invasions. Cambridge University Press, Cambridge, pp. 120–136. Myers, R.L., 1983. Site susceptibility to invasion by the exotic tree Melaleuca quinquenervia in southern Florida. J. Appl. Ecol., 20: 645–658. Navaratnam, S.J. and Catley, A., 1986. Quarantine measures to exclude plant pests. In: R.H. Groves and J.J. Burdon (Editors), Ecology of Biological Invasions. Cambridge University Press, Cambridge, pp. 106–112. Niswander, S.F. and Mitsch, W.J., 1995. Functional analysis of a 2year-old in-stream wetland: hydrology, phosphorus retention, and vegetation survival and growth. Wetlands, 15: 212–225. Ojima, D.S., Schimel, D.S., Parton, W.J. and Owensby, C.E., 1994. Long- and short-term effects of ﬁre on nitrogen cycling in tallgrass prairie. Biogeochemistry, 24: 67–84. Olson, S. and James, H., 1982. The role of Polynesians in the extinction of the avifauna of the Hawaiiana islands. In: P.S. Martin and R.G. Klein (Editors), Quaternary Extinctions: A Prehistoric Revolution. University of Arizona Press, Tucson, pp. 768–780. Palmer, J.P., 1994. Fallopia japonica (Japanese knotweed) in Wales. In: L.C. de Waal, L.E. Child, P.M. Wade and J.H. Brock (Editors),

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Chapter 18

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE Dan BINKLEY

INTRODUCTION

Temperate and boreal forests cover ~2.4×109 hectares, almost 20% of the terrestrial surface of the earth (FAO, 1995). The separation of temperate and boreal forests is unclear, as statistics are commonly compiled by country rather than ecosystem type. The majority of temperate forest areas fall within Canada, China, Russia, and the United States. Temperate forests typically occur where annual temperatures average between 6 and 18ºC, and precipitation averages more than 500 mm yr−1 (Archibold, 1995). Cooler conditions characterize boreal forests, and warmer or drier conditions favor grasslands or shrublands. Within this overall range, the occurrence and character of temperate forests varies with seasonality of precipitation, and the frequency and intensity of disturbances. A full account of these forests will be found in other volumes of this series (Ovington, 1983; R¨ohrig and Ulrich, 1991). The aboveground biomass of mature temperate forests ranges from about 10 kg m−2 for poor-quality sites (with poor supplies of water or nutrients, or severe environmental conditions), to more than 100 kg m−2 for old-growth conifer forests (>300 kg m−2 for Sequoia forests) (Cannell, 1982). The rate of aboveground net primary production typically ranges from 0.4 to 2.5 kg m−2 yr−1 . The time between stand-replacing disturbances is commonly on the order of centuries, with typical ages for mature trees of 50 to >2000 years. The current extent of temperate forests in the northern hemisphere is probably about 20% less than in preagricultural times (Williams, 1994), and most of the remaining forests have been moderately to heavily affected by human activities. These human-related changes include widespread alteration of ﬁre regimes, conversion of old-growth forests to younger forests,

conversion of diverse forests into intensively managed plantations, and introduction of exotic species. The patterns of forest disturbances have been substantially altered, even where human inﬂuences have been relatively slight. The basic features of natural disturbances in temperate forests are brieﬂy mentioned in this chapter, as many additional details are available in other chapters. The focus then shifts to human-related disturbances (particularly forest harvesting), highlighting the similarities and differences between these and more natural disturbances. Ideally, these subjects might be woven together with a theme of natural vs. humanrelated disturbances, but the wide range of scales and impacts in both categories minimize any useful generalizations.

PATTERNS AND PROCESSES OF NATURAL DISTURBANCES

Foresters and ecologists have always been aware that disturbances can shape the development of forests, but a deeper appreciation of the pervasive importance of disturbances has developed only in the past few decades. Across the United States, for example, the annual rate of tree mortality averages about 18% of the rate of forest growth (on a volume basis: United States Department of Agriculture Forest Service, 1981). This mortality is driven by many processes, and patterns range from death of single trees to widespread disturbances across landscapes. Old-growth forests commonly occupied much of the temperate-forest region before widespread industrialization, and old-growth forests typically experienced

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high rates of mortality. For example, gaps in the canopies of old-growth forests of amabilis ﬁr (Abies amabilis), western hemlock (Tsuga heterophylla), and mountain hemlock (Tsuga mertensiana) in coastal British Columbia account for 18% of the ground area, and another 60% of the forest area is close to a gap – at a distance less than the canopy radius of a single tree (Lertzman and Krebs, 1991). In addition, in North America before European settlement, natural disturbance regimes led to substantial areas of younger forests within landscapes dominated by older forests (Whitney, 1994). White and Pickett (1985) developed some terms for describing the major features of this broad array of disturbances, including distribution (in space, and along environmental gradients), frequency or return interval, rotation period (time needed for entire area to be disturbed at least once), predictability, size, intensity, and severity. They also noted that occurrence of one disturbance, such as a drought, could increase the likelihood of another (such as an insect outbreak). These features can be grouped broadly in terms of major disturbances (called stand-replacing disturbances, where most of the dominant plants are removed) and minor disturbances (which leave the majority of dominant plants alive: Oliver and Larson, 1996).

WIND

Most forests of the world experience wind events that damage trees (see Webb, Chapter 7, this volume, for more detail). These wind disturbances span scales from removal of branches on some trees by moderate winds (often associated with heavy loading of branches with snow or ice) to removing most of the dominant trees across landscapes by hurricanes. The risk of wind damage to an individual tree (or stand) tends to increase with age, as a result of increasing size and resistance to the wind (Oliver and Larson, 1996). The resistance of a tree to wind damage depends in part on the wind regime experienced during the development of the tree; trees exposed to strong winds develop stem forms (taper) that accommodate greater stress from swaying canopies. Changes in stand structure, such as creation of gaps, may increase the wind velocities experienced by individual trees, leading to either breakage of stems or uprooting of the entire tree. Uprooted trees create pit and mound

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microtopography (covering up to 30% of a forest area), mixing soil horizons and providing raised microsites that may favor tree regeneration (Pritchett and Fisher, 1987). Many case studies document the widespread impacts of winds. Whitney (1994) tabulated estimates of major wind damage for forests in the eastern United States before European settlement. The return interval (reciprocal of the annual proportion of an area devastated by windstorms) averaged between 500 and 1300 years for the northeastern and midwestern U.S.A. This range overlaps with the return interval for ﬁre, which ranged from 130 to 14 000 years, aside from an interval of 80 years for jack pine (Pinus banksiana). Franklin (1988) cited unpublished data from across the Paciﬁc Northwest of the United States on mortality of trees in old-growth stands. Winds were responsible for 80% of mortality in coastal forests of Sitka spruce (Picea sitchensis) and western hemlock (Tsuga heterophylla), compared with 40% of the mortality forests of Douglas-ﬁr (Pseudotsuga menziesii) and western hemlock in the Cascade Mountains, and 20% of mortality in drier, less dense forests of ponderosa pine (Pinus ponderosa) to the east. Large windstorms can create mosaics of disturbance across large landscapes. For example, a windstorm in 1921 damaged forests across 25 000 ha of the Olympic Peninsula of Washington (U.S.A.), with severity of damage ranging from slight to almost complete removal of dominant trees on different portions of the landscape (Oliver and Larson, 1996). This storm had two major effects: accentuating landscape heterogeneity by creating a mosaic pattern of old forests mixed with regenerating forests, and insuring some continued homogeneity into the future by synchronizing the regeneration time across large portions of the landscape. The condition of stands may play a large role in determining the effects of high winds. Diseased or damaged trees tend to be less wind-ﬁrm, and some species are more prone to wind damage than others. For example, Hurricane Hugo in 1989 toppled mature loblolly pine (Pinus taeda) more readily than similarsized longleaf pine (P. palustris) (Gresham et al., 1991). Wind disturbance may lead to a synergistic interaction (sensu White and Pickett, 1985) with other disturbances, particularly with ﬁre and insect outbreaks. Accumulations of large quantities of dead wood can greatly increase the risk of subsequent ﬁres. A strong windstorm in the Rocky Mountains of

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE

Colorado (U.S.A.) in 1939 is thought to have provided the conditions necessary for an outbreak of spruce bark beetles (Dendroctonus ruﬁpennis), and the associated blue-stain fungus (Leptographium engelmannii) that devastated almost 300 000 ha over the following decade (Veblen et al. 1991; see section “Insects and Diseases” below). Nutrient supplies may increase following major wind disturbances, as a result of decreased uptake by plants, perhaps lower supplies of labile carbon for microbial immobilization, and higher rates of mineralization of organic matter. I know of only one study that examined net nitrogen mineralization in a wind-generated age sequence of temperate forests. The high-mortality zones of ﬁr waves1 of both Fraser ﬁr (Abies fraseri) in North Carolina (U.S.A.) and balsam ﬁr (Abies balsamea) in New York (U.S.A.) showed rates of net nitrogen mineralization that were similar to those in adjacent mature stands (Sasser and Binkley, 1989). More information would be needed to develop a general picture. Wind disturbances generally leave some vegetation relatively intact, and some root systems or broken stems may sprout. Disturbances to the soil are generally slight, except where roots have been upturned. Revegetation is relatively rapid, and nutrient losses are probably minimal.

INSECTS AND DISEASES

Insect populations are generally present in forests at low-to-moderate levels (Schowalter, Chapter 9, this volume). Low populations may have substantial effects on trees, and landscape-scale disturbances typically result when populations increase by orders of magnitude above background levels. These major disturbances typically have a synergistic component – population irruptions often depend in part on weather patterns and on susceptibility of host trees (which may in turn relate to weather patterns or other stresses). Forest diseases, such as root-rot fungi, generally affect smaller portions of a landscape than do major insect outbreaks, but also show large synergistic effects. In addition, synergies are common between insects and diseases. Bark beetle infestations are the indirect cause of tree death in many instances; the direct cause is the proliferation of blue 1

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stain fungal hyphae developing from spores transported by the beetles. Many temperate regions are experiencing unusually severe outbreaks of insects and disease pathogens as a result of human-related introductions to new forest areas. Exotic bark beetles (Scolytus multistriatus) introduced the exotic Dutch elm disease (Ceratocystis ulmi) which decimated American elm (Ulmus americana) over much of its native range in North America. Similarly, the chestnut blight (Endothiella parasitica) from Asia was introduced to forests of the eastern United States early in this century, and within a few decades American chestnut (Castanea dentata) was eliminated as a dominant component of these forests. Gypsy moths (Porthetria dispar) have greatly inﬂuenced rates of defoliation (particularly for oaks – Quercus spp.) in some forests of the same region, but with much lower rates of mortality. Nutrient losses in stream water may be increased after disturbance by insects or diseases, but the increases are relatively small. For example, Swank (1988) noted that defoliation of a hardwood forest (up to one-third of the canopy removed through the growing season for several years) raised stream-water concentrations of nitrate nitrogen from background levels of about 10 g °−1 to about 40 g °−1 . Although the response of stream-water nutrient concentrations was dramatic, the increase in loss of nitrogen was quite small (<0.5 kg nitrogen ha−1 yr−1 ). More severe disturbances, such as near-total removal of all tree cover by bark beetles in stands of lodgepole pine (Pinus contorta) may lead to greater nitrogen losses (perhaps 1–3 kg N ha−1 yr−1 : Knight et al., 1991); but, even so, these losses are relatively minor. Disturbances generated by insects and diseases typically leave portions of the vegetation relatively undamaged; responses by surviving trees and understory plants often lead to rapid recovery of vegetation on a site. Soils are generally not disturbed, except where root-rot fungi have caused trees to blow down more easily.

FIRE

Fire is probably the most pervasive major disturbance of temperate forests (for general summaries, see Wright

“Fir waves” are a wind-induced pattern of mortality and regeneration that “moves” across the high-elevation landscape, both in North America and in Japan.

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and Bailey, 1982; Agee, 1993; Pyne et al., 1996). The historic role of ﬁre and humans in ecosystems is blended. In many temperate-forest areas, indigenous people were a major source of ﬁre ignition (e.g., Barrett and Arno, 1982; Pyne, 1993). Pyne et al. (1996) suggested that Europe was “all but uninhabitable” without the use of ﬁre. The role of humans in other regions remains unclear (Agee, 1993; Whitney, 1994). A wide variety of studies have estimated the occurrence of ﬁres in North America prior to European settlement, and these rates may represent common ranges for other temperate forests as well (Table 18.1). Table 18.1 Estimated ﬁre-return interval in temperate forests of North America 1

Fig. 18.1. Fuel consumption (a measure of ﬁre intensity) generally increases with the ﬁre return interval in North American temperate forests (based on Olson, 1981; Christensen, 1987).

Major forest type (location) 2

Return interval (yr) 2,3

Pine forests (southeastern USA)

2–25

Deciduous forests (northeastern USA/Canada)

10–350+

Spruce/hemlock (eastern Canada)

200–1000

Pine/mixed hardwood (mid-continent)

15–300

Mixed conifer (Rocky Mountains)

10–300

Ponderosa pine (western USA)

2–25

Sequoia (California)

3–40

continental-scale return interval of 125 to 250 years. Fire suppression has substantially altered the structure and species composition of many temperate forests; the area burned each year has decreased, but the number of ignitions and the consumption of biomass per unit area have increased (Kauffman et al., 1993). In Canada, ﬁres consume about 2.5×106 ha of forests annually (Stocks and Simard, 1993), the majority of ﬁres being in boreal forests (see Oesterheld et al., Chapter 11, this volume) rather than Canada’s southern temperate forests. In pre-European times, the ﬁre return intervals in Canadian forests probably averaged from 50 to 200 years, whereas current rates of burning correspond with average return intervals of 400 to >2000 years (Pyne et al., 1996). Information is spottier for other temperate regions. Heavily settled parts of Europe have very low ﬁre frequencies; ﬁre regimes in Russia and other countries of the former Soviet Union vary substantially among regions (with changes in settlement patterns and ﬁre-control policies), and the relatively low frequency of ﬁres in temperate forests in China was overshadowed by the 3×106 ha of forest consumed by the Black Dragon ﬁre in 1987 (Pyne et al., 1996).

Conifer forests (northwestern USA) 50–750 1

Based on Heinselman (1981); Martin (1982); Agee (1993); Clark and Robinson (1993); Oliver and Larson (1996). 2 This broad range of forest types and ﬁre regimes is probably representative of similar forest types in other temperate regions. 3 Note that any particular hectare may burn more or less often than the long-term average across the landscape.

Forests with short return intervals for ﬁre commonly experienced surface ﬁres that did not kill the dominant vegetation, but probably shaped the composition of the understory and the long-term dynamics of the forests. Return intervals of a century or more typically led to high-intensity, stand-replacing ﬁres (Fig. 18.1). At a continental scale, the 48 contiguous United States contain about 250×106 ha of temperate forests, and about 15×106 ha of forests burned each year in the ﬁrst third of the 20th Century (MacCleery, 1992), for an average return period on the continental scale of about 15–20 years. This ﬁre pattern was unusual, reﬂecting a substantial role of humans in igniting ﬁre, but little effort at ﬁre suppression. Since the 1960s, only about 1–2×106 ha have burned annually, for a

Surface ﬁres Low-intensity, frequent surface ﬁres were characteristic minor disturbances of pine forests (Pinus palustris, P. taeda) in the southeastern part of North America, and in western forests of ponderosa pine (P. ponderosa). As a result of these ﬁres, accumulations of material (and nitrogen) in the forest ﬂoor (litter layer, O horizon) were low, there were lower densities of trees per

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE

hectare (particularly for ponderosa pine forests), and high production and diversity of herbaceous species. The biomass of pine forests in the southeastern United States is typically 15 000–30 000 kg ha−1 , with 50– 150 kg nitrogen ha−1 ; repeated burning reduces forestﬂoor biomass to 10 000–15 000 kg ha−1 and nitrogen content to 20–35 kg ha−1 (Binkley et al., 1992). The loss of nitrogen to the atmosphere during surface ﬁres averages about 5 kg nitrogen per megagram of fuel consumed (Fig. 18.2). Ponderosa pine forests with natural ﬁre regimes in

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the southwestern United States have a tree density of about 30 to 150 trees ha−1 , compared with 700 to >2000 trees ha−1 following decades of ﬁre suppression (Covington and Moore, 1994). Low-density stands have major understory communities (containing 50 to 200 species in a hectare) that contribute substantially to ecosystem productivity, whereas in high-density, unburned stands of ponderosa pine productivity is concentrated primarily in the overstory trees (Covington, 1994). Scorched pines may be more susceptible to attacks by bark beetles (Agee, 1993), another example of synergy between disturbance factors. Stand-replacing ﬁres

Fig. 18.2. Nitrogen loss is proportional to fuel consumption (about 5 kg N lost per megagram of fuel consumed) for both surface ﬁres (upper) and slash ﬁres (lower). Pita1 (loblolly pine) from Kodama and Van Lear (1980); Pita2 from Richter et al. (1982); Pita3 from Schoch and Binkley (1986); Pipo1 (ponderosa pine) from Nissley et al. (1980); Pipo2 from Covington and Sackett (1984); Laoc (larch (Larix occidentalis)Douglas-ﬁr stand) from Jurgensen et al. (1981); data for slash ﬁres from Little and Ohmann (1988).

Stand-replacing ﬁres vary widely, in the amounts of fuel consumed, temperature regimes (aboveground and belowground), and subsequent environmental effects. Once forests have accumulated enough fuel, the occurrence, intensity, and size of stand-replacing ﬁres depend greatly on weather conditions. The scale of stand-replacing ﬁres ranges from several hectares up to hundreds of thousands (or even millions) of hectares. Within a single burned area, variations in fuels, topography, and winds typically lead to a variety of ﬁre intensities and subsequent implications for ecosystem recovery. These ranges in scale are illustrated well by the work of W. Romme and his colleagues in Yellowstone National Park (Figs. 18.3, 18.4). The extent of ﬁres (and the resultant stand size) appeared to depend strongly on weather conditions. The proportion of an area burned during 20-year intervals (Fig. 18.4) showed great variation, again largely resulting from weather patterns. The very dry summer of 1988 led to ﬁres that burned (to varying degrees) about one-third of the 800 000 ha Park. Climate varies on time scales that are shorter than the life-span of many temperate forests. For example, Johnson and Larsen (1991) examined the ﬁre return interval in an area of 500 km2 of the Canadian Rocky Mountains. Prior to 1730, the climate was warm and dry, with an average ﬁre return interval of about 50 years for forests consisting predominantly of lodgepole pine. Cooler and moister conditions lengthened the return interval to 90 years. Few data are available for the distribution of ﬁre sizes in temperate forests. One of the best data sets comes from the National Forests of the Sierra Nevada range in California (McKelvey and Busse, 1996). Between 1900 and 1992, about 20 to 30%

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Dan BINKLEY

Fig. 18.3. Map of lodgepole pine/subalpine ﬁr (Abies lasiocarpa) stands in an area of 73 km2 in Yellowstone National Park (Wyoming, U.S.A.), as reconstructed for 1778. Number = stage of stand development (1, post-ﬁre; 2, seedling/sapling; 3, immature; 4, mature; 5, transition; 6/7, old-growth; 8, meadows; 9, riparian; 10, steep slopes); approximate stand ages in brackets. Average patch size of historic ﬁres varied substantially (from Romme, 1982), for both the pre-European settlement period (prior to 1870) and during a period of intensive ﬁre suppression (1900–1960).

of the landscape at elevations between 500 m and 1200 m burned, compared with 5–10% of the landscape between 1500 m and 2200 m. The majority of the areas burned were consumed in ﬁres of modest size (<10 000 ha: Fig. 18.5), and about 75% of the burned areas burned only once in this century (and about 15% burned twice). Fuel combustion heats soils, commonly to temperatures in excess of 500ºC in the litter layer, and 200ºC in the upper mineral soil (Agee, 1993). Direct effects of heating include death of ﬁne roots (and perhaps

coarser roots), disruption of microbial populations, oxidation and volatilization of nitrogen, and often wind-blown removal of other nutrients in ash. Losses of nitrogen are probably most critical to the future productivity of temperate forests, and these losses range from a few kilograms per hectare for surface ﬁres to several hundred kilograms per hectare in severe ﬁres (Fig. 18.2). Soils tend to be warmer after ﬁres, as a result of decreased canopy interception of sunlight and reduced albedo of blackened surfaces. Daytime temperatures

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE

Fig. 18.4. Fire patterns for an area of 1300 km2 in Yellowstone National Park; large variations in percent of area burned and in corresponding ﬁre return interval demonstrate the critical role of weather in determining ﬁre regime in any 20-year period (after Romme and Despain, 1989).

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residual organic material on the soil surface can pose a challenge for revegetation (Hungerford, 1980). Accelerated rates of erosion following ﬁre can cause signiﬁcant nutrient losses (Wright and Bailey 1982). Erosion rates may increase on slopes after ﬁres as a consequence of ﬁre-caused changes in vegetation, soil properties, hydrology, and geomorphic processes (Swanson, 1981). However, the actual amount and duration of increase in erosion varies widely among sites as a consequence of ﬁre intensity, soil inﬁltration capacity, topography, climate, and patterns of vegetation recovery. Loss of plant cover and litter exposes soil to increased kinetic energy of raindrops, which may increase particle movement. Water inﬁltration rates may diminish after ﬁre as surface pores are plugged and water repellence develops (Crams and DeBano, 1965). Increased sediment loss as a consequence of the formation of hydrophobic soil layers has been documented in ponderosa pine forests (Campbell et al., 1977; White and Wells, 1981). Increased overland ﬂow and loss of soil binding by root systems result in increased rill and sheet erosion, as well as facilitation of debris ﬂows (Swanson, 1981; Wells, 1987). Post-ﬁre loss of ash due to wind can also lead to important loss of nutrients (see Ewel et al., 1981). Revegetation after ﬁre depends on the reproductive strategies of available plants and propagules, as well as the conditions generated by the ﬁre. Many deciduous tree species can sprout from surviving root-stocks. Some conifers, such as lodgepole pine and jack pine, have serotinous cones which remain viable on trees for many years, and release seeds following ﬁre. Other species depend on seeds remaining viable in the soil, or being transported from survivors within or beyond the ﬁre boundary.

FOREST HARVESTING

Fig. 18.5. Proportion of total ﬁres during each period by area size classes for an area of 850 000 ha in the Sierra Nevada range in California (from McKelvey and Busse, 1996).

of topsoils in summer in burned areas are commonly several degrees warmer than in unburned areas (Christensen and Muller, 1975; Hungerford, 1980; Wilbur, 1985). In some cases, excessive warming of

The history of harvesting in temperate forests has drastically altered almost all disturbance regimes. The eastern United States lost 96% of its old-growth forest to harvest (including land-clearing for agriculture as well as tree-harvest for wood products: Whitney, 1994). The states of Indiana, Maine, Ohio, and Pennsylvania now retain less than 0.05% of their original old-growth forests. In other areas, larger proportions of old-growth forests have been retained; the National Forests of the Sierra Nevada range of California still have about 25% of their old-growth cover (Sierra Nevada Ecosystem

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Project, 1996). Major summaries of these large-scale and pervasive changes have been provided by Williams (1989) and Whitney (1994) for North America; Darby (1956) for Europe; Rackham (1986) for the British Isles; and French (1963, 1983) for Russia. Most temperate forests are managed to remain in use as forest land; the world-wide extent of temperate forests has remained nearly constant since 1950 (Williams, 1994). Much of this territory has been used for intensiﬁed forest management. Forest harvesting (and subsequent forest recovery) is the dominant disturbance for most temperate forests of the world. In the United States, semi-natural disturbances account for a loss of about 110×106 m3 yr−1 of wood volume, compared with 500×106 m3 yr−1 of harvest for wood products (United States Department of Agriculture Forest Service, 1981; MacCleery, 1992). In Russia, total forest area increased by 68×106 ha from 1961 through 1993, primarily through aggressive reforestation efforts (Shvidenko and Nilsson, 1997). During this period, heavy harvesting reduced the extent of mature and overmature conifer forests by about 20%, despite an overall increase in the standing volume of wood across all forests. Average wood volumes in European Russia have increased by about 40% in the past 30 years. The disturbance associated with forest harvesting spans a wide spectrum from relatively minor impacts of selective removal of individual trees to complete removal of the forest followed by intensive site preparation (which may include burning of residual organic matter, soil scariﬁcation or plowing, weed control, and fertilization) and planting of genetically selected, single-species stands. In this section, I brieﬂy describe major silvicultural systems for regenerating forests, from forest harvest through site preparation and tree regeneration. Silviculture is the application of ecological understanding in manipulating the composition and growth of forests to meet management objectives (which typically include revenue from wood products, but may also include multiple-resource objectives such as wildlife habitat). Major English-language discussions of silviculture have been published by Barrett (1995); Nyland (1996); Oliver and Larson (1996) and Smith et al. (1997); these sources provide details about the general topics discussed in this section. The disturbances associated with forest harvest have many features in common with some natural disturbances, including reduction (or loss) of canopies, removal of nutrients, and large increases in resources

Dan BINKLEY

for residual or regenerating vegetation. Forest harvesting tends to be unique in only two ways: the removal of large amounts of wood, and in some cases the compaction of soil. Natural disturbances, such as wildﬁres, may consume the foliage of the forest and leave huge amounts of wood behind; harvesting removes huge amounts of wood, and often (though not always) leaves the foliage and ﬁner materials behind (Kimmins, 1992). The presence of large amounts of woody debris may have important effects on some features of terrestrial ecosystems (such as smallmammal habitat, and regeneration sites for certain tree species), and on most aspects of aquatic ecosystem within forests (Gregory, 1997). The risk factors for disturbance by forest harvest are also different from those for natural disturbances; for example, forest harvest is more likely to be close to roads and population centers (Schmidt et al., 1996). Stand development and disturbances A common characterization of forest-stand development begins with a stand-initiation stage [see Oliver and Larson (1996) and Smith et al. (1997) for elaboration], and this may involve either the planting of genetically selected seedlings, or natural regeneration from on-site propagules. Environmental conditions that inﬂuence the establishment of young trees (and competing vegetation) are particularly important during the stand-initiation stage. Post-harvest environmental conditions may be favorable for tree regeneration in some situations, or very exacting in others. The ecological effects of clear-cutting have been well reviewed by Keenan and Kimmins (1993). These effects include major changes in microclimate. The diurnal range in temperatures typically increases by a factor of two, soil temperature at a depth of 50 cm typically increases by 4–5ºC, and soil water increases as a result of increased inputs (because of reduced evaporation from the canopy surface) and lower evapotranspiration. Hungerford (1980) provided a classic description of the microenvironmental conditions at the soil surface following various types of forest harvest. On a sunny summer day at high elevation in Montana, the high and low temperatures at the soil surface beneath an intact forest were 35ºC and 2ºC, and the frost-free season (between the last frost of spring and ﬁrst frost of autumn) lasted 112 days. A clear-cut stand (where all the trees were harvested) experienced a high of 56ºC (sufﬁcient to kill seedlings) and a low of −5ºC

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE

(sufﬁcient to freeze seedlings) on the same day, and a frost-free season of just 20 days. Soil water content is typically higher after removal of a forest, and nutrient supply may or may not be higher (see review by Ryan et al., 1997). As the canopy and root system of the regenerating forest begin fully to exploit site resources, forests enter a stem-exclusion stage where further recruitment of new trees is unlikely. This stage is characterized by severe competition between trees, and mortality of suppressed individuals. Growth rates commonly peak in this stage, and competition-driven mortality becomes substantial. Selective removal of trees at this stage is called thinning or spacing, and it provides the beneﬁts of current revenue and improvement of the future value of the stand by manipulating its composition and density. With further aging, the deaths of individual trees may open the canopy (and perhaps the soil system) enough to allow successful recruitment of new trees, in an understory-reinitiation stage. This stage is typically (but not universally) characterized by increased diversity in canopy structure, species composition, and spatial heterogeneity. Most forests that are managed with an emphasis on wood production would be harvested either later in the stem exclusion stage or early in the understory-reinitiation stage. An old-growth stage may then be reached where the death of single, large trees creates gaps that are large enough to provide a mosaic of age classes in the stand. Forest-management programs have often started with the harvest of old-growth forests (with great economic value), and conversion of the site to production of wood on shorter rotations. Optimal rotation lengths may be set, based on the timing of the maximum average annual increment, calculated in terms of either wood yield or economic value. Stand-replacing forest harvesting Forest-harvest regimes have been designed to achieve both desired wood products and stand regeneration. Clear-cutting removes essentially all the trees, and favors establishment of shade-intolerant species (which are commonly preferred for high rates of growth). Seed-tree cuts retain enough mature trees to provide seeds for regeneration, but with little effect on the microenviromental conditions of the site. Shelterwood cuts retain enough canopy coverage to moderate microenvironmental conditions (and to provide seeds in

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some cases). These harvest regimes are generally used to produce forests with one age class or cohort, and most commonly with one favored species. Selectionharvest regimes remove a relatively small proportion of a stand in any given time period, perpetuating a multiple-age structure in the forest, and, commonly, a multi-species composition (including some that may be shade-tolerant and slow-growing). All of these harvest regimes are classically referred to as regeneration methods, although this may sound disingenuous to nonforesters. The degree of disturbance associated with forest harvest is probably proportional to the amount of canopy reduction, and is further inﬂuenced by the extent of any soil impacts. Soil disturbance results from inadvertent impacts of machines used in the harvesting operation, and from intentional manipulations as part of the site preparation for the regenerating forest. Heavy machinery can compact soils, reducing aeration and water holding capacity, and increasing soil strength (making it difﬁcult for ﬁne roots to penetrate). Site preparation may involve manipulating residual organic matter, such as piling of wood and stumps into windrows (followed by burning), reshaping of the soil surface (plowing up low beds of raised soil), and sometimes improving soil drainage by digging of ditches. Under the most intensive treatments, forests subject to management resemble agricultural land more than natural forests. Many examples are available to show reduced site productivity following severe site-preparation treatments which include piling of nutrient-rich organic matter into windrows for burning. For example, this windrowing/burning treatment in a plantation of Pinus radiata in New Zealand reduced the growth of trees (through age 17) by 28% (Dyck et al., 1989). Moderateto-severe compaction of soils typically impacts 5% to 30% of a harvested area, reducing productivity of these portions by 5 to >50% (Keenan and Kimmins, 1993). Burning of organic matter after harvesting can remove very large quantities of nutrients (Fig. 18.2), sometimes more than the quantities removed in biomass products. More careful application of intensive site-preparation treatment usually increases the growth of crop trees, but whether this increased growth results from increased site fertility or simply from reduced competition with non-crop tree vegetation (summarized in Walstad and Kuch, 1987) remains largely undetermined. Where sitepreparation treatments include ditches to drain excess

462

water and improve soil aeration, substantial increases in total ecosystem productivity result (Binkley, 1986). The removal of nutrients in biomass may lead to lower productivity following the harvesting disturbance, but empirical data are surprisingly limited. Nutrient removals depend strongly on the types of products removed from a site; stem-only harvests typically remove about half the quantities of nutrients that are lost in whole-tree operations (Binkley, 1986). Forest fertilization is a widespread practice, both at the time of regeneration and later in the forest rotation. The disturbances associated with forest harvest often lead to increased rates of soil erosion, much of the increase depending on characteristics of road construction, and the degree of disturbance of the forest ﬂoor (Binkley, 1986; Binkley and Brown, 1993). In general, erosion rates after harvesting are too low to affect the productivity of the next generation of forest, but often high enough to reduce water quality (especially in association with poor road design). The leaching losses of nutrients typically increase as a result of harvest disturbances, but these increases are always far smaller than the nutrient losses in harvested biomass or through ﬁre (Binkley, 1986). Forest-management programs may be designed to ensure the maximum yield of wood production. Natural sources of mortality are minimized by control of competition (including tree harvest to reduce competitiondriven mortality), and active suppression of ﬁre and insects. Intensively managed forests probably lead to harvest of >80% of the wood produced by the forest, with other disturbance factors accounting for <20% of the wood volume. Forest-harvesting systems can also be designed to promote a wide variety of social objectives, including characteristics of the residual forest that inﬂuence wildlife, aesthetic values, and plant diversity. Recovery of vegetation after harvesting of temperate forests tends to take from 5 to >50 yr, depending on the proportion of trees removed and the parameter of interest. Leaf area and productivity of trees may recover as quickly as 6 years after complete clearcutting for some hardwoods (Zavitkovski and Newton, 1971; Covington and Aber, 1980), or more typically 20–70 years for many conifers (Switzer et al., 1966; Turner and Long, 1975; Long and Smith, 1992; Ryan et al., 1997). These rates are probably similar to those following some natural disturbances (such as windthrow), but faster than some (such as severe wildﬁre). Silvicultural regimes may also be aimed to

Dan BINKLEY

accelerate the development of old-growth structure in younger forests (Newton and Cole, 1987) by selectively removing smaller competing trees to increase the growth of larger trees. Silvicultural regimes may include green tree retention as a part of clear-cutting operations to increase the similarity between clearcutting and natural disturbances (such as ﬁres) that tend to leave some surviving trees (Rose and Muir, 1997; Franklin et al., 1997). The effects of harvesting on the diversity of the subsequent plant community depends on many factors, such as post-harvest site preparation (using ﬁre, heavy machinery, herbicides), planting of selected species, and management of the density (and overall structure) of the population of regenerating trees (Keenan and Kimmins, 1993). Overall, harvesting tends to foster increased diversity of species other than trees for a period that typically lasts until full canopy closure, after which diversity may be reduced owing to heavy shade. This suggestion of a general trend has so many exceptions, however, that it may not be useful. Structural diversity of the forest may be increased by harvesting where single trees or small patches are harvested, or greatly reduced where large cutting units are regenerated with even-aged plantations. Generalization about the effects of disturbances on wildlife is not possible. Individual species may be favored or harmed by disturbances (Patton, 1992; Harris and Harris, 1997). Species relying on high productivity and diversity of understory plants typically increase in population after disturbances that reduce habitat suitability for species with strong afﬁnities for old-growth forests. The response of a species to disturbance in one portion of its range may not apply to other situations or areas. Given this breadth of wildlife responses, useful insights are likely to be developed only for speciﬁc situations, and then only when considered in a landscape context (Keenan and Kimmins, 1993). Another difference between natural stand-replacing disturbances and forest management is the landscape scale of forestry operations. Forest cutting operations tend to have well-deﬁned boundaries, whereas many natural disturbances (such as ﬁre) have irregular and fuzzy boundaries (Hunter, 1990). In the past, most forestry practices have emphasized objectives such as yield of wood products, and, in some operations, maintenance of diverse habitats across landscapes. Spies et al. (1994) provide one of the few quantitative assessments of the effects of forest practices on

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE

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Fig. 18.6. Distribution of the sizes of patches in a 2600 km2 landscape in Oregon, in 1972 and 1988, showing substantial reductions in the average size of the interior closed-canopy forests (>100 m from forest edge) (A). The reduction in large patches corresponded with an increase in the total number of patches in the landscape (B) (from Spies et al., 1994).

landscape-scale patterns. These authors used Landsat imagery to characterize changes between 1972 and 1988 in a landscape in Oregon (U.S.A.) covering 2600 km2 . In 1972, much of this area contained oldgrowth forests (on both private and public lands), with mature, closed-canopy forests occupying 71% of the landscape. Logging over the next 16 years lowered the coverage of closed-canopy forests to 58%. The declines in forest cover were greater on privately owned lands (declining from 50% in 1972 to 28% in 1988) than on publicly owned, non-wilderness lands (where forest cover declined from 79% to 68%). These net changes were the balance between removals by logging (moving landscape units from closed forest to other

conditions), and addition by the succession of formerly cleared areas into closed-canopy conditions. Over the 16-year period, the average rate of conversion from closed-canopy forest was 1.7% yr−1 , and the rate of reestablishment of closed-canopy conditions was 1.3%. These changes in forest cover included major changes in the patch size of the landscape, and in the amount of edges between closed forest and other land types. With “edge” deﬁned as 100 m of closed forest adjacent to another land type, the amount of edge increased from 9% to 13% of the landscape in the publicly owned land. The amount of interior closed forest (more than 100 m from an adjacent land type) declined from 60% to 42% on publicly owned lands, and from 33% to

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12% on privately owned lands. Across the landscape, the amount of the total interior forest in large patches declined (Fig. 18.6), and the number of individual patches (typically of smaller size) increased. Once forest harvesting has generated a landscape pattern of edge and interior space, these patterns tend to be very persistent and difﬁcult to erase (Wallin et al., 1994). Overall, this example demonstrated that forest harvesting creates novel (unprecedented) patterns of forest conditions at landscape scales, as compared with natural disturbance regimes. The implications of these novel changes are not at all clear. Recent research has provided new evidence of the effects of forest harvesting, including better insights about the effects of silvicultural activities on a broad range of forest features, about the spatial and temporal scales of impacts, and about the case-speciﬁc nature of many interactions. Current research and practice in forestry will continue to expand the knowledge base on the effects of harvesting disturbances, including retention of live trees (aggregated and dispersed across a harvesting unit), varying rotation lengths, mixtures of species, and interactions of wildlife, plant diversity, and wood production [see the broad range of chapters in Kohm and Franklin (1997)]. The application of this new knowledge to land management activities will, of course, depend on social and political choices.

CONCLUSIONS

Disturbances are fundamental to temperate forests, across scales of time (from seasons to millennia) and space (from single trees to landscapes). The impact of disturbances depends on the interactions of many ecological factors, and these complex interactions prevent any simple characterization of likely changes in forests. Human-related disturbances have come to dominate most of the temperate-forest landscapes, through harvesting trees (and inﬂuencing revegetation), suppressing ﬁre, introducing exotic species (including pests and pathogens), and altering landscape patterns (size and distribution of stands). Forests are generally robust relative to these disturbances, but the character of the forests after human disturbance typically differs substantially from those following natural disturbances. Major differences include increased production of wood, narrower distributions of age-classes across landscapes, and altered diversity of species and stand structures within the forests.

Dan BINKLEY REFERENCES Agee, J.K., 1993. Fire Ecology of Paciﬁc Northwest Forests. Island Press, Washington, DC., 493 pp. Archibold, O.W., 1995. Ecology of World Vegetation. Chapman and Hall, London, 510 pp. Barrett, J.W., 1995. Regional Silviculture of the United States. Wiley, New York, 643 pp. Barrett, S.W. and Arno, S.F., 1982. Indian ﬁres as an ecological inﬂuence in the northern Rockies. J. For., 80: 647–651. Binkley, D., 1986. Forest Nutrition Management. Wiley, New York, 290 pp. Binkley, D. and Brown, T., 1993. Management Impacts on Water Quality of Forests and Rangelands. USDA Forest Service GTRRM-239, 114 pp. Binkley, D., Richter, D., David, M.B. and Caldwell, B., 1992. Soil chemistry in a loblolly-longleaf pine forest with interval burning. Ecol. Appl., 2: 157–164. Campbell, R.E., Baker Jr., M.B. and Ffolliott, F., 1977. Wildﬁre effects on a ponderosa pine ecosystem: an Arizona case study. USDA Forest Service Research Paper RM-191, Ft. Collins, Colorado. Cannell, M.G.R., 1982. World Forest Biomass and Primary Production Data. Academic Press, London, 389 pp. Christensen, N.L., 1987. The biochemical consequences of ﬁre and their effects on the vegetation of the Coastal Plain of the southeastern United States. In: L. Trebaud (Editor), The Role of Fire in Ecological Systems. SPB Academic Publishing, The Hague, pp. 1–21. Christensen, N.L. and Muller, C.H., 1975. Effects of ﬁre on factors controlling plant growth in Adensotoma chaparral. Ecol. Monogr., 45: 29–55. Clark, J.S. and Robinson, J., 1993. Paleoecology of ﬁre. In: P.J. Crutzen and J.G. Goldammer (Editors), Fire in the Environment: The Ecological, Atmospheric, and Climatic Importance of Vegetation Fires. Wiley, New York, pp. 193–214. Covington, W.W., 1994. Implications for ponderosa pine/bunchgrass ecological systems. In: Sustainable Ecological Systems: Implementing an Ecological Approach to Land Management. USDA Forest Service GTR-RM-247, Ft. Collins, Colorado. Covington, W.W. and Aber, J.D., 1980. Leaf production during secondary succession in northern hardwoods. Ecology, 61: 200–204. Covington, W.W. and Moore, M.M., 1994. Postsettlement changes in natural ﬁre regimes and forest structure: ecological restoration of old-growth ponderosa pine forests. In: R.N. Sampson and D.L. Adams (Editors), Assessing Forest Ecosystem Health in the Inland West. Haworth Press, New York, pp. 153–181. Covington, W.W. and Sackett, S., 1984. The effect of a prescribed ﬁre in Southwestern ponderosa pine on organic matter and nutrients in woody debris and forest ﬂoor. For. Sci., 30: 183–192. Crams, J.S. and DeBano, L.F., 1965. Soil wettability: a neglected factor in watershed management. Water Resourc. Res., 1: 283–286. Darby, H.C., 1956. The clearing of woodland in Europe. In: W.L. Thomas (Editor), Man’s Role in Changing the Face of the Earth. University of Chicago Press, Chicago, pp. 183–216. Dyck, W.J., Mees, C.A. and Comerford, N.B., 1989. Medium-term effects of mechanical site preparation on radiata pine productivity in New Zealand – a retrospective approach. In: W.J. Dyck and

DISTURBANCE IN TEMPERATE FORESTS OF THE NORTHERN HEMISPHERE C.A. Mees (Editors), Research Strategies for Long-term Site Productivity, Bulletin 152. Forest Research Institute, Rotorua, New Zealand. Ewel, J., Berish, C., Brown, B., Price, N. and Raich, J., 1981. Slash and burn impacts on a Costa Rican wet forest site. Ecology, 62: 816–829. FAO, 1995. Forest Resources Assessment, 1990: Global Sythesis. UN Food and Agriculture Organization, Rome. Franklin, J., 1988. Paciﬁc Northwest forests. In: M.G. Barbour and W.D. Billings (Editors), North American Terrestrial Vegetation. Cambridge University Press, Cambridge, pp. 103–130. Franklin, J., Berg, D.R., Thornburgh, D.A. and Tappeiner, J.C., 1997. Alternative silvicultural approaches to timber harvesting: variable retention harvest systems. In: K.A. Kohm and J.F. Franklin (Editors), Creating a Forestry for the 21st Century. Island Press, Washington, D.C., pp. 111–149. French, R.A., 1963. The making of the Russian landscape. Adv. Sci., 20: 44–56. French, R.A., 1983. Russians and the forest. In: J.H. Bater and R.A. French (Editors), Studies in Russian Historical Geography. Academic Press, London. Gregory, S.V., 1997. Riparian management in the 21st Century. In: K.A. Kohm and J.F. Franklin (Editors), Creating a Forestry for the 21st Century. Island Press, Washington, D.C., pp. 69–85. Gresham, C.A., Williams, T.M. and Lipscomb, D.J., 1991. Hurricane Hugo wind damage to southeastern U.S. coastal forest tree species. Biotropica, 23: 420–426. Harris, E. and Harris, J., 1997. Wildlife Conservation in Managed Woodlands and Forests. Wiley, New York, 342 pp. Heinselman, M.L., 1981. Fire intensity and frequency as factors in the distribution and structure of northern ecosystems. In: Fire Regimes and Ecosystem Properties. USDA Forest Service GTRWO-26, Washington, DC, pp. 7–57. Hungerford, R., 1980. Microenvironmental response to harvesting and residue management. In: Environmental Consequences of Timber Harvesting in Rocky Mountain Coniferous Forests. USDA Forest Service General Technical Report INT-90, Ogden, Utah, pp. 37–74. Hunter Jr., M.L., 1990. Wildlife, Forests, and Forestry. Regents/ Prentice Hall, Englewood Cliffs, 370 pp. Johnson, E.A. and Larsen, C.P.S., 1991. Climatically induced change in ﬁre frequency in the southern Canadian Rockies. Ecology, 72: 194–201. Jurgensen, M.F., Harvey, A.E. and Larsen, M.J., 1981. Effects of Prescribed Fire on Soil Nitrogen Levels in a Cutover Douglasﬁr/western Larch Forest. USDA Forest Service Research Paper INT-275, Ogden, Utah. Kauffman, J.B., Christensen, N.L., Goldammer, J.G., Justice, N.L., May, J., Pyne, S.J., Stocks, B.J., Trabaud, L.V., Trollope, W., Weiss, K. and Williams, M., 1993. Group report: the role of humans in shaping ﬁre regimes and ecosystem properties. In: P.J. Crutzen and J.G. Goldammer (Editors), Fire in the Environment: The Ecological, Atmospheric, and Climatic Importance of Vegetation Fires. Wiley, New York, pp. 375–388. Keenan, R.J. and Kimmins, J.P., 1993. The ecological effects of clearcutting. Environ. Rev., 1: 121–144. Kimmins, J.P., 1992. Balancing Act: Environmental Issues in Forestry. UBC Press, Vancouver, 244 pp. Knight, D.H., Yavitt, J.B. and Joyce, G.D., 1991. Water and nitrogen

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outﬂow from lodgepole pine forest after two levels of tree mortality. For. Ecol. Manage., 46: 215–225. Kodama, H. and Van Lear, D.H., 1980. Prescribed burning and nutrient cycling relationships in young loblolly pine plantations. South. J. Appl. For., 4: 118–121. Kohm, K.A. and Franklin, J.F. (Editors), 1997. Creating a Forestry for the 21st Century. Island Press, Washington, D.C., 475 pp. Lertzman, K.P. and Krebs, C.J., 1991. Gap-phase structure of a subalpine old-growth forest. Can. J. For. Res., 21: 1730–1741. Little, S. and Ohmann, J., 1988. Estimating nitrogen lost from forest ﬂoor during prescribed ﬁres in Douglas-ﬁr/western hemlock clearcuts. For. Sci., 34: 152–164. Long, J. and Smith, F., 1992. Volume increment in Pinus contorta var. latifolia: the inﬂuence of stand development and crown dynamics. For. Ecol. Manage., 53: 53–64. MacCleery, D.W., 1992. American Forests: A History of Resiliency and Recovery. USDA Forest Service FS-540, Washington, DC., 59 pp. Martin, R.E., 1982. Fire history and its role in succession. In: J.E. Means (Editor), Forest Succession and Stand Development Research in the Northwest. Oregon State University, Corvallis, pp. 92–99. McKelvey, K.S. and Busse, K.K., 1996. Twentieth-Century ﬁre patterns on Forest Service lands. In: Sierra Nevada Ecosystem Project: Final Report to Congress, Vol. II, Assessments and Scientiﬁc Basis for Management Options. University of California, Centers for Water and Wildland Resources, Davis, pp. 1119– 1138. Newton, M. and Cole, E.C., 1987. A sustained-yield scheme for old-growth Douglas-ﬁr. West. J. Appl. For., 2: 22–25. Nissley, S., Zasoski, R. and Martin, R., 1980. Nutrient changes after prescribed surface burning of Oregon ponderosa pine stands. In: Proceedings Sixth Conference on Fire and Forest Meteorology. Society of American Foresters, Washington, DC, pp. 214–219. Nyland, R.D., 1996. Silviculture: Concepts and Applications. McGraw-Hill, New York, 633 pp. Oliver, C.D. and Larson, B.C., 1996. Forest Stand Dynamics. Wiley, New York, 520 pp. Olson, J., 1981. Carbon balance in relation to ﬁre regimes. In: Fire Regimes and Ecosystem Properties. USDA Forest Service GTRWO-26, Washington, DC., pp. 327–378. Ovington, J.D. (Editor), 1983. Temperate Broad-leaved Evergreen Forests. Ecosystems of the World 10. Elsevier, Amsterdam, 241 pp. Patton, D.R., 1992. Wildlife Relationships in Forested Ecosystems. Timber Press, Portland, 392 pp. Pritchett, W.L. and Fisher, R.F., 1987. Properties and Management of Forest Soils. Wiley, New York, 494 pp. Pyne, S.J., 1993. Keeper of the ﬂame: a survey of anthropogenic ﬁre. In: P.J. Crutzen and J.G. Goldammer (Editors), Fire in the Environment: The Ecological, Atmospheric, and Climatic Importance of Vegetation Fires. Wiley, New York, pp. 245–266. Pyne, S.J., Andrews, P.L. and Laven, R.D., 1996. Introduction to Wildland Fire. Wiley, New York, 769 pp. Rackham, O., 1986. The History of the British and Irish Countryside. Dent, London. Richter, D., Ralston, C. and Harms, W., 1982. Prescribed ﬁre: effects on water quality and forest nutrient cycling. Science, 215: 661–663.

466 R¨ohrig, E. and Ulrich, B. (Editors), 1991. Temperate Deciduous Forests. Ecosystems of the World 7. Elsevier, Amsterdam, 635 pp. Romme, W.H., 1982. Fire and landscape diversity in subalpine forests of Yellowstone National Park. Ecol. Monogr., 52: 199–221. Romme, W.H. and Despain, D.G., 1989. Historical perspective on the Yellowstone ﬁres of 1988. BioScience, 39(10): 695–699. Rose, C.R. and Muir, P.S., 1997. Green-tree retention: consequences for timber production in forests of the western Cascades, Oregon. Ecol. Appl., 7: 209–217. Ryan, M.G., Binkley, D. and Fownes, J.H., 1997. Age-related decline in forest productivity: pattern and processes. Adv. Ecol. Res., 27: 213–262. Sasser, C.L. and Binkley, D., 1989. Nitrogen mineralization in highelevation forests of the Appalachians. II. Patterns with stand development in ﬁr waves. Biogeochemistry, 7: 147–156. Schmidt, T.L., Spencer Jr., J.S. and Hansen, M.H., 1996. Old and potential old forest in the Lake States, USA. For. Ecol. Manage., 86: 81–86. Schoch, P. and Binkley, D., 1986. Prescribed burning increased nitrogen availability in a mature loblolly pine stand. For. Ecol. Manage., 14: 13–22. Shvidenko, A. and Nilsson, S., 1997. Are the Russian forests disappearing? Unasylva, 48: 57–64. Sierra Nevada Ecosystem Project, 1996. Status of the Sierra Nevada, Vol. I, Assessment Summaries and Management Strategies, Sierra Nevada Ecosystem Project, Final Report to Congress. University of California, Centers for Water and Wildland Resources, Davis, California. Smith, D.M., Larson, B.C., Kelty, M.J. and Ashton, P.M.S., 1997. The Practice of Silviculture: Applied Forest Ecology. Wiley, New York, 537 pp. Spies, T.A., Ripple, W.J. and Bradshaw, G.A., 1994. Dynamics and pattern of a managed coniferous forest landscape in Oregon. Ecol. Appl., 4: 555–568. Stocks, B.J. and Simard, A.J., 1993. Forest Fire Management in Canada. Disaster Manage., 5: 21–27. Swank, W.T., 1988. Stream chemistry response to disturbance. In: W.T. Swank and D.A. Crossley Jr (Editors), Forest Hydrology and Ecology at Coweeta. Springer-Verlag, New York, pp. 339–357. Swanson, F.J., 1981. Fire and geomorphic processes. In: Fire Regimes and Ecosystem Properties. USDA Forest Service GTR WO-26, Washington, DC, pp. 421–444. Switzer, G.L., Nelson, L.E. and Smith, W.H., 1966. The

Dan BINKLEY characterization of dry matter and nitrogen accumulation by loblolly pine (Pinus taeda L.) Soil Sci. Soc. Am. Proc., 30: 114–119. Turner, J. and Long, J.N., 1975. Accumulation of organic matter in a series of Douglas-ﬁr stands. Can. J. For. Res., 5: 681–690. United States Department of Agriculture Forest Service, 1981. An Assessment of the Forest and Range Land Situation in the United States. Forest Resource Report 22, Washington, DC. Veblen, T.T., Hadley, K.S., Reid, M.S. and Rebertus, A.J., 1991. The response of subalpine forests to spruce beetle outbreak in Colorado. Ecology, 72: 213–231. Wallin, D.O., Swanson, F.J. and Marks, B., 1994. Landscape pattern resonse to changes in pattern generation rules: land-use legacies in forestry. Ecol. Appl., 4: 569–580. Walstad, J.D. and Kuch, P.J. (Editors), 1987. Forest Vegetation Management for Conifer Production. Wiley, New York, 523 pp. Wells, S.G., 1987. The effects of ﬁre on the generation of debris ﬂows in southern California. Geol. Soc. Am. Rev. Eng. Geol., 7: 105–114. White, P.S. and Pickett, S.T.A., 1985. Introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Orlando, pp. 3–13. White, W.D. and Wells, S.G., 1981. Geomorphic effects of the La Mesa ﬁre. In: The La Mesa ﬁre symposium. Los Alamos National Laboratory LA-9236-NERP, pp. 73–90. Whitney, G., 1994. From Coastal Wilderness to Fruited Plain: A History of Environmental Change in Temperate North America from 1500 to the Present. Cambridge University Press, Cambridge. Wilbur, R.B., 1985. The effects of ﬁre on nitrogen and phosphorus availability in a North Carolina coastal plain pocosin. Dissertation, Duke University. Williams, M., 1989. Americans and Their Forests. Cambridge University Press, Cambridge. Williams, M., 1994. Forests and tree cover. In: W.B. Meyer and B.L. Turner II (Editors), Changes in Land Use and Land Cover: A Global Perspective. Cambridge University Press, Cambridge, pp. 97–124. Wright, H.A. and Bailey, A., 1982. Fire Ecology: United States and Southern Canada. Wiley, New York. Zavitkovski, J. and Newton, M., 1971. Litterfall and litter accumulation in red alder stands in western Oregon. Plant Soil, 35: 257–268.

Chapter 19

“Ecology becomes a more complex but far more interesting science when human aspirations are regarded as an integral part of the landscape.” (Ren´e Dubos, 1980)

ANTHROPOGENIC DISTURBANCE AND TROPICAL FORESTRY: IMPLICATIONS FOR SUSTAINABLE MANAGEMENT G.S. HARTSHORN and J.L. WHITMORE

INTRODUCTION

Anthropogenic disturbance in tropical-forest landscapes spans a huge spatial range as well as an intensity gradient. For example, log extraction by heavy machinery damages many standing trees and the fragile soil (Johns et al., 1996). At the other extreme, mega-scale clearing of tropical forests in Brazil for commercial plantations (Moran, 1981; McNabb et al., 1997), cattle ranching (Uhl and Buschbacher, 1987; Serrao et al., 1996), and surface mining (Parrotta et al., 1997a) have produced extensive deforested landscapes (World Resources Institute, 1992). Even islands of intact forest in a sea of disturbance are inﬂuenced by the severity and type of disturbance around them (Lovejoy and Bierregaard, 1990; Bierregaard et al., 1992; Turner, 1996). The several thousands of years during which indigenous peoples living in tropical-forest regions (G´omezPompa et al., 1987; Bush et al., 1992) were long believed to have had negligible effects on those forests and their biodiversity. Except in urban centers such as those developed by the Mayan culture in northern Mesoamerica, it has been thought that indigenous peoples were present at low population densities and subsisted by harvesting local resources (particularly game, ﬁsh, and fruit) and practicing shifting cultivation (Fig. 19.1) of staple crops [manioc (Manihot spp.), maize (Zea mays)]. Even in wet forest areas, shifting cultivation with an ample fallow period (>20 yr) allowed natural rebuilding of the patch of felled forest. Recent paleoecological evidence (charcoal, pollen, and maize phytoliths) from the Dari´en area of Panama – a famous wilderness area rich in endemic species

(Gentry, 1982) – indicates a 4000-year history of human disturbance and settlement (Bush and Colinvaux, 1994). The data are consistent with depopulation and abandonment of most agricultural areas following the conquest by the Spanish, Portuguese, and other Europeans. The well-developed forests appear to have regrown within 350 years (Budowski, 1970). Nevertheless, the extraordinary biodiversity of the Dari´en area seems not to have been seriously affected by 4000 years of indigenous agriculture. Disturbance in tropical-forest landscapes has become an important focus for ecology and management over the past few decades, largely because disturbance is increasingly viewed as a normal process in most forest ecosystems (Pickett and White, 1985). Foresters and land stewards have long managed forests and trees not only for commercial production of timber, but often to assist or enhance site recovery after disturbance (Parrotta et al., 1997b). Recognition of disturbance regimes has become integral to improved understanding of tropical-forest dynamics and potential land uses. Natural disturbance regimes are the basis for developing more ecologically sound models for managing complex tropical forests (Hartshorn, 1995). A great deal of ecological research and environmental awareness about tropical forests has been generated over the last 30 years. But often it has fallen short in offering solutions to local issues. It is only at this local level that real solutions can be implemented. Local issues must be addresssed one by one and resolved as unique problems, rather than viewing them from afar and trying to prescribe global approaches. The critical questions concern the interactions and synergies of disturbance as they inﬂuence the growth,

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Fig. 19.1. Traditional shifting cultivation of subsistence crops where a small patch of forest is cut and burned (upper photo). Note that the clearings are few and well-spaced in the matrix of forest. Hand-held aerial photo near San Carlos de R´ıo Negro, Amazonian Venezuela. The lower photo shows typical slash-and-burn advancement of the agricultural frontier into tropical forests. Note that older ﬁelds are not abandoned and new ﬁelds are not isolated clearings in the forest matrix. Hand-held aerial photo in the Chapare of Amazonian Bolivia (Hartshorn).

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reproduction, and mortality of the trees comprising forest communities. Not only does disturbance, sensu lato, play a prominent role in natural forests, but it also is of considerable importance to tree plantations. Most of the tree species commonly favored in plantation forestry are early or mid-successional – that is, they require some degree of prior site disturbance to thrive. Tree plantations usually require site preparation, weeding during the early years, and thinning after crown closure, all of which constitute programmed disturbance (Homfray, 1936; Evans, 1992). This chapter focuses on anthropogenic disturbances in tropical-forest ecosystems (including natural forests and woodlands, as well as tree plantations) and their implications for sustainable management of these important natural resources. The tropics are considered broadly here, corresponding to that part of the planet that has supported tropical forests and woodlands over the past few centuries (cf. Richards, 1996). More important than latitude is the moisture regime of tropical-forest and woodland ecosystems, which may span a range of mean annual precipitation from well below 1000 mm to over 10 000 mm. Seasonality of rainfall is usually a principal determinant of forest type as well as the choice of tree species for plantation forestry (Evans, 1992). Inter-annual variations in rainfall seasonality are often signiﬁcant factors in the dynamic biological processes and ecological interactions in tropical forests. Deﬁnitions Global concerns about tropical deforestation, compatibility of forestry and conservation, and sustainability of forest management practices, inter alia, require that the terminology used should be as clear as possible. Therefore, we offer several brief deﬁnitions to aid in understanding what we mean when we use terms like “the tropics,” “forest management,” or “enrichment planting.” The tropics and subtropics include most of the land within the equatorial latitudes 23.5º north and south, including higher altitudes and drier zones. The tropics are not only the lowland, wetter areas such as the classic “rainforests” of the Amazon and Congo basins, but also the much more extensive seasonally dry areas peripheral to the core of wetter forests. The subtropics (roughly from 12–13º to 23–25º latitude) are more prone to tropical cyclones (including hurricanes), tend to be more seasonal in rainfall and temperature,

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and have generally lower species richness than the inner tropics (Holdridge, 1967; Migenis and Ackerman, 1993). Thus, the scope of this chapter is necessarily broad, including a wide variety of ecosystems, soil types and climatic regimes, as well as native and introduced tree species. Forestry is the science, business, and art of creating, conserving, and managing forests and forest lands for the continuing or sustainable use of these resources (Ford-Robertson, 1971). Forestry is one of the key tools in management of lower-latitude forest landscapes. Without appropriate forestry practices in response to or in coordination with anthropogenic disturbances, these forest resources will not survive, as has been repeatedly demonstrated throughout the forested tropics over the last few decades. The determination of which forestry technique (or techniques) might be appropriate in a speciﬁc forest depends on several factors, such as the present condition of the resource, past land use or abuse, disturbance frequency, commercial pressures, poverty levels, ratio of land to population, land tenure, and management objectives. Forest management is the practical application of scientiﬁc, social, and economic principles to the administration and working of forest land for speciﬁed objectives (Ford-Robertson, 1971). Particularly since World War II, deforestation in the tropics has become an increasingly serious problem. Most of the harvesting of tropical hardwoods and the conversion of forests to other land uses has occurred in the absence of forest management (Wadsworth, 1997). It should be noted that prior to 1950 the temperate forest region suffered far greater deforestation than did the tropics (Dubos, 1980), often also in the absence of forest management. Silviculture is applied forest ecology – that is, application of the theory and practice of controlling forest establishment, composition, structure, and growth, for commercial or other objectives (Ford-Robertson, 1971; Smith et al., 1997). Extensive silviculture relies mainly on natural regeneration, sometimes with enrichment planting (Spurr, 1979). Ecologically-sound examples in the tropics exist in Trinidad (Clubbe and Jhilmit, 1992), Surinam (Vega Condori, 1987), Peru (Hartshorn and Pariona, 1993, 1997), Malaysia and Ghana (Moad and Whitmore, 1994). Intensive silviculture on the other hand usually involves tree planting. Productive commercial plantations in the tropics include: eucalypts in Congo (formerly Zaire) and Brazil; pine plantations by the company AMCEL in the Brazilian Amazon, at Usutu in Swaziland, and in Fiji; and Swietenia in

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Table 19.1 Ten top ranked tropical countries for area of natural forests and for area of plantation forests 1 Country

Bangladesh

Total area (km2 )

Natural forests 1990 (km2 )

Rank

Plantation forests 1990 (km2 )

142 776

7 690

Bolivia

1 098 582

493 170

7

2350 280

Brazil

8 506 663

5 611 070

1

49 000

Colombia

1 138 339

540 640

5

1260

Rank

8 3

Cuba

114 494

17 150

2450

7

India

3 287 588

517 290

6

132 300

1

Indonesia

2 042 034

1 095 490

3

61 250

2

Mexico

1 972 546

485 860

8

1090

678 034

288 560

Peru

1 285 215

679 060

4

Sudan

2 505 809

429 760

10

Thailand

513 998

127 350

Venezuela

912 050

456 900

Myanmar

Vietnam Zaire (Congo) 1

332 569

83 120

2 344 113

1 132 750

2350

9 2

8

1840 2030

10

5290

5

2530

6

14 700

4

420

Data from World Resources Institute (1996).

Indonesia and Fiji (Evans, 1992). Silvicultural systems turn the harvest of wood into a tool for establishing the next forest. Plantation silviculture has a much longer history than natural forest management in the tropics. Selective logging is a non-technical term currently in vogue. It is also called selective cutting, creaming, or high-grading (Ford-Robertson, 1971), and refers to commercial harvest of logs, usually only the very best, with no regard for the future of the forest. The selection system, on the other hand, is a silvicultural system that harvests some trees in all size classes either singly, in small groups or strips, in order to promote natural regeneration and stand rebuilding on a nearly continuous basis. The objective is maintenance of an uneven-aged stand – that is, sustainable forest management close to the natural condition – while extracting wood products (Burns, 1983; Society of American Foresters, 1994). Due to the similarity of terms, selective logging and the selection system of harvest tend to be confusing, but they obviously represent quite different practices. A plantation is a forest crop or stand raised artiﬁcially, either by sowing of seed, or planting of seedlings or vegetatively-grown planting stock (FordRobertson, 1971). In 1990, there were 30.8×106 ha of tropical timber plantations (FAO, 1995); this statistic

excludes non-timber tree crops such as cacao (Theobroma cacao) or coffee (Coffea arabica). Tropical timber plantations are dominated by three favorites – eucalypts (Eucalyptus spp.), pines (Pinus spp.) and teak (Tectona grandis). In comparison to the FAO estimate of 1987×106 ha of tropical forests in 1990 (Lanly, 1995), tropical tree plantations cover very little (1.4%) of the tropical-forest landscape (Table 19.1). Despite this huge discrepancy in proportional areas of timber plantations and of natural forests, plantation forestry is the principal source of industrial wood. Plantation silviculture for environmental or protection objectives is also practiced. Examples include eucalypts for windbreaks in Sudan, Prosopis for reclamation in the coastal desert of Peru, Casuarina for quarry reclamation in Kenya, pines in Malawi to protect catchments, and a wide variety of agroforestry plantings to promote better crop production or protect sources of irrigation water. When local communities depend on the forest for products, such plantations tend to yield fuelwood and other wood-based beneﬁts, but only as secondary goals (Evans, 1992). Enrichment planting involves disturbance of the existing forest cover, usually in the form of a line cut through secondary or degraded forest, to facilitate planting of preferred species. After establishment, the

TROPICAL FORESTRY AND DISTURBANCE

canopy is further opened to promote faster growth of the planted trees. The intervening understory vegetation is usually altered only minimally in order to reduce the competition from weedy species, while keeping costs down. Coppice is a regeneration method in which standing trees are cut and subsequent crops originate mainly from adventitious or dormant buds on living stumps, but also as suckers from roots and rhizomes (Burns, 1983). This system is only appropriate for those species that coppice naturally after disturbance. Eucalyptus globulus in the Andes is often regenerated by coppicing. Agroforestry is a collective name for all land-use systems and practices where woody perennials are deliberately grown on the same land-management unit as crops and/or domestic animals, either in spatial mixture or in time sequence, assuming there are both ecologic and economic interactions between the woody and the non-woody components (Khurana and Khosla, 1993). For the past few decades, agroforestry has been a critical component in the efforts to lessen the disturbance effects of traditional agricultural practices (e.g., slash and burn) in the tropics. Many agroforestry practices are based on some of the more successful models developed by indigenous peoples over the centuries (or millennia) that integrate tree planting with agriculture (Wint, 1978; Denevan et al., 1987; Vayda, 1987). Old-growth forest is usually considered to be synonymous with virgin or primary forest, and to be undisturbed by humans (Clark, 1995). This has become increasingly contentious as widespread evidence of human occupation or use has been found even in remote tropical forests (Horn and Sanford, 1993). Although many tropical forests may have been disturbed by indigenous peoples, the scale and intensity of disturbance is difﬁcult to quantify. Most old-growth forests have well-developed ecosystem functions (e.g., tight nutrient cycling, gap dynamics) and harbor impressive biodiversity in spite of previous disturbance. Secondary forest traditionally refers to forest regrowth after catastrophic disturbance. In the past few decades, however, there has been increasing use of this term to describe logged forest (e.g., Brown and Lugo, 1990). Corlett (1994) has asked the key question, “How much disturbance is necessary to make a forest secondary?” Logged forests usually still maintain signiﬁcant functions as ecosystems and for conservation of biodiversity, thus they are excellent

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examples of the difﬁculty in differentiating old-growth from late secondary forest types along a disturbance gradient. In contrast, early to late secondary forests can have signiﬁcantly lower biodiversity than is seen in old-growth forests (Young and Wang, 1989; Zou et al., 1995). One hears a lot about tropical deforestation, particularly in the temperate zone. There certainly are a series of local forest management problems in tropical countries which merit immediate attention, and which, when viewed nationally or regionally, constitute a serious global problem. Each of these local problems is different, with solutions unique to each locality, and must be resolved at the local level. Unfortunately, this chapter by necessity must address the issue broadly. But the far more interesting endeavor is to analyze and solve a particular forest management problem on the ground. Too much global analysis, with too little local application of current knowledge to speciﬁc problems, is a major failing. Armchair resource management might be a good term for this phenomenon; it has become an important issue in dialogue between developed and less-developed countries.

DISTURBANCE EFFECTS

The causes and effects of disturbance in tropical forests are numerous and complex. A review of natural disturbance phenomena in tropical forests is beyond the scope of this chapter. There is a considerable body of scientiﬁc literature on speciﬁc aspects, such as scale (cf. Nelson et al., 1994; van der Meer and Bongers, 1996), magnitude (cf. Janzen, 1988; Murphy and Lugo, 1995), frequency (cf. Hartshorn, 1992; Walsh, 1996), and fragmentation (cf. Vitousek, 1988; Lovejoy and Bierregaard, 1990; Turner, 1996). For more general reviews of natural disturbances in tropical forests, the interested reader is referred to Johns and Skorupa (1987), Arriaga (1988), Appanah (1993), Inoue et al. (1993), Alvarez-Buylla (1994), Boose et al. (1994), Bennett and Dahaban (1995), Denslow (1995), Thiollay (1996) and Walker et al. (1996). In this section we review the interactions between anthropogenic disturbances and tropical forests. Impacts to soils and soil organisms The physical effects of disturbance are often most pronounced on soils. Common phenomena such as

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logging, slash-and-burn agriculture, and conversion to pasture may have profound effects on both physical and chemical properties of soils (Malmer and Grip, 1994; McNabb et al., 1997). Physical effects include compaction, increased bulk density, reduced organic matter, and greater erosion. Chemical effects include loss of nutrients, higher acidity, and lower base saturation. Timber harvesting – particularly the use of heavy machinery to extract logs – can cause serious compaction, rutting, and erosion of tropical soils. While the felling of one or a few trees and the creation of canopy openings are analogous to natural tree falls, the extraction of logs usually causes appreciable damage to soils and remaining trees. Estimates of logging damage range widely. In a detailed quantitative study of direct and indirect logging damage in Sarawak, Nussbaum et al. (1995) calculated that, of a total area of 300 ha surveyed, 5% was used for log landings, 25% for skid trails, 30% was occupied by debris piles, 20% was disturbed forest, and 20% was left as residual undisturbed patches. Damage estimates are comparable in the eastern Amazon (Uhl and Vieira, 1989). Congdon and Herbohn (1993), in northern Queensland (Australia), found that the effects of selective logging were still apparent after 25 years, disturbed soils having higher bulk densities and pH, and lower cation exchange capacity, and concentrations of Kjeldahl nitrogen and available phosphorus; their data suggest that recovery from selective logging is dependent on soil fertility and on the intensity of disturbance. Mycorrhizas play an integral role in most tropical forests, greatly enhancing nutrient uptake by trees through mutualistic associations (Janos, 1980). In west Malaysian forests, mycorrhizal infection of tree seedlings was reduced by 25% after selective logging and 75% after heavy logging (Alexander et al., 1992). Mycorrhizas are also important to the establishment of successional species in abandoned pastures or agricultural ﬁelds (Fischer et al., 1994). In the Ivory Coast, soil mycoﬂora varied with major soil types, but showed clear resilience to drought in 1982–83, recovering quickly after the return of rains (Maggi et al., 1990). Impacts on plants The abrupt opening of the forest canopy, whether by natural tree falls or the felling of timber trees, increases light levels reaching the forest ﬂoor below the canopy

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gap as well as in the forest understory bordering the gap (Williams-Linera, 1990; Denslow, 1995). For shadetolerant understory plants and seedlings, the abrupt exposure to direct sunlight may cause sun-scalding of leaves or even death. In an experimental study of the possible causes of mass mortality of understory seedlings, Lovelock et al. (1994) found that fatalities are due to a combination of photoinhibition with moisture stress. Shade-tolerant species show a greater degree of photoinhibition in forest gaps at midday than do shade-intolerant species. As some of these understory plants may be potential canopy species, their loss can inﬂuence the development of the future forest on that site. In theory, one might suppose that harvesting of a highly prized species would threaten or endanger that species; however, there is limited documentation of such a direct relationship. The true mahoganies (Swietenia spp.) are a noteworthy example. Littleleaf mahogany (S. mahagoni), native to the Caribbean region, has been commercially unavailable for several decades, yet this species is not threatened with biological extinction. Large-leaf mahogany (S. macrophylla), now the focal timber species harvested in the southern Amazon (Ver´ıssimo et al., 1995), also has extensive populations in northern Mesoamerica that have been exploited for at least three centuries (Lamb, 1966; Hartshorn et al., 1984), without this species being seriously threatened with extinction. Genetic erosion in such cases is potentially a problem, but again there are only limited data to support this concern (but see Hall et al., 1994, 1996; Murawski et al., 1994). Intensive harvesting of all size classes of a species can lead to local disappearance, as with Myristica malabarica and Syzygium gardneri in southern India (Daniels et al., 1995). The key to not overharvesting a species is the setting and enforcement of minimum cutting limits which allow individual trees to reach sexual maturity (i.e., ﬂowering and fruiting) before harvesting. Even seasonality can be important, as in the case of large-leaf mahogany, which is mostly logged in the dry season, but whose seeds are not dispersed until the end of the dry season. Thus, a much reduced seed crop is available to provide natural regeneration after logging. The silvicultural method must then ﬁrmly rely on regeneration established prior to harvest, which may or may not be present, and could be destroyed during harvest. Fire is one of the most common agents of disturbance in the tropics. Except in the very wettest

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Fig. 19.2. Tropical forest landscape converted to grassland by annual burning. Note the skeleton trees still standing on the burned slopes along the Lae–Wau road, Papua New Guinea (Hartshorn).

regions, ﬁre is an integral component of slash-and-burn agriculture – used by colonists as well as indigenous shifting cultivators. In seasonally dry ecosystems, ﬁre is typically an annual event (Fig. 19.2), often extending into intact forest. Usually the later in the dry season that ﬁre occurs, the more severe are the effects, killing more standing trees and decreasing the proportion of coppicing by the survivors (Uhl and Kauffman, 1990; Sampaio et al., 1993). The great Borneo ﬁres of 1982–83 were exacerbated by an El Ni˜no drought and fueled by logging slash (Leighton and Wirawan, 1986). Beaman et al. (1985) estimated that the Borneo ﬁres destroyed about 106 ha of forest in Sabah. Tree mortality attributable to the logging ranged from 6 to 12%; post-logging drought caused an additional 12– 18% mortality, while drought and ﬁre together caused 38–72% mortality (Woods, 1989). Impacts on animals Selective logging seems to have only modest effects

on animal diversity (Johns and Skorupa, 1987; Johns, 1992; Bennett and Dahaban, 1995; Frumhoff, 1995; Johns, 1996; Thiollay, 1996). Although mobile animals may migrate away from areas of active logging, they seem to return fairly quickly once logging ceases and to re-establish normal population numbers and guild structures within 3–8 years (Lambert, 1992). Far more serious is the improved access which logging roads offer to hunters and colonists (Fig. 19.3). Commercial or subsistence hunting of bush-meat can seriously deplete popular animal species to the point of causing local extinctions (Wilkie and Finn, 1990; Redford 1992). Heavily-hunted large frugivores are often key agents of seed dispersal for tropical forest trees (Johns, 1987; Terborgh, 1995). When large, previously undisturbed tracts of forest are modiﬁed, large animals such as jaguars (Panthera onca) and tapirs (Tapirus spp.) disappear. In eastern Ecuador, Canaday (1996) found that insectivorous birds of the interior forest were more likely to be absent from disturbed forests and non-forest

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Fig. 19.3. Forest access roads under construction for intensive logging of lowland tropical rainforest, Gulf Province, Papua New Guinea (Hartshorn).

habitats than non-insectivores. Canaday’s disturbance gradient spanned about 3.5 km, and included a coffee plantation on the edge of the study forest. Feinsinger et al. (1988) analyzed the assemblage of hummingbird species along a disturbance gradient in a Costa Rican cloud forest. They found that hummingbird species interactions are most organized in the mature-phase forest, less organized in small tree-fall gaps, and almost undeﬁned in large gaps. The pattern is attributed to greater availability of food (ﬂoral nectar) in the large gaps, yet the hummingbird assemblages in these disturbed patches within old-growth forest were not at all comparable to the more “weedy” hummingbird assemblages in anthropogenic old-ﬁeld habitats. Thus, the authors suggested that responses of consumers to disturbance mosaics in old-growth forests may often be subtle and complex. Many tropical bird species that are generally rare throughout their range are vulnerable to forest fragmentation and disturbance (Terborgh et al., 1990; Hagan and Johnston, 1992). Fragmentation in tropical forests and/or in temperate forests is a concern

because of the effect on neotropical migratory birds, who spend time in both. Selective logging affects species diversity and abundance of butterﬂies in tropical forests (Hill et al., 1995). In Maluku, Indonesia, species richness, abundance, and evenness of butterﬂies were all signiﬁcantly higher in unlogged forest than in selectively logged forest. Six butterﬂy species with restricted geographic distributions were found only in the unlogged forest, constituting a complex butterﬂy community there. The authors suggested that the distributional pattern of tropical butterﬂy species may be used as an indicator of forest disturbance. In Sabah, moths show signiﬁcant loss of diversity (especially at higher taxonomic levels) with forests having been disturbed or converted to plantations (Holloway et al., 1992). Eggleton et al. (1995) studied the species richness of termites in ﬁve forest plots in southern Cameroon with differing disturbance levels. Severe disturbance resulted in a large reduction of termite species, whereas there was little difference in termite species richness between

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slightly disturbed and old-growth forests. Soil-feeding termites dominated in the old-growth and regenerating forests, but were greatly reduced in the severely disturbed forests. Wood-feeding termites appeared to be more resilient to disturbance than soil-feeding termites. Interestingly, the authors found no evidence of secondary invasion of disturbed forests by savannaassociated termite species. In Sumatra, Indonesia, a population of phytophagous beetles (Epilachna vigintioctopunctata) exploded during abnormally low rainfall (1982–83 El Ni˜no), but reproduction was suppressed by normal rainfall (Inoue et al., 1993). The beetle population was limited by food shortage at the end of the favorable dry period, and by high mortality during normal rainy periods. Impacts on ecosystems Along with clearing for agriculture and pasture establishment, logging is well known to be a serious disturbance factor in most tropical forests. In addition to the obvious effects on forest structure, major anthropogenic disturbances such as these also affect ecosystem processes. In a study of three different levels of selective logging in Sabah, Malaysia, Douglas et al. (1992) found a four-fold increase in stream sediment yield after a logging road was built across the upper catchment, and an increase of 5-fold to 18-fold in sediment yield after the 0.54 ha catchment was logged. One year after logging the sediment yield had declined to 3.6 times the amount from a control catchment, indicating fairly rapid partial recovery. In an experimental study of water yields from paired catchments in tropical rainforest in Sabah subjected to different methods of clear-felling, cutting and burning secondary vegetation led to 50% more runoff, while clear-felling and mechanical logging followed by burning increased runoff by about 60% (Malmer, 1992). Extensive surface runoff caused surface and gully erosion along tractor tracks (Malmer, 1996). The ecological effects of selective logging are often downplayed because the resulting opening of the canopy is similar or identical to that caused by natural tree falls. However, selective logging is clearly additional to the background level of natural tree falls. Changes in local wind patterns and eddy effects may erode the gap edge by causing additional tree falls. In Kibale forest in Uganda, heavy logging resulted in large areas of herbaceous tangle, attracting elephants which suppressed forest regeneration by damaging young

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trees, thus perpetuating the herbaceous dominance of the understory (Struhsaker et al., 1996). Of more serious concern is the potential degradation of the general forest canopy by fairly intensive logging of most of the large trees. Ecologists have long recognized the importance of edges, or ecotones. There have been few quantitative studies of synergies between forest fragmentation and edge effects, especially in the tropics. In a study of forest fragments ranging from 1.4 to 590 ha in Queensland, Australia, Laurance (1991) showed that forest fragments had higher canopy and subcanopy damage than non-fragmented forests, as well as exceptional abundance of heavy lianas, climbing rattans and weedy species. Although the most striking changes occurred within 200 m of an edge, change was detectable up to 500 m inside forest fragments. A similar swamping of the margins of forest isolates has been documented in the Brazilian Amazon (Lovejoy and Bierregaard, 1990; Bierregaard et al., 1992). By far the most pervasive and serious disturbance facing tropical forests in general is the extensive conversion of forests to non-forested landscapes (Fig. 19.4). In tropical America, a primary reason for this conversion has been the establishment of pastures for beef cattle. The magnitude and severity of the conversion process have been so great that some scientists have questioned the possibility of forest recovery (e.g., G´omez-Pompa et al., 1972). The overwhelming trend of forest conversion to other land uses, and the widespread use of ﬁre, have generated much scientiﬁc concern about the role of tropical deforestation in the atmospheric build-up of carbon dioxide and the concomitant effects on global climate change. In contrast, the potential effects of global climate change on tropical ecosystems have been largely ignored. Of particular concern in global-change scenarios is the increasing frequency and severity of droughts in the tropics, as has been documented in Panama (Leigh et al., 1990; Condit et al., 1996) and Sabah (Walsh, 1996). Fluctuations in rainfall seasonality affect plant phenology, so that scarcity of fruit may cause local famine among fruit-eating animals (Hartshorn, 1992). Far too little is known about the effects of climate change on tropical forests (cf. Colinvaux et al., 1996; Fuller and Prince, 1996). Recovery and restoration During the last three decades it has been popular to de-

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Fig. 19.4. Highly fragmented tropical dry forest in the western Central Valley, Costa Rica (Hartshorn).

scribe the tropics as fragile (e.g., Farnworth and Golley, 1974; Ayling, 1991). Scientiﬁc evidence, however, does not support the claim of general tropical fragility. The idea of resilient natural systems, even disturbed tropical forests, appears to be a more appropriate interpretation (Dubos, 1980; Orians, 1982; Lugo, 1995). One needs to understand tropical forest resiliency better, for whether forests are resilient or fragile will be the key to how they can be managed. Hurricane damage provides a good example of the resistance and resilience of different types of tropical forests to these devastating events (cf. Walker et al., 1996). In structurally complex rainforest on Guadeloupe (Imbert et al., 1996), high canopy trees served to shield smaller trees, and clusters of tall trees protected forest structure from hurricane damage. In contrast, the structurally uniform mangrove forest showed patchy damage related to species susceptibility. Floristic composition and forest structure may be the principal determinants of the effects of hurricane damage. Hurricanes remove considerable amounts of foliage from surviving trees, increasing

the light intensity in the understory 2–3-fold (Turton, 1992). Some of the best research is being done in the eastern Amazon, particularly in the Brazilian State of Par´a where vast areas were converted to cattle pastures, though many were subsequently abandoned. Researchers noted that areas used intensively for cattle ranching before abandonment were very slow to recover. Nepstad et al. (1996) carried out a comparative study of tree establishment in abandoned pasture and in experimental gaps in adjacent old-growth forest. Treeseedling emergence and coppicing were both about 20 times more in the forest than in abandoned pasture. Principal limitations were the low numbers of tree seeds dispersed into the pasture and much higher seed and seedling predation in the pasture than in the forest. Even physically protected transplants fared poorly in the abandoned pasture, as a result of inimical conditions such as higher air temperatures, lower humidity, and greater soil moisture stress. However, those lands pastured with fewer cattle, or less abused, tended to recover to secondary forest rather rapidly.

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Puerto Rico, on the other hand, has had appreciable abandonment of agricultural lands over the past 60 years. Aide et al. (1995) found that forest recovery in pastures is greatly delayed in comparison with recovery after natural disturbances. Species richness and density of woody species were quite low during the ﬁrst 10 years of pasture succession; woody biomass did not increase substantially for 15 years after abandonment. As is often the case, the presence of grasses and the dominance of ferns signiﬁcantly retard the establishment of secondary forest. When disturbance occurs to such an extent that patches of natural forest become island remnants without a vegetation matrix, recovery becomes of critical importance. What is the threshold beyond which remnant patches lose a signiﬁcant amount of their biodiversity and ecosystem functions? Islands of forest habitat not only lose signiﬁcant numbers of species, but they also are more susceptible to invasion by pioneer or exotic species (Vitousek, 1988; Lovejoy and Bierregaard, 1990; Bierregaard et al., 1992). Restoration of forest landscapes from remnant islands has serious management implications. Colonizing species can facilitate the successional or restorative process. Particularly important is the role of pioneer species such as Cecropia. Early successional phases are characterized by extremely rapid tree growth, biomass accumulation, and leaf turnover. Near San Carlos de R´ıo Negro, Venezuela, it was found that maximum net photosynthesis, leaf nitrogen content and speciﬁc leaf area all peaked within 3 years and declined signiﬁcantly over the ﬁrst 10 years of succession (Ellsworth and Reich, 1996). Changes in species composition and in resource availability combine to produce the common pattern of decreasing leaf nitrogen concentrations and photosynthetic rates during early forest succession. At Los Tuxtlas, Mexico, the dominant pioneer tree, Cecropia obtusifolia, has fast rates of seed-bank turnover, as a result of high seed predation and pathogen attacks (Alvarez-Buylla and Garcia-Barrios, 1991). The patchy establishment of robust, unpalatable shrubs in degraded or abandoned pastures often serves to enhance succession by providing attractive fruits for volant seed dispersers (which defecate or regurgitate seeds of other forest species), as well as more favorable microhabitats for tree establishment. Some examples include Cordia multispicata in the Brazilian Amazon (Vieira et al., 1994) and Melastomataceae in the pine savanna of Belize (Kellman, 1979). Probably more

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frequently than is currently understood, the causes of disturbance may have synergistic or even multiplicative effects. For example, it is hypothesized that two scales and types of disturbance – ﬁre and gaps – in the context of soil patchiness, control the pattern of forest islands in savannas (San Jos´e et al., 1991). Multiplicative disturbances can be devastating. Drought, hurricanes, ﬁres, logging, El Ni˜no events, and others are usually studied singly, which may not lead to a very solid understanding of tropical-forest resilience or fragility.

TROPICAL FOREST MANAGEMENT

In the temperate region, forestry is a fairly simple operation compared to the tropics; and it has not changed much in the last two centuries. The writings of Gifford Pinchot (e.g., 1899; 1903), based on over 100 years of European experience, are rather similar to management guidelines of today. Sustainability and ecosystem management are not new ideas. On the other hand, wishes and goals of public landowners have certainly changed, which is part of the political process. The same forest-management principles, and some of the same techniques, are used to attain these new goals. In the tropics, however, with vastly more complex ﬂora, fauna, and socio-economic conditions, successful forest-management techniques have been much slower to evolve (Vanclay, 1992; Hartshorn, 1996). Colonial foresters, well versed in temperate forestry, tried repeatedly to transplant those practices to the tropics, with nearly 150 years of failures to show for it (see, for example, volumes of the Indian Forester dating back to the mid-1800s). Given the difﬁculties encountered in managing the existing forests, their tendency was to “simplify” or “Europeanize” the complexity of the tropical forest. In other words, they replaced the diverse native forest with monoculture plantations, generally using better-known exotic species. To this day, the most successful tropical forestry, from the perspective of timber production, involves monoculture plantations of very few exotic species (Ewel, 1991; Evans, 1992). Just as in the temperate region, tropical goals are changing; wood production – important as it may be – is only one of many beneﬁts that are now required from tropical forests. Indigenous management systems were, and future management systems for tropical forests need to be, based on programmed disturbance using knowledge

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of the local forest ecosystem (cf. G´omez-Pompa and Bainbridge, 1995). It is unlikely that many of the currently valid indigenous management systems will be actively pursued far into the 21st century, no matter how successful they have been. Sad as it may be, these societies, that have been resilient over millennia, are not likely to fare well unless major and costly efforts are made to encourage their survival. In spite of past attempts at this, they are disappearing one by one. Forest-based indigenous cultures remaining 100 years from now will be more like museum pieces than management systems. Much the same can be said for old-growth tropical forests, and future research and management projects need to respond to these realities. Research programs concentrating on pristine forests tend to identify and describe in detail the many problems in trying to maintain them in an “undisturbed” state. However, researchers rarely get involved in management approaches that can help to solve these problems. Forestry, forest management, silviculture, and related disciplines have the responsibility of converting existing knowledge to solve such problems, wherever the political will and socio-economic realities allow. Forest management aspirations may well be shown in healthy secondary forests that are productive according to 21st century needs and goals, with a certain portion (10–20%) of forests preserved in old-growth natural areas. Although it is now reasonably well documented that disturbance is a central factor in the dominance– diversity relations of tropical forests and that maximum diversity should occur at intermediate levels of disturbance (Janzen, 1970; Connell, 1978; Clark and Clark, 1984), there has not been much success in using the growing knowledge about disturbance to develop more ecologically sophisticated models for forest management (Roberts and Gilliam, 1995; Uhl et al., 1997). Two exceptions are the strip-cut model (Hartshorn and Pariona, 1993, 1997) and medium-scale disturbance to promote natural regeneration of mahogany (Snook, 1996). The former is based on the relatively high proportion of shade-intolerant canopy tree species that require gaps for successful establishment and rapid growth (Hartshorn, 1978; Gorchov et al., 1993). The latter takes advantage of mahogany’s excellent natural regeneration in larger disturbances caused naturally by hurricanes, ﬁres, and river meanders. This feature of mahogany offers promise for regeneration of this valuable species using silvicultural techniques based

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on artiﬁcial disturbance. A recent paper by Whitman et al. (1997) describes a mahogany-logging operation in northern Belize as “not silviculturally sustainable” because its disturbance may be insufﬁcient to promote adequate mahogany regeneration. Unless a stand has plenty of advanced regeneration prior to harvest, “gentler” harvest techniques may leave openings too small for regeneration of the desired species. Post-disturbance management objectives that call for continuation of forest cover normally require forestry practices, more speciﬁcally forest management and silviculture. Managers are seldom willing simply to allow forest to return naturally without an effort to speed up the process, control species composition, or otherwise inﬂuence the establishment of a postdisturbance forest. Rarely are they patient enough to wait the decades required by nature, or able to prevent further disturbance during those decades. This is true for a variety of management objectives, whether wood production, biodiversity values, or watershed protection. Managerial responses to disturbance follow an intensity gradient, just as disturbances do. At one extreme is the low-budget, extensive approach to allow nature to heal the wound by natural regeneration. This can be appropriate in zones of low human population density, where site abandonment for decades is a feasible management tool, and/or where biodiversity values constitute the primary management objective. A slightly more intensive method would be to intervene with techniques that encourage desired species and discourage undesired species. A more intensive approach would be enrichment planting in the secondary forest, whereby lines are cut approximately every 5 m, and seedlings (or seeds) of desired species are planted along the cut lines. Clearing of the lines is then done for several years to allow enough sunlight for the young plants to thrive (Moad and Whitmore, 1994). Several steps up the intensity gradient might involve some form of clear-cuts as a harvest/regeneration technique. Ideally, silvicultural techniques to regenerate new forests are soundly based on knowledge of the local forest and its components. The responses of forests to disturbance offer solid clues as to how new forests can be created, using either natural (extensive) or artiﬁcial (intensive) methods. The methods are many (cf. Evans, 1992; Pinard et al., 1995; ter Steege et al., 1996; Wadsworth, 1997), and are far broader than a choice between a clear-cut or a selection system. They may apply to gaps in healthy forests, restoration of

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disturbed or unhealthy forests, healing of catchments, afforestation of non-forest land, or a variety of other landscape situations. On sites where disturbance has led to clearing of the site, but has not yet resulted in secondary forest regrowth, where food or land is scarce due to heavy human populations, and/or where medicinals or other forest beneﬁts are desired by local inhabitants, agroforestry methods can be used to increase the productivity of the site in terms of food, ﬁber, fuelwood, or other products. There are a wide variety of agroforestry approaches, most of which have evolved from indigenous practices (Lojan Idrobo, 1992; Moad and Whitmore, 1994; Denevan et al., 1987). The agroforestry approach to tropical land use carries with it many advantages. While it often departs from the natural ecosystem more than some would wish, it is far better than many of the more destructive options, and some agroforestry systems can be quite similar to the natural, original forest. Agroforestry can create forests in areas long devoid of trees, as in the successful restoration of native Prosopis to northern coastal Peru, where irrigation assists with establishment on shifting sand dunes. According to Valdivia and Cueto (1979), Prosopis produces beans (362 kg ha−1 yr−1 ), livestock fodder (90 kg tree−1 yr−1 ), and honey (40 kg hive−1 yr−1 ). In general, agroforestry systems have the following important attributes: (1) they tend to be very peopleoriented, with beneﬁts principally targeting local residents; (2) they can help fund, or justify, regeneration of new forests; (3) they employ appropriate technologies; (4) they offer multiple beneﬁts; and (5) they are wildlife-friendly – usually by providing habitat. Agroforestry is generally practiced in rural, small operations controlled by land-owners, but there have been several cases of industrial agroforestry (Whitmore and Burwell, 1986). A variety of private ﬁrms have found agroforestry to be advantageous in their landmanagement projects. Some instances of disturbance are so severe that unorthodox methods of restoration may be required. Scale of disturbance obviously contributes to the severity, with direct consequences for either natural successional processes or human restoration efforts. In the latter case, the use of monocultures or mixedspecies plantations of fast-growing trees may be appropriate techniques, whether the goal is to restore the native ecosystem as rapidly as possible, or to restore productivity to the site for human beneﬁt (Lugo and Liegel, 1987; Parrotta, 1992; Wormald, 1992; Parrotta

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et al., 1997a,b). Some planted tree species can mature and complete the site-restoring cycle in 10–15 years. Whether native or exotic, species must be selected that will rapidly restore site conditions to a point where either the goal of restoration or that of productivity may be met. Indeed, the question arises, is a native species really native to a severely degraded site? At what point must a manager rely on a plantation of certain robust exotics to help restore a site to where native species can re-occupy it? Some exotic weed pests become so dominant on a disturbed site that site recovery is impaired, and restoration can be economically out of the question. Obviously, species that might become uncontrollable weeds must be avoided. Fortunately, current silvicultural knowledge allows good predictability for a wide range of tree species (Brown and Lugo, 1994). Having the technical method worked out, however, is usually not sufﬁcient. The cost of restoring large areas may well be prohibitive, especially in areas where labor costs are high. Tax or other incentives to promote industry involvement in restoration may succeed in some countries (e.g., Brazil, Costa Rica, Indonesia). In cases where the management objective is to restore the disturbed site to a natural condition, abandonment may permit nature to run its course. In cases where recovery is too slow in human terms, or where the site has been badly abused to the point that natural recovery is impeded, then soil restoration and planting of tree seedlings may be required. Restorative plantations often improve microclimate conditions on a site to the point that native vegetation can then get a foothold. Interestingly, it seems to matter little whether native or exotic species are employed in this nurse-crop approach; indeed the natives might not thrive under harsh conditions where exotics often will. This is one of the reasons why certain exotics are widely planted. Obviously, exotics that reproduce vigorously on a given site should not be used, as they can become dominant weeds (Evans, 1992; Brown and Lugo, 1994; Parrotta et al., 1997a). Plantations can be used where the management goal is to produce wood for later harvest, whether for fuelwood, poles, or other products for local use, or for pulp, veneer, or lumber for local use or export. In such cases, any of several techniques and species can be employed, often quite successfully, and often while addressing secondary goals such as site improvement, biodiversity, water harvest values, ecotourism, or wildlife habitat. In the past, most such plantations have

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involved exotic species, often pines, eucalypts, teak, or Leucaena (Evans, 1992). Recent studies (Espinoza and Butterﬁeld, 1989; Russo and Sandi, 1995; Butterﬁeld, 1996) have shown that plantation success with native species can also be promising. Mixed-species (multiple species) plantations have proven successful in many tropical and subtropical countries (Wormald, 1992). Production forestry by plantations can represent an extreme and intensive use of a disturbed site. Special cases are worth noting, where disturbed sites are in need of more intense management in order to restore them to productivity for the beneﬁt of local people. The higher montane regions of the tropics, such as the Andean or Himalayan chains, are often heavily populated with communities totally dependent on fuelwood (Nepal–Australia Forestry Project, 1980; Lojan Idrobo, 1992; United Nations Conference on Environment and Development, 1992; Sharma and Chaudhry, 1997). The original forests are depleted, and, because of steep slopes, soil erosion is a serious problem (Young, 1994). However, plantations often can stabilize the site while offering fuel and other amenities, such as fodder. Reforestation efforts in lowlatitude arid zones, where fuelwood is often the vital forest product for local inhabitants and the resource is frequently depleted when demand exceeds supply, are made more difﬁcult by slow tree growth rates and frequent disturbance by ﬁre and grazing (Evans, 1992; Moad and Whitmore, 1994; Schroth et al., 1996). Perhaps of lesser magnitude is disturbance by hurricanes in Caribbean island nations. High population density and dry ecosystems are common, but forests that have evolved under such a disturbance regime usually respond and recover on their own, albeit slowly. When recovery is too slow, intervention may be required. In each of these cases, montane zones, arid zones, and hurricane-prone forests, plantation forestry is likely to be an important tool in addressing disturbances such as ﬁre, wind, overgrazing, and excessive harvest of fuelwood. Managing native tropical forests for wood harvest using an uneven-aged silvicultural system requires considerable planning, as well as a gentle approach. Low-impact harvesting techniques are gaining in acceptance, particularly where they protect and/or facilitate natural regeneration of preferred species. Such techniques generally employ inventory of harvestsize stems, a survey of advance regeneration, cutting of woody vines (lianas, bush ropes) a year or more before harvest, careful and parsimonious placement of

G.S. HARTSHORN and J.L. WHITMORE

extraction routes, directional felling, and follow-up to assure adequate regeneration of the managed forest. Plantation harvest typically involves clear-cutting as part of the site preparation for the next generation of planted trees. Clear-cutting is an extreme form of disturbance, especially when the site is burned, but can be an aid for regeneration of some species. Certain species of wildlife are also favored by clear-cuts. While large clear-cuts can be ecologically unsound as well as visually unpleasant, smaller clear-cuts can, in some cases, serve as a valuable management tool. Even where patch clear-cuts are used to favor the regeneration of a desired species, it may be necessary to leave advance regeneration to form part of the new forest. The problem in northern Belize as described by Whitman et al. (1997) indicates a need for patch clearcuts to provide better regeneration of mahogany. Managers need to keep two things in mind: (1) If clear-cuts are used, they should be silviculturally justiﬁable and kept small; and (2) forest land needs to be managed for multiple beneﬁts, not just for one or two, even though one of these beneﬁts will surely be considered the dominant one in a given landmanagement plan.

CONCLUSIONS

Disturbance is an integral cause of tropical forest dynamics and regeneration. Tropical foresters and land stewards increasingly are using disturbance to improve ecological forest management, site restoration, and the selection of tree species for plantation forestry. Disturbance of tropical-forest landscapes spans broad spatial and intensity gradients, which are not easily categorized in discrete units. Disturbance parameters such as scale (e.g., from gap to landscape), frequency (from daily to centuries), magnitude (from patch to regional), severity (from extensive to intensive), and patchiness (from mosaic to uniform) all contribute to the impacts and effects. Yet there is very little documentation of their synergistic effects. Disturbance effects are most noticeable in soils and plants, but much less in animals because of their mobility. Common phenomena such as logging, slash-andburn agriculture, and conversion of forests to pasture can degrade or destroy forests, but also may have profound effects on soils. When carelessly performed, mechanized logging is particularly destructive to soils and to many of the remaining trees. Less obvious

TROPICAL FORESTRY AND DISTURBANCE

are the potentially damaging effects of global climate change, particularly its effects on the seasonality or intensity of rainfall. There is increasingly strong evidence of a multi-decade drying trend exacerbated by increasing severity of drought. Seasonally dry areas are quite susceptible to ﬁre, causing very serious burns in logged forests, and many ﬁres are anthropogenic in origin. Fragmentation also has serious effects on forest remnants, with edge effects detectable up to a few hundreds of meters into the remnant. Selective logging has modest effects on much wildlife, although logging roads often permit commercial or subsistence hunters to decimate newly accessible game animals. Responses to disturbance are mostly individualistic, although some broad generalizations are possible. Mycorrhizas are the most studied of soil organisms; they show a decline in abundance due to selective logging and other disturbances, but are fairly resilient to drought. Even though wind disturbance such as hurricanes can wreak havoc with subtropical forests, the partial defoliation of the canopy can trigger gregarious ﬂowering of understory trees and shrubs. Forest types differ markedly in their resistance and ability to recover from hurricanes, largely related to the capacity of the species to sprout new branches or stems. Animals vary greatly in their responses to disturbances. Species and guilds of the forest interior tend to be more susceptible to loss than groups typical of disturbed habitats. Ecosystems are stressed by largescale or high-intensity disturbances. Severely altered systems such as extensive pastures in tropical-forest landscapes may be extremely slow to recover to forest. Robust colonizing species often play key roles in facilitating the successional process. Although tropical forests appear to be fairly resilient to disturbance, there must be adequate seed sources and dispersal agents for natural regeneration of the new forest to succeed. Repetitive or too-frequent disturbance can degrade tropical forests to a state where it may be very difﬁcult to restore well-developed forest habitat. There are many elements involved in managing a tropical-forest resource successfully, whether globally (conceptually) or locally (in practice), whether for wood production or for other products/beneﬁts. The well-being of tropical forests in general is of vital interest to people everywhere. But that well-being is directly tied to the well-being of the people who reside in or near a given forest. One cannot address one without the other. The well-being of the local, forest-

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dependent people requires a productive and healthy forest (Whitmore, 1992). To be sustainable, forests must be managed by nature or nurture. The form of management will depend on the objectives. Silviculture (applied forest ecology), is a principal tool to maintain a healthy, vigorous and productive forest (Smith et al., 1997). Silviculture leads to renovation of the forest, and often requires disturbance (tree removal) to accomplish its goal. Harvest in a managed forest should be conducted in a way to promote regeneration of the new forest. Silvicultural practices include restoring biodiversity, and this is a correct and proper primary goal of forest management in some forests. If 10–20% of a nation’s forests have this as the primary goal, 80–90% of the forest resource is then available in which other goals can take precedence (Whitmore, 1992). There is a place for plantation silviculture, just as there is for intensive cultivation of corn (Zea mays) or rice (Oryza sativa). Preference should be given to native species, more than has been the case to date, but exotics may be better for a given objective and set of conditions. Given widespread deforestation and conversion of forest to other uses, it makes no sense to fell a healthy forest in order to establish plantations. Highly productive plantations of tropical tree species will likely out-compete temperate-zone sources of wood during the next century. Temperate-zone forestry experience does not transfer well to the tropics, where conditions are so different. However, one thing that has been learned in temperatezone forestry merits attention in the tropics. After the primary forest is cut in an area, one often studies the new, secondary forest to develop best management practices. Then, when that secondary forest is harvested, one expects these best management practices still to be valid for the third forest. But the third forest is usually quite unlike the second one, just as the second forest was unlike the primary forest. The third forest will need its own studies and best management practices. In the tropics, where rotations can be much shorter, this lesson is even more critical than in the temperate zone. There is also a tendency to confuse forest management (e.g., selection harvest) with lack of forest management (e.g., selective logging). Many forests are logged with no attempt at management, and the discipline of forest management is credited with “another failure” (Wadsworth, 1997). This is one of the many blind spots to be overcome before

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proceeding with sustainable approaches to tropicalforest management. Swietenia macrophylla is certainly far less abundant now than it used to be in localized areas such as northern Costa Rica. Many tropical forest ecosystems have been converted to agriculture, affecting virtually all native species, not just mahogany. Assuming that the mahogany there is gone as a result of international trade may sidestep the real issue, and lead to false conclusions and ineffective solutions. Repeated attempts to list S. macrophylla in the Convention on International Trade in Endangered Species have prompted accelerated harvest of the resource to the point of extermination. Landowners in tropical nations perceive the Convention on International Trade in Endangered Species as a threat by foreign nations who want to control their property, just as landowners in the United States perceive conservation groups or government agencies as impinging on their rights. The key word here is perception. The intentions of those supporting the Convention on International Trade in Endangered Species, while considered highly suspect by some, are not the issue. Relegating local forest management to international decision-making bodies, when it excludes local decision-makers or managers from the process, cannot lead to rational, local resolution of naturalresource issues. One must ask how useful knowledge can be shared, so that logical solutions can be reached beneﬁtting the planet as well as local inhabitants. In an interesting paper contrasting science and environmental ethics, Sarukh´an (1996) reviewed the work of others in this area and concluded that ecology must be solidly based on science, that most scientists hope to contribute to the preservation of “our common heritage, the Earth”, and that there is an absolute need to assist poorer nations to reach sustainable development comparable to that on the rest of the planet. He warned against belief in a nature that does not really exist, and against using pseudo-scientiﬁc knowledge as a cornerstone in one’s thinking – as do many environmentalist groups. Tropical-forest management has not progressed much, despite considerable ecological research over the past few decades. Several subjects are still in need of research. Research on restoration of productivity to degraded or disturbed sites is one of the most pressing needs (Brown and Lugo, 1994). More study is needed on the suitability of native species for plantations, the optimum utilization of tropical forests and tropical

G.S. HARTSHORN and J.L. WHITMORE

woods, and of non-wood products. However, much of what is already known is not now being used; such implementation would be sufﬁcient to resolve many forest-management problems in land use and ecology. Research alone will serve no purpose without solid programs of extension, education, and technology transfer (Whitmore, 1992). Plantation silviculture with exotics continues to be the main approach to tropical forestry oriented to wood products. Apart from plantations, most harvesting of tropical wood is still based on the cutting of forests not managed in any systematic fashion. Much more attention will need to be devoted to integrating the principles of disturbance ecology into transdisciplinary efforts to manage complex tropical forests, secondary forests, and tropical tree plantations sustainably.

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486 Zagt, R.J., 1996. Ecology and Logging in a Tropical Rain Forest in Guyana: With Recommendations for Forest Management. Tropenbos Series 14, Wageningen, 123 pp. Terborgh, J., 1995. Wildlife in managed tropical forests: a neotropical perspective. In: A.E. Lugo and C. Lowe (Editors), Tropical Forests: Management and Ecology. Springer-Verlag, New York, pp. 331– 341. Terborgh, J., Robinson, S.K., Parker III, T.A., Munn, C.A. and Pierpont, N., 1990. Structure and organization of an Amazonian forest bird community. Ecol. Monogr., 60: 213–238. Thiollay, J.-M., 1996. Distributional patterns of raptors along altitudinal gradients in the northern Andes and effects of forest fragmentation. J. Trop. Ecol., 12: 535–560. Turner, I., 1996. Species loss in fragments of tropical rain forest: a review of the evidence. J. Appl. Ecol., 33: 200–209. Turton, S., 1992. Understorey light environments in a north-east Australian rain forest before and after a tropical cyclone. J. Trop. Ecol., 8: 241–252. Uhl, C. and Buschbacher, R., 1987. Potential productive capacity of abandoned pasture lands in the Brazilian Amazon. In: A.E. Lugo (Editor), People and the Tropical Forest. US Man and the Biosphere Program, Washington DC, pp. 35–36. Uhl, C. and Kauffman, J.B., 1990. Deforestation, ﬁre susceptibility, and potential tree responses to ﬁre in the eastern Amazon. Ecology, 71: 437–449. Uhl, C. and Vieira, I.C.G., 1989. Ecological impacts of selective logging in the Brazilian Amazon: a case study from the Paragominas region of the state of Par´a. Biotropica, 21: 98–106. Uhl, C., Barreto, P., Ver´ıssimo, A., Vidal, E., Amaral, P., Barros, A.C., Souza Jr., C., Johns, J. and Gerwing, J., 1997. Natural resource management in the Brazilian Amazon: an integrated approach. BioScience, 47: 160–168. United Nations Conference on Environment and Development, 1992. Managing fragile ecosystems: sustainable mountain development. Agenda 21, Chapter 13. United Nations, New York. Valdivia, S. and Cueto, L., 1979. Settlement and rural development in the eriazas zones of the north coast of Peru. In: G. De Las Salas (Editor), Proceedings, Workshop on Agroforestry Systems in Latin America, 26–30 March 1979, Turrialba, Costa Rica. CATIE and UNU, pp. 163–169. van der Meer, P.J. and Bongers, F., 1996. Patterns of tree-fall and branch-fall in a tropical rain forest in French Guiana. J. Ecol., 84: 19–29. Vanclay, J.K., 1992. Species richness and productive forest management. In: F.R. Miller and K.L. Adams (Editors), Wise Management of Tropical Forests 1992, Proceedings of the Oxford Conference on Tropical Forests 1992. Oxford Forestry Institute, Oxford, UK, pp. 1–9. Vayda, A.P., 1987. Shifting cultivation and patch dynamics in an upland forest in East Kalimantan, Indonesia. In: A.E. Lugo (Editor), People and the Tropical Forest. US Man and the Biosphere Program, Washington DC, pp. 16–17. Vega Condori, L., 1987. Crecimiento de cedro Cedrela odorata manejado en fajas de rastrojo y en asocio inicial con cultivos. CONIF INFORMA, No. 10. Bogota, Colombia, 20 pp.

G.S. HARTSHORN and J.L. WHITMORE Ver´ıssimo, A., Barreto, P., Tarifa, R. and Uhl, C., 1995. Extraction of a high-value natural resource in Amazonia: the case of mahogany. For. Ecol. Manage., 72: 39–60. Vieira, I.C.G., Uhl, C. and Nepstad, D.C., 1994. The role of Cordia multispicata Cham. as a ‘succession facilitator’ in an abandoned pasture, Paragominas, Amazonia. Vegetatio, 115: 91–99. Vitousek, P.M., 1988. Diversity and biological invasions of oceanic islands. In: E.O. Wilson and F. Peter (Editors), Biodiversity. National Academy Press, Washington DC, pp. 181–189. Wadsworth, F.H., 1997. Forest Production for Tropical America, Agriculture Handbook 710. USDA Forest Service, Washington DC, 563 pp. Walker, L.R., Silver, W.L., Willig, M.R. and Zimmerman, J.K. (Editors), 1996. Special issue: long term responses of Caribbean ecosystems to disturbance. Biotropica, 28: 414–613. Walsh, R.P.D., 1996. Drought frequency changes in Sabah and adjacent parts of northern Borneo since the late nineteenth century and possible implications for tropical rain forest dynamics. J. Trop. Ecol., 12: 385–407. Whitman, A.A., Brokaw, N.V.L. and Hagan, J.M., 1997. Forest damage caused by selection logging of mahogany (Swietenia macrophylla) in northern Belize. For. Ecol. Manage., 92: 87–96. Whitmore, J.L., 1992. Advances in sustainable forest management of natural tropical forest, and in plantation of native species. In: Juan Razali Wan Mohd, Shamsudin Ibrahim, S. Appanah and Mohd. Farid Abd. Rashid (Editors), Proceedings, International Symposium on Harvesting and Silviculture for Sustainable Forestry in the Tropics. Oct. 5–9, 1992. FRIM, Kuala Lumpur, Malaysia. pp. 1–9. Whitmore, J.L. and Burwell, B., 1986. Industrial agroforestry. Unasylva, 38: 28–34. Wilkie, D.S. and Finn, J.T., 1990. Slash-burn cultivation and mammal abundance in the Ituri Forest, Zaire. Biotropica, 22: 90–99. Williams-Linera, G., 1990. Origin and early development of forest edge vegetation in Panama. Biotropica, 22: 235–241. Wint, S.H., 1978. Report on man-made forests in Burma. In: Forest News for Asia and the Paciﬁc, Volume II. FAO, Bangkok, pp. 5–8. Woods, P., 1989. Effects of logging, drought, and ﬁre on structure and composition of tropical forests in Sabah, Malaysia. Biotropica, 21: 290–298. World Resources Institute, 1992. World Resources 1992–93. Oxford University Press, New York, 385 pp. World Resources Institute, 1996. World Resources 1996–97. Oxford University Press, New York, 365 pp. Wormald, T.J., 1992. Mixed and pure forest plantations in the tropics and subtropics. FAO Forestry Paper 103. FAO, Rome, 152 pp. Young, K.R., 1994. Roads and the environmental degradation of tropical montane forests. Conserv. Biol., 8: 972–976. Young, S.S. and Wang, Z.J., 1989. Comparisons of secondary and primary forests in the Ailao Shan region of Yunnan, China. For. Ecol. Manage., 28: 281–300. Zou, X., Zucca, C.P., Waide, R.B. and McDowell, W.H., 1995. Longterm inﬂuence of deforestation on tree species composition and litter dynamics of a tropical rain forest in Puerto Rico. For. Ecol. Manage., 78: 147–157.

Chapter 20

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA ´ Claudio M. GHERSA and Rolando J.C. LEON

INTRODUCTION

The processes and patterns of changes through ecological time at a site are called succession. During this process, the physical environment continually changes along with plant composition, generating a dynamic system characterized by complex biotic–abiotic interactions. Odum (1969) and Whittaker (1975) codiﬁed many of the features of a classical successional model for progressive community development. They suggested that species diversity, community complexity, biomass, and ﬂoristic stability increased with successional time. The concept of succession and the features of a progressive model are applicable to human-disturbed systems (Vitousek and Walker, 1987). Peet (1992) stated that community change is often categorized as either episodic or gradual; it is episodic when the change is discontinuous and generated by exogenous factors (e.g., tillage), and it is gradual, when change is continuous and generated by endogenous factors (e.g., competition). He also stated that the ﬁrst type of community change is generally viewed by ecologists as disturbance, whilst the second type is viewed as succession. Because he believed that episodic and gradual changes could be confounded depending on the scale of observation, he thought that it was not possible fully to separate disturbance from succession. Therefore, both should be included in any treatment of vegetation dynamics. These considerations are especially important for agroecosystems where an annual agriculture cycle is superimposed on natural colonization processes (Soriano, 1971); each year begins with tillage, followed by sowing and crop production, and ﬁnishes with a harvest. As a consequence, a weed community (a weed is considered

here as a plant growing in a cultivated area that is not harvested or grazed) must be responsive to several patterns of change in the physical environment: (1) seasonal variation; (2) agricultural cycles; and (3) long-term environmental trends, such as increasing soil erosion or climate change. Very little is known about the signiﬁcance of succession for the functional properties of agroecosystems. Agroecosystems are characterized by the establishment and management of a modiﬁed and simpliﬁed plant community, often comprising exotic species. This changes the ecosystem by altering the composition and activities of associated herbivore, predator, symbiont, and decomposer communities (Swift and Anderson, 1992). The composition, diversity, structure, and dynamics of agroecosystems may differ in many respects from those of the original ecosystem that dominated the landscape before the onset of agricultural activities. Little information exists on how shifts in weed ﬂora affect function in these systems, and even less information exists enabling one to test whether a succession can really exist in this highly disturbed environment. Most information is skewed by the perception that weeds, pests, and diseases are invaders (Williamson, 1996) which do not follow the generally accepted stages of succession and which have only negative effects on ecosystem function. In agroecosystems, land is grazed, burned, or cultivated in a cyclic way. This means that land is exposed to regular disturbances and has periods with low soil cover, but with high levels of resource availability. Because during these periods nutrient absorption is low and mineralization of organic matter is high, a great proportion of the mineralized nitrogen from organic matter is lost by leaching or denitriﬁcation

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(Tivy, 1990; Smith et al., 1992; Swift and Anderson, 1992). These conditions with high resource availability and a simpliﬁed biotic system make the crop–weed community susceptible to invasion. Each invasion of the community creates a new scenario of instability, which Williamson (1987) calls “press” perturbation – a situation in which the structural and functional properties of the community are modiﬁed by new species rather than by the extinction of ones already present. Forcella and Harvey (1983) showed that both species richness and equitability of the alien arable weeds in the north-western United States have increased since the turn of the century. This suggests that alien arable weeds are following the expected features for a classical successional model, but there is no information on how the invasion of new species has modiﬁed the functional properties of the community. Decomposer organisms play a major role in driving succession in both natural ecosystems and in agroecosystems (Peet, 1992; Schulze and Mooney, 1993; Coleman and Crossley, 1996). In agroecosystems, organic matter is produced in pulses, and a great proportion of its decomposition occurs while the soil has no plant cover. Therefore, a short successional process is started after the crop cycle is ended, which depends on organic matter produced previously. This is called heterotrophic succession (McNaughton and Wolf, 1984) and may be considered a retrogressive process, controlled by the available substrate. Swift and Anderson (1992) hypothesized that the relationship between the number of plant species and ecosystem functions follow a hyperbolic pattern, in which a plateau is reached with a low species number. They believed that this is the point at which the plants take control of the decomposer system. They also considered that species composition is important because different plants can give different physical or chemical signals to the ecosystem. Plants differing in physical structure create different spatial interactions, changing, for example, the volume of the resource space to be exploited. The chemical signals originate both from the productive capacity of the plant (i.e., the input of carbon and energy to the system, the ability to compete for water and nutrients) and the patterns of synthesis of chemicals (e.g., allelopathic molecules, ratios of carbohydrate to lignin). In spite of the fact that decomposition of soil organic matter is considered as a critical factor in ecosystem stability (Smith et al., 1992), the relationship between the number of plant species and composition is often

´ Claudio M. GHERSA and Rolando J.C. LEON

overlooked when agronomists and soil scientists try to understand nutrient availability for crops. Disturbance is expected to stop successional processes and modify the diversity and complexity of the community, depending upon its spatial and temporal dimensions (Glenn-Lewin and van der Maarel, 1992). Recovery of agroecosystems from disturbance is considered as secondary succession, and the dominant mechanisms are population processes (Peet, 1992). Many different kinds of disturbance have been studied (White, 1979; Pickett and White, 1985), but few studies focus on changes occurring in the species composition of weeds at different scales of time and space, and even fewer try to understand the relationships between the processes governing those changes. Recently, Swift et al. (1996) proposed a set of hypothetical models addressing the linkage between agricultural intensiﬁcation and total agroecosystem biodiversity (Fig. 20.1). These models predict that the end result of the intensiﬁcation of agricultural activities will be a system with very low diversity (e.g., intensive cereal production). In Fig. 20.1A, Model IV is an application of the intermediate disturbance hypothesis (Connell, 1978), in the sense that biodiversity remains high and even increases with agricultural intensiﬁcation, until a critical stage of intensiﬁcation is reached. Increase in intensity after this stage will produce rapid declines in biodiversity. Both model IV and model II in Fig. 20.1A allow some intensity of disturbance without losing appreciable diversity. The “hump-backed model” proposed by Grime (1973), and a similar one proposed by Tilman (1982), may also describe successional patterns and the relation between disturbance and diversity (Fig. 20.1B) According to both models species density (the average number of species in a unit area) will increase over successional time to a maximum, after which it will decrease. In both models, disturbance reduces the level of stored live and dead biomass, and allows for release of resources to the soil. Intermediate levels of disturbance thus facilitate maximum diversity, because biomass accumulation is stopped, competition is relaxed, and soil fertility is high. Grime (1979) also considered that life-history strategies replace one another during successional time, starting with ruderal species, going through competitors, and ending with late-successional stress-tolerant species. This last strategy appears when most of the soil resources available for plants during the competitor stage are stored in biomass or necromass of the ﬂora. Grime used his model to discuss maintenance

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we present data from croplands of the rolling pampa at regional or landscape and ﬁeld or patch scales which reveal changes in weed species composition, morphotypes, and physiotypes. We then discuss ideas about the mechanisms that control successional changes in agroecosystems.

DESCRIPTION OF LANDSCAPES AND CROP HISTORY OF THE STUDY SITE

Fig. 20.1. (A) Hypothetical relationships between agricultural intensiﬁcation and total agroecosystem biodiversity. Note that the x-axis is non-quantitative. The four curves illustrate four different scenarios, representing differential effects of agricultural management on total biodiversity with differing implications for conservation. Adapted from Swift et al. (1996). (B) Hypothetical relationships between biodiversity and increasing crop and litter or resources, and stress or disturbance. Vertical arrows indicate probable management intensity and disturbance level for the croplands in the rolling pampa. Curve (a) represents the scenario of Connell, (b) that of Grime and Tilman. Adapted from Connell (1978), Grime (1973) and Tilman (1982).

of monocultures in agricultural systems, and argued that dominance and reductions in species richness are attainable, if practices would allow the development of a large standing crop and minimize the frequency of cropping. Adequate models and essential data do not exist for understanding how patterns of spatial heterogeneity, at large or small scales, regulate population growth or the build-up of communities during secondary succession in agroecosystems. Such understanding is crucial for the design of cultural practices and systems that generate environments which are, at least temporarily, unsatisfactory to certain weed species. In this section,

The pampas [also described elsewhere in this series (Soriano, 1991)] occupy a vast area of Argentina, including the Province of Buenos Aires, and parts of the Provinces of Entre R´ıos, Santa F´e, C´ordoba, La Pampa, and San Luis. This area corresponds to a subregion of the R´ıo de la Plata grasslands, which extends over 70 million hectares of Argentina, Brazil, and Uruguay. The size and shape of opal phytoliths in the soil show that grassland vegetation and the types of grasses composing the vegetation have been invariable throughout the period of pedogenesis (Tecchi, 1983). The entire region has been developed, especially during this century, for the livestock industry and agriculture. The area in Argentina dedicated to cropland increased from ~6 million hectares during the ﬁrst 5 years of this century, to 26 million hectares in 1984 (FAO, 1986). This activity destroyed most of the natural grasslands in the arable areas of the pampas. The ﬂora of the pampas comprises about 1000 species of vascular plants, including several that have been introduced (Parodi, 1947). Although the pampas are considered to be of uniform physiognomy and topography, several subunits are recognized on the basis of geomorphology, drainage, geology, physiography, soils, and vegetation (Soriano, 1991). All the information and the discussion of successional changes in arable land included in this chapter refer exclusively to the subunit called the rolling pampa, which is the main cropland of Argentina (hatched area in Fig. 20.2). This area is gently rolling (Fig. 20.3) and is located between 34ºS and 36ºS latitude and 58ºW and 62ºW longitude. It is bounded by the R´ıo de la Plata and the R´ıo Paran´a on the northeast, by the R´ıo Salado on the southwest, and by the R´ıo Matanza on the southeast (Soriano, 1991). The climate is temperate and humid, without a dry season, but with a hot summer. The annual average rainfall is ~1000 mm and the mean annual temperatures range between 16ºC and 17ºC.

490

´ Claudio M. GHERSA and Rolando J.C. LEON

Fig. 20.3. Summer view of a successional grassland growing in the well-drained Argiudol soil type in a gently rolling landscape in San Antonio de Areco (Buenos Aires Province). The dominant grasses are: Bothriochloa laguroides, Paspalum dilatatum, and Stipa papposa. Other frequent species are Briza subaristata, Eragrostis lugens, Melica brasiliana and Sporobolus indicus. Croplands are visible in the distant background. Trees at the roadsides and around houses are planted.

The representative soil is a Mollisol, the most common type being Argiudol. Wheat (Triticum aestivum), maize (Zea mays), and linseed (Linum usitatissimum) were the most important crops during the ﬁrst stages of the cropping boom, which commenced in 1875; the area harvested for these crops rose to ~4.5, 3.0, and 1.4 million hectares, respectively, at the start of World War II. In the mid-1930s, crops of sunﬂower (Helianthus annuus) became signiﬁcant. After the mid 1950s, grain sorghum (Sorghum bicolor) became widespread and the area dedicated to soybeans (Glycine max) increased from 0.1 million hectares in 1972 to 2.5 million hectares in 1985. Minimum-tillage and zero-tillage practices increased greatly after 1980. Irrigation and fertilization are nearly nonexistent, but recent increases in commodity prices, and the need of farmers to augment revenue, are rapidly changing this situation. Potential vegetation and plant communities of arable land It is difﬁcult to reconstruct the composition and structure of the original vegetation of this region. By Fig. 20.2. Geographic location of the rolling pampa.

491

Number of species

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA

Community Fig. 20.4. Temporal pattern of increase of species producing high quantities of secondary metabolites (alkaloids, terpenoids, and cyanophoric glucosides), relative to the change in total number of species; C, cropland; G, grassland. Adapted from Su´arez (1997).

the time of the ﬁrst ﬂoristic surveys (Parodi, 1930), virgin grasslands were already rare. In the rolling pampa (Fig. 20.2) vegetation structure corresponds to a prairie in humid years and to pseudo-steppe during dry periods (Fig. 20.3). Winter temperatures are rarely a serious limitation, whereas drought in summer frequently inhibits growth of most species. The species that characterize the dominant community in the fertile arable soils are: Bothriochloa laguroides, a grass with short rhizomes which grows only during the high-temperature period of the year; Stipa neesiana, a bunch-grass up to 50 cm high; and three grasses forming small tufts, Aristida murina, Piptochaetium montevidense, and Stipa papposa. Other common grasses are Melica brasiliana, Paspalum dilatatum, and Piptochaetium bicolor. Agrostis montevidensis, Briza subaristata, Bromus unioloides, Danthonia montevidensis, Eragrostis lugens, Panicum bergii, Paspalum notatum, Poa bonariensis, Schizachyrium spicatum, Setaria parviﬂora, and Stipa hyalina form a set of less frequent grass species. Shrubs and suffruticose plants are poorly represented. The most frequent species in this category are Baccharis articulata, B. coridifolia, B. notosergila, B. trimera, Eupatorium subhastatum, Heimia salicifolia, Hedeoma multiﬂora, Margyricarpus pinnatus, and Vernonia rubricaulis. Small broad-leaved herbs and sedges including Adesma bicolor, Berroa gnaphalioides, Carex bonariensis, Chaptalia spp., Chevreulia sarmentosa, Conyza spp.,

Facelis retusa, Hypochaeris spp., Micropsis spathulata, Oxalis spp., Phyla canescens, Polygala australis, Tragia geraniifolia, Verbena spp., and Vicia spp. are interspersed among the grasses (Soriano, 1991).

CHANGES IN THE DIFFERENT AGROECOSYSTEMS

Successional changes in cropped land Regional-landscape scale A clear pattern can be discerned based on the ﬂoristic composition in maize croplands in the rolling pampa, using the data of Parodi (1926, 1930) and two phytosociological surveys carried out in an area of approximately 2.5×106 ha by Le´on and Suero (1962) and Su´arez et al. (1995). Weeds present in maize ﬁelds of the rolling pampa are in a non-equilibrium state similar to that observed in Denmark by Haas and Streibig (1982) and in the northwestern United States by Forcella and Harvey (1983), in which total species richness and species equitability have increased since the turn of the century. The number of species in the original grasslands of the rolling pampa growing in well-drained soils was ~222. It was dramatically impoverished by early agricultural activities to ~53 by 1926 (32 of the original ﬂora plus 21 new) (Parodi, 1926) (Fig. 20.4). This

´ Claudio M. GHERSA and Rolando J.C. LEON

492

B

Species number

Species number

A

Community

Community

Fig. 20.5. (A) Temporal changes in morphotypes of species in the grassland community (G) and the maize crop weed community (C); solid bar, dicots; shaded bar, monocots. (B) Temporal changes in the number of species of different origin; hatched bar, native; shaded bar, exotic; solid bar, cosmopolitan. Adapted from Ghersa et al. (1996).

impoverishment in the ﬂora was not unexpected, as the native habitat for the grassland species was lost by soil tilling. Thereafter, the agricultural landscape was continually invaded by weeds. In 1960, the total number of species in maize crops was 79 (Le´on and Suero, 1962) – 34 from the original grassland and 45 exotic to the grassland; at the present time, the total number of species present in maize crops is 99 (Su´arez et al., 1995), 54 from the original grassland and 45 exotic (Fig. 20.4). Species number increased during the period from 1926 to 1960, and again from 1960 to 1995, at rates of 1.32 and 1.50 species per year, respectively. The average net rate of species increase per year since 1926 is 1.01 (Ghersa et al., 1996). Through this process, a great proportion of the ﬂoristic richness of the rolling pampa has been restored, even though the native perennial grasses have nearly disappeared (Fig. 20.5). The cultivated landscape is characterized as a mosaic, where patches of bare ground are interspersed among areas experiencing different frequencies and intensities of disturbance of the soil surface. The ratio of the landscape area to the disturbance area is crucial in determining the dynamics of the landscape as a whole (Prentice, 1992; Harvey and May, 1997). Cropping activities continually provide propagules of weed species (Radosevich et al., 1997), relax competition by reducing plant density, and mineralize nutrients in the soil organic matter. Thus, communities in arable land can be characterized as

highly invasible (Crawley, 1987). This can explain why species richness has increased over time. Nevertheless, early in the century, most of the species already present in the rolling pampa were excluded from the weed community in maize croplands; but with time an increasing number of the species belonging to the original grassland reinvaded the weed community (Fig. 20.5B). This observation prompts the question as to whether species that were excluded early in the century by cultivation have evolved and adapted to the disturbance regime of maize croplands, or whether the disturbance regime and the environment as a whole have changed, and now are suitable for the weed species originally excluded. Although there is no way of directly answering these questions, some insight can be generated by considering how groups of species distinguished by origin (native–exotic), morphotypes (dicotyledon–monocotyledon), and physiotypes (C3 – C4 metabolic pathways, or production of secondary metabolites: essential oils, coumarins, and alkaloids), have changed over time. The greatest increase in species richness in the rolling pampa was registered for native and exotic dicotyledons. Weeds with high production of secondary metabolites also increased in relation to the total number of species (Fig. 20.4), but the ratio of C3 to C4 species has remained at approximately 1 from 1930 to 1990 (Ghersa et al., 1996). Because the greatest load of herbicide applied in the region was used to control broad-leaved species in cereal crops (Hall

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA

et al., 1992), a reduction in dicotyledonous weeds in relation to grasses would have been expected (Fryer and Chancellor, 1970). At a regional scale, this was not true. The new species to invade the region were frequently dicotyledons (Fig. 20.5A). In agricultural lands, C4 species should increase in abundance with respect to C3 , because they are better adapted to water and nitrogen stress caused by soil deterioration (Baker, 1974). This could be particularly important in the conditions of the rolling pampa, where cropping is carried out in rain-fed systems with low fertilization. In contrast, the C3 :C4 ratio remained unchanged through the study period, despite the changes that occurred in agroecosystems (Hall et al., 1992) and in the speciﬁc composition of the weed community. Climate may be governing the invariant C3 :C4 ratios (Stowe and Teeri, 1978; Fowler, 1981). The seasonal change in the rolling pampa, from a hot summer with a negative water balance, to a temperate winter with positive water balance, allows the existence of both C3 and C4 physiotypes by curtailing any selective advantage experienced by one physiotype in a particular scenario. Holzner (1982) has suggested that climate is a primary factor determining the geographic distribution of weeds, and that anthropogenic and ecological variables occupy subordinate levels, relevant to ﬁtness at the patch scale. If one accepts that the C3 :C4 ratio is controlled by climate, then what is the explanation for the higher relative increase in dicotyledonous species in spite of the negative environments created anthropogenically by herbicides? In the original grassland, dicotyledonous species were more numerous than monocotyledons, and they still remain more frequent in old pastures (Fig. 20.5A). Nevertheless, if one accepts that plant density in cropland is lower than that in undisturbed grassland, and that the crop is the dominant species in the community (all the weed communities that we have described were in well-managed ﬁelds, and most weeds were small and subordinated, usually covering less than 20% of the ground), then one could speculate that the relative increase in dicotyledons relative to monocotyledonss is generated by differences in competitive ability. Monocotyledons are well adapted to high-density environments, with high levels of irradiance. In contrast, dicotyledons are better adapted to lower densities and lower irradiance levels (Koner, 1993). Since most of the light is intercepted by the

493

dominant crop species, dicotyledons should have a competitive advantage. Production of secondary metabolites increases when plants are grown in environments imposing biotic or abiotic stress (Coley et al., 1985; Herms, 1992; Wink, 1993). Species producing high levels of secondary metabolites are well represented in mature stages of succession (Grime, 1979). This can be interpreted as a successional trend imposed by species interactions, as in any secondary succession, or by degradation of the land caused by cropping. In the ﬁrst case, competition and density-dependent mortality increase over time, favoring stress tolerant strategies in response to increases in live and dead biomass. In the second interpretation, stress conditions should appear by reduction in water and nutrient availability for plant growth caused by impoverishment of the physical and chemical properties of the soil. Field and patch scale A patch of vegetation can be deﬁned as an area small enough for all the individual plants growing in it to have strong interactions. This means that the area of a patch can extend only to as much as a few square meters in herbaceous systems or to a thousand square meters in tree systems (Prentice, 1992). In arable land, interactions may be either reduced or enhanced by cultural activities. Cultural activities disperse pests and diseases, facilitate herbivory and source–sink relations, and relax competition. For this reason we are considering that a ﬁeld under a particular cropping system is similar to a patch in a natural system. This similarity should be high when ﬁeld area is small and it decreases with extension of ﬁeld area. At the ﬁeld or patch scale, it is less probable to ﬁnd stability in the community than at the regional or landscape scale. This means that changes in species abundance and richness over short time intervals are more readily expected. At a regional scale, changes in climate and soils control vegetation dynamics, but at the patch scale disturbance has a greater impact. Disturbance related to cropping activities should generate non-equilibrium systems (White, 1979). The pattern of succession should appear only if one looks at long records of successive disturbance cycles (Soriano, 1971; Delcourt et al., 1983). Species richness of weeds increased since 1930, and there is a particular pattern of change in relation to land-use history and system of tillage. In a survey (Le´on and Suero, 1962) of ﬁelds with maize crops grown under conventional

´ Claudio M. GHERSA and Rolando J.C. LEON

494 Table 20.1 Species richness and crop yield for soybean and maize crops in the rolling pampa Crop 1

Maize Maize Soybean 4 Soybean 5

1 2 3 4 5

Year

1960 1990 1990 1990

Number of ﬁelds surveyed

Field yield (kg ha−1 )

15

3360

15

3030

11

Relative yield 2

Species richness 3 Field

Community

1.05

11.2

44

0.95

15.7

48

7940

1.12

22.0

67

10

5840

0.83

15.7

49

15

3470

1.08

13.8

45

26

2910

0.73

10.5

51

13

1630

1.13

16.1

44

6

1260

0.87

14.3

33

The ﬁrst record of each crop system corresponds to high-yield ﬁelds, and the second to low-yield ﬁelds. Relative yield was calculated as the ratio of the yield of a given ﬁeld to the average yield of all ﬁelds surveyed in that year. Field, average number of species recorded per ﬁeld; community, total number of species recorded in ﬁelds with the same community. Conventional tillage. Reduced tillage.

tillage (ploughing and harrowing), values of species richness at the level of ﬁeld and community were higher in ﬁelds producing a low yield of grain (relative to the mean production for the year of the survey) than in ﬁelds with a high yield. The opposite relation was registered in the 1990 survey (Table 20.1; Su´arez et al., 1995). In addition, richness of weed species was analysed in soybean crops grown under conventional tillage and under reduced tillage. In this case, species richness at point and community level increased when a conventional tillage system was used. There was an increase in both ﬁeld species richness and relative yield of soybean seed in ﬁelds cultivated with conventional tillage, but community richness was lower when relative yields were high. In reduced tillage systems, relative yield of soybean seed and species richness increased. In contrast, community richness was higher when relative yields were high (Table 20.1). In the 1990 surveys, ﬁeld and community richness of weeds in maize were higher than those in soybean. This means that the communities of weed species were distinct in spite of the fact that all surveys were conducted in ﬁelds with the same soil types. All these differences in weed species richness show how heterogeneity at the scale of ﬁelds (50–150 ha) induced anthropogenically contributes to community and landscape diversity (McNaughton, 1983). At a regional scale this heterogeneity provides for a diversity of species, allowing succession to continue (Prentice, 1992).

The 1960 and 1995 phytosociological surveys in maize crops (Le´on and Suero, 1962; Su´arez et al., 1995) can be used to classify species according to their constancy, and to observed changes in composition of their groupings in trying to understand successional changes. It can be considered that species present in a high proportion of stands (high constancy values) are well adapted to the regional climate and soil conditions. Those having low constancy values are only adapted to particular local conditions, or are newly invading species. In the 1960 survey Amaranthus quitensis, Anoda cristata, Chenopodium album, Datura ferox, Digitaria sanguinalis, Echinochloa colonum, Paspalum distichum, Physalis viscosa, and Setaria geniculata had constancy values greater than 50% (Group I). Euphorbia lasiocarpa, E. ovalifolia, Polygonum aviculare, Portulaca oleracea, Tagetes minuta, Xanthium cavanillesii, and X. spinosum had constancy values between 20 and 50% (Group II). Species of less than 20% constancy (Group III) added up to 64% of the total number of species, and are not presented here. In the 1990 surveys, Group I remained unchanged except for the loss of three species: P. viscosa and S. geniculata moved to Group II, and P. distichum, to Group III, and the gain of four species: E. lasiocarpa, P. oleracea, and T. minuta, from Group II, and Sorghum halepense from Group III. In the 1990 survey Group II comprised Bidens subalternans, Coronopus didymus, Dichondra microcalyx, Oxalis chrysantha, Physalis viscosa, Setaria geniculata, Sonchus oleraceus, Stellaria media,

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA

495

Fig. 20.6. Multivariate analysis for Tagetes minuta chemiotypes (production and composition of terpenoids) for plants collected from different provinces and grown in Buenos Aires under experimental conditions: principal components analysis by the geographic origin of each plant collection. S, Salta; J, Jujuy; B, Buenos Aires; M, Mendoza; R, R´ıo Negro; A, San Juan provinces. Adapted from Gil et al. (1999).

and Veronica persica. All of these species had low constancy values in the previous survey. Group III in the 1990 survey included 69% of the total species. This can be seen as a replacement of species caused by a successional trend responding to climate and/or soil changes. There is no information suggesting that climate has effectively changed during this period, but it is known that the soil changed. In Argentina very little fertilizer is applied, and tillage frequency has increased since 1970 (Hall et al., 1992). Changes in constancy of some weed species shows that the soil environment in cultivated land is becoming unsuitable for some species, whilst it is becoming better for others. An increase in species richness in the 1990 survey, particularly in Groups II and III, suggests that there is a slow process whereby new habitat is created, and species adapted to it are invading. Tagetes minuta was one of the species that increased its constancy in the course of 30 years. It is known to produce terpenoids and tiophenes that are biologically active as biocides, repellents, and attractors for several insects (Soule, 1993). This South American species was present in the area of the city of Buenos Aires, and in Chile, where it was collected in 1724 and described as a medicinal plant used by American natives (Dillenius, 1732; Cabrera, 1967). Although distributed in a wide geographic area, it was not present in the original

grassland community of the rolling pampa (Parodi, 1926), but now is one of the dominant species in maize ﬁelds. Recent studies show that T. minuta obtained from different locations in Argentina differ in the production and composition of secondary metabolites, thus showing some degree of allopatric speciation. When the collections from different provinces were grown in a single controlled environment, the plants from the rolling pampa (Buenos Aires Province) had the greatest variability in the amount and quality of secondary metabolites compared to the rest of the collections. The range of variation observed for the rolling pampa collection was nearly the same as that observed for all the collections together (Fig. 20.6). Production of secondary metabolites by plants from the rolling pampa was also sensitive to changes in stress related to competition and the presence of nematodes (Gil et al., 1999). The variability and the plasticity in the rolling pampa populations may have appeared as an adaptive adjustment to the agricultural environment. Variability found in many weed populations has been explained by gene ﬂow, enhanced by cultural practices, among populations that had undergone some process of allopatric speciation (Ghersa et al., 1994). Through this process, populations may gain phenotypic plasticity, as in the case of T. minuta plants from the rolling pampa, which allocate energy to defence only under biotic

496

stress. This strategy probably has allowed this species to survive and invade an agricultural environment characterized by pest outbreaks and an alternation of resource abundance (when the soil is ploughed) and restriction (when most of the resources are monopolized by the crop). Differences in point richness and community richness of weeds are partly due to differences in the soil factors that control weed germination, according to observations by de la Fuente (1997) and Su´arez (1997). These authors conducted reciprocal seed plantings in ﬁelds with high and low levels of soil degradation. Seeds came from the weeds present in the community associated with each level of degradation. They also observed germination response of weed species to manipulated soil conditions. Differences in soil temperature in the 5-cm surface layer differentially affected weed germination. The differences in soil temperature were caused by changes in the soil and in the accumulation of litter. Surfaces of less degraded soils were darker and absorbed more radiation than did more degraded soils. The difference in color was caused by higher organic matter content and lower clay content of the soils subjected to less cultivation and erosion. In the degraded soils, clay content in the plough layer increased because the eroded organic horizon was mixed with the clay-rich B horizon. Litter on the soil surface was related to the reduced tillage systems. Promotion or reduction in seed germination of weed species was in accordance with differences observed in point and community richness. In soybean crops grown with conventional tillage, for example, constancy of Bidens subalternans was 4% when the soil was highly degraded, as against 19% when the soil was less degraded. Experiments on how seed germination of this species is regulated by soils with characteristics similar to those in the ﬁelds surveyed revealed that the changes in species constancy were associated with differences in both seed germination and seedling emergence (Fig. 20.7). When changes in the soil environment are not related to the mixture of soil horizons caused by erosion and tillage, the most important changes for the seeds are caused by the effect of tillage on litter accumulation. The effect of tillage system, fertilization, and crop residues on carbon balance and respiration, as well as on the distribution of organic matter and microfauna in the soil, was evaluated in various studies conducted in different sites of the rolling pampa (Pilatti et al., 1988;

´ Claudio M. GHERSA and Rolando J.C. LEON

Alvarez et al., 1995a–e; Alvarez et al., 1996). These studies showed that the main changes are related to soil temperature and stratiﬁcation in mineralization. Litter accumulation on the soil surface following reduced tillage diminishes the average maximum and daily range of soil temperatures. In conventional mouldboard ploughing systems, both organic matter and decomposer activity are distributed throughout the 15– 20 cm plough layer, whereas in the undisturbed and reduced-tillage systems, activity is concentrated in the top 5-cm layer of soil, and decreases exponentially with depth. In spite of these differences in stratiﬁcation of mineralization activities, emission of carbon dioxide and the overall carbon dynamics in bulk soil were not changed by tillage system or nitrogen fertilization during the 5 years of observation (Alvarez et al., 1996).

Fig. 20.7. Cumulative seedling emergence of Bidens subalternans in the following treatments: high soil degradation, light surface (solid triangles); low soil degradation, dark surface (solid squares), and high soil degradation, artiﬁcial dark surface (solid circles). Lines represent regression models ﬁtted to the observed values. From de la Fuente (1997).

Successional changes in pastures We consider pastures separately from croplands because they differ in frequency and intensity of anthropogenic disturbance. When sown, pastures are like any cropland, but subsequently succession occurs under the less disturbed conditions of grazing. Succession in these conditions is fairly rapid at the beginning, but then slows down, depending on the intensity and

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA

Field and patch scale Successional studies of the rolling pampa (Le´on and Oesterheld, 1982) show that ﬁelds sown as pastures,

Relative cover (%)

Years Fig. 20.8. Changes in cover of planted (solid line) and spontaneous (dotted line) species in pastures of the rolling pampa. From Le´on and Oesterheld (1982).

Diversity

Regional and landscape scales Phytosociological surveys carried out in 1995–1997 in 45 ﬁelds with old pastures (>15 years since sowing) made it possible to compare their ﬂoristic composition with one of the natural grasslands described by Parodi (1930). Surprisingly, both the 1930 and the 1995– 1997 communities had ~220 species (Fig. 20.5A). Although changes in species composition occurred, because total species number of the old pastures is composed of 147 of the grassland original species and 75 new ones (Fig. 20.5B), a successional climatic stability with regard to diversity persists. The average rate of increase of new species was 1.15 species yr−1 . The structure of the community remained unchanged, but replacement occurred in about 25% of rare and subordinate species. The most important species that were replaced included Cenchrus myosuroides, Desmanthus sp., Trifolium argentinense, and T. polymorphum. The new species in the community are weeds such as Crepis sp., Hypochaeris radicata, Senecio burchelii, planted ornamentals like Leucanthemum vulgare, and forage species such as Agropyron elongatum, Festuca arundinacea, Lotus tenuis, Medicago sp., Melilotus alba, Phalaris aquatica, and Trifolium sp. It is important to note that the rate of replacement in a grassland community through the last 70 years is quite similar to that observed during the same period for the maize weed community. This rate of species replacement has been observed in other grasslands as a response to grazing or anthropogenic disturbances (Sala et al., 1986; Swift et al., 1996). An important question remains unanswered: do the new introductions perform the same function as do the species that they replaced [probably not], and what are the consequences of changes on the overall behavior of the system?

with similar management practices and soil types, experience a reduction in the cover of the planted species and increases in species richness and diversity (as measured by the Shannon–Weaver index: Whittaker, 1977; see Figs. 20.8, 20.9). At late stages, the dominant species are those of the native community (Le´on et al., 1984). Apparently, the combined effect of cattle grazing and planted species does not inhibit or delay succession, which appears to be driven by soil and climatic conditions.

Number of species

frequency of grazing. Vegetation is relatively stable so long as abiotic stress or biotic factors inhibit or at least delay further succession (van Andel et al., 1993). As soon as the density of herbivores is reduced below the level of carrying capacity by an external factor, the rate of vegetation succession increases. Moderate grazing generates a patchy system, where some patches are heavily grazed while others are untouched. Succession proceeds in the ungrazed patches (Bakker et al., 1983).

497

Years Fig. 20.9. Changes in species richness and diversity in pastures of the rolling pampa. Shannon–Weaver index of diversity (solid line), species richness (dotted line). From Le´on and Oesterheld (1982).

Successional changes in wastelands, relicts, and corridors The rolling landscape of this region is shaped by geomorphological processes, highly distorted by human activities. Anthropogenic modiﬁcation of the landscape takes place in different ways. The region

498

´ Claudio M. GHERSA and Rolando J.C. LEON

Fig. 20.10. Succession in wasteland and a fence corridor in the rolling pampa. The dominant woody species at road borders are Baccharis articulata and B. notosergila, and, near the fence, Gleditsia triacanthos and Parkinsonia aculeata

is divided by fences to prevent animals from grazing grain crops, as well as by roads, railways, and electric or telephone lines. Together, these generate a network of corridors of semi-natural (non-cropped) vegetation. Because the average size of ﬁelds ranges between 50 and 100 ha, even narrow corridors (1– 300 m wide) cover a signiﬁcant proportion of the landscape. Moreover, the region is characterized by small areas of abandoned land near corrals, silos, houses, and railway stations, as well as on the outskirts of small towns and cities. These small areas differ in size and frequency, and range from a few square meters to a few hectares. In these areas, the structure of the community is modiﬁed because poorly represented shrubs and trees from the semi-natural community invade disturbed and abandoned areas. The grass community is dominated by perennial grasses, such as Cortaderia selloana and Paspalum quadrifarium, forming large tussocks mixed with weeds like Dipsacus fullonum and Sorghum halepense. The main woody species are native taxa such as Acacia bonariensis, Baccharis sp., Discaria longispina, Aloysia gratissima, Parkinsonia aculeata, and Sphaeralcea bonariensis, and exotic species such as Broussonetia papyrifera, Gleditsia triacanthos, Ligustrum sp., Melia azedarach,

and Morus alba. Gleditsia triacanthos is also invading riparian zones along the many rivers and creeks of the rolling pampa landscape. These areas are important habitats for wildlife, woody species, and perennial grasses from the original community. Small mammals such as rodents and armadillos, as well as ﬂying and walking birds, use these corridors to escape cropping activities and predation. The absence of agricultural activities, together with sporadic ﬁre and dispersal of tree seeds by birds that perch on fences (Montaldo, 1993), are the main factors enhancing the invasion of corridors, waste land, and relicts by woody species (Fig. 20.10). Together, these factors drive succession away from the original grassland. Facelli and Le´on (1986) and Mazia et al. (1996) carried out experiments in the inland pampa to evaluate if the grassland could be invaded by woody species. Their results support the idea that, when soil tillage is absent, and some removal of above-ground biomass of grass takes place, Gleditsia triacanthos, Prosopis caldenia, and Ulmus pumila can become established. The successional process, driven by the invasion of woody species of the waste land and corridors, generates areas with dense populations of shrubs and trees, many of them having thorns. This creates

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA

problems of visibility, particularly at road intersections and bridges, and impedes access to fences. There is also the perception that these areas are habitat for unwanted mammals such as foxes, mice, and armadillos, as well as for weeds and diseases. For this reason, land owners and government are forced to remove trees to keep fences and roads clean. Nevertheless, some patches have escaped this removal and are now small forests.

CONCLUDING REMARKS

We have shown that, at the regional scale, there has been an increase in species number in the spontaneous vegetation of croplands in the rolling pampa. The increase has occurred without loss of the dominant weeds, and some functions (e.g., the C3 and C4 photosynthetic pathways, and production of secondary metabolites). This corresponds to the “press” perturbation model (Williamson, 1987). In contrast, total species number in the old pastures was quite stable over time, in spite of the fact that some species replacement occurred. In these old pastures, the dynamics of the community ﬁts better with a classical model for secondary succession, in which changes are driven by competitive exclusion. Less frequent soil tillage in pastures than in cropland allows for high accumulation of biomass in pastures and increases in plant density, so that competition among plants is increased. Nevertheless, in both the croplands and the pastures, total number of species and the structure of the community are apparently constrained by soil and climate. At the regional scale, the rate at which agricultural land gained species was surprisingly similar to the rate at which new species invaded pastures, replacing the original species of the rolling pampa grassland (Parodi, 1930). This could just be a coincidence, or the same process could be regulating the invasion of new species in the highly disturbed cropped ﬁeld and the less disturbed pastures. It is possible that, in both the cropped land and the pastures, the arrival of propagules of new species and soil and climatic constraints are governing the dynamics of the weed ﬂora at similar rates. In cropland, the press perturbation caused by the dominant crop, which is added each year to the spontaneous weed community, creates a “crop climatic environment” allowing for an over-representation of

499

dicotyledonous species. In the old pastures dicotyledons are also over-represented, probably because of the effects of grazing. In spite of the fact that woody species have in the past been poorly represented in the rolling pampa, the successional process observed in the wastelands and relicts unequivocally suggests that, should agriculture stop, the grassland would change to some kind of woodland or a savanna type vegetation. Data at the regional scale do not correspond to any of Swift’s models, all of which predict a regressive succession, in which diversity decreases as intensity of disturbance increases (Fig. 20.1). After agriculture expanded over the grasslands in the rolling pampa at the beginning of this century, species richness and landuse intensity increased together (Hall et al., 1992). Although ﬁelds of the rolling pampa are not irrigated, and still have a relatively low load of pesticides and fertilizers as compared to Europe and the United States, today’s agriculture in the rolling pampa is intensively managed. On the other hand, the situation in the 1930s was characterized by low-intensity management. Pastures with disturbance intensity lower than that experienced by ﬁelds with annual crop species have a higher species richness. This would be in agreement with a decay in diversity in relation to an increased disturbance intensity. Considering the patch or ﬁeld scale, our data could ﬁt Connell’s Model IV because: (1) community richness of the weed species grew with an increase in the period under maize production (low relative grain yield is related to high richness of the weeds); (2) the opposite occurred in the 1990 survey, (low relative yields of maize are related to low richness of the weeds); and (3) in soybean croplands, richness under conventional tillage is higher than under reduced tillage (Table 20.1). Low yields are directly related to soil degradation, and this in turn is related to the intensity of disturbance. According to Grime’s (1979) concepts of life strategies, the replacement of life strategies over time in the agroecosystem we studied probably follows a similar trend to that in natural succession. This is based on the assumption that availability of soil resources follows a similar trend in natural and agricultural ecosystems, but that the change is caused by different mechanisms. In the natural system, soil is depleted by biotic consumption, thus stress-tolerant strategies are favored in mature successional stages. In croplands, soil resources are scarce as a consequence of erosion, and because nutrients are exported with the harvested crops. If the increase in constancy of species that produces a

500

high level of secondary metabolites (Fig. 20.4) reveals an increase in stress tolerance, then soil degradation may be forcing replacement of competitors by stresstolerant species. Recent invasion of cultivated lands by native species of the original grassland, which in early stages of the cropping history were excluded from the maize weed community, further supports this idea. If the original grassland was mature, species of that community would have been stress-tolerant. When soils were highly fertile, native species were excluded by crops and by weeds with competitive strategies. Now that the soils represent different degrees of degradation, native species that are stress-tolerant can reinvade. The overall effect of increased representation of the stresstolerant strategy in the ﬂora of the agroecosystem is unknown, but stability should increase over time, as indicated by increase in species number and/or species replacement. A weed community dominated by plants producing terpenoids, thyophenes, coumarins, latex, and alkaloids should have an important biological impact, and could negatively affect pest outbreaks (Wink, 1993). Stress-tolerant weed species themselves are more difﬁcult to control, as evident from the stability of the major weeds present in maize crops, despite the efforts invested in controlling them. Both of these attributes (i.e., stress-tolerance, and production of secondary metabolites) have consequences to pest and weed management of agroecosystems. A tradeoff may be possible; the beneﬁcial effect of stresstolerant weeds in pest control may be balanced against the negative effect on crop yield. This trade-off differs between the landscape and the ﬁeld level. There are several processes operating simultaneously in the agroecosystems of the rolling pampa. There are changes in the landscape related to human activities, creating variability in “gamma” diversity. Abandoned land, roads, corridors, wasteland, and a diversity of crops and farming practices generate a large diversity of habitats. This structural complexity at the regionallandscape level creates conditions suitable for invasion processes, whereby the dynamics and composition of the weed community become insensitive to changes in agricultural practices over time. Habitat diversity creates refuges for native and exotic plant species, impeding species extinction and allowing the development of a population to a threshold size, above which it can withstand the environmental stochasticity of the agricultural landscape and sustain recolonization of empty patches (Mack, 1995)

´ Claudio M. GHERSA and Rolando J.C. LEON

Therefore, large-scale invasion processes are regulated by climatic constraints to succession, and are important in maintaining species diversity. This diversity functions as the genetic and species bank from which weed invasions on the patch or ﬁeld scale and species evolution can be nourished continually. At the ﬁeld or patch scale, the composition of the weed ﬂora associated with a crop is dynamic and responsive to changes in agricultural practices over time. Community dynamics and structure are controlled by the crop understorey environment, and by the physical and chemical properties of the soil. Most of the hypotheses describing how agricultural disturbances affect biodiversity are based on differences in diversity among plots differing spatially in the intensity and frequency of the disturbance (Pimentel et al., 1992; Swift and Anderson, 1992). It is clear that disturbance reduces biodiversity, and this effect can be thought to end in a monoculture when disturbance is extreme. Alternatively, if changes in biodiversity are observed following each type of disturbance over time, the effect of disturbance may be quite different. Successional and evolutionary processes will tend to reduce the original differences among levels of disturbance. For example, in the rolling pampa in 1930, the grassland community had 169 more species than did the maize community. In 1996, the difference was reduced to 123 species, and in areas cultivated with maize the number of species that belonged to the original grassland community increased from 25 to 46. It is notable that successional changes in the stresstolerant weed community in the arable land of the rolling pampa can lead to stability similar to that which can be expected for a mature successional stage. The successional trend observed for the rolling pampa is similar to that observed for weeds in Europe and in the United States. Not only is succession not stopped by agricultural activities, but it follows the expected trend for natural systems: regional diversity increases with time. At the regional and patch scale, physiotype (C3 :C4 ratio) is more stable than species number, morphotype, or chemiotype. Unfortunately, weed researchers have often overlooked the longterm ecological processes involved in colonization, succession, and equilibrium. This oversight can lead to misinterpretation of the factors causing species shifts and limit predictive ability concerning future weed invasions, species losses, or evolution of herbicideresistant populations. Clearly, it will be a long time before accurate predictions are commonplace. Until the

SUCCESSIONAL CHANGES IN AGROECOSYSTEMS OF THE ROLLING PAMPA

key factors regulating successional changes in agricultural lands can be identiﬁed, adequate management of weed or other pest populations will not be possible.

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502 inﬂuence of environmental stochasticity. In: Weeds in a Changing World. BCPC Symposium Proceedings, 64: 65–74. Mazia, C.N., Chaneton, E.J., Le´on, R.J.C. and Ghersa, C.M., 1996. Tree species colonization in pampean grasslands and forest plant communities. Proc. Annu. Meet. Ecol. Soc. Am., 77: 290. McNaughton, S.J., 1983. Serengeti grassland ecology: the role of composite environmental factors and contingency in community organization. Ecol. Monogr., 53: 291–320. McNaughton, S.J. and Wolf, L.L., 1984. Ecolog´ıa General. Omega, Barcelona, 713 pp. Montaldo, N.H., 1993. Dispersi´on por aves y e´ xito reproductivo de dos especies de Ligustrum (Oleaceae) en un relicto de selva subtropical en la Argentina. Rev. Chil. Hist. Nat., 66: 75–85. Odum, E.P., 1969. The strategy of ecosystem development. Science, 164: 262–270. Parodi, L.R., 1926. Las malezas de los cultivos en el partido de Pergamino. Rev. Fac. Agron. Vet. Univ. Buenos Aires, 5: 75–188. Parodi, L.R., 1930. Ensayo ﬁtogeogr´aﬁco sobre el partido de Pergamino. Estudio de las praderas pampeanas en el norte de la Provincia de Buenos Aires. Rev. Fac. Agron. Vet. Univ. Buenos Aires, 271. Parodi, L.R., 1947. La estepa pampeana. La vegetaci´on de la Rep´ublica Argentina. Geograf´ıa de la Rep´ublica Argentina. An. Soc. Argent. Est. Geograf., 8: 143–207. Peet, R.K., 1992. Community structure and ecosystem function. In: D.C. Glenn-Lewin, R.K. Peet and T.T. Veblen (Editors), Plant Succession, Theory and Prediction. Chapman and Hall, London, pp. 103–151. Pickett, S.T.A. and White, P.S. (Editors), 1985. The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Orlando, 472 pp. Pilatti, M.A., de Orellana, J.A., Priano, L.J., Felli, O.M. and Grenon, D.A., 1988. Incidencia de manejos tradicionales y conservacionistas sobre propiedades f´ısicas, qu´ımicas y biol´ogicas de un argiudol en el sur de Santa Fe. Cienc. Suelo, 6: 19–30. Pimentel, D.A., Stachow, U., Takacs, D.A., Burbaker, H.W., Dumas, A.R., Meaney, J.J., O’Neil, J.A.S., Onsi, D.E. and Corzilius, D.B., 1992. Conserving biological diversity in agricultural and forestry systems. Bioscience, 42: 354–364. Prentice, I.C., 1992. Climate change and long-term vegetation dynamics. In: D.C. Glenn-Lewin, R.K. Peet and T.T. Veblen (Editors), Plant Succession. Theory and Prediction. Chapman and Hall, London, pp. 293–339. Radosevich, S.R., Holt, J. and Ghersa, C.M., 1997. Weed Ecology. Implications for Management. Wiley, New York, 589 pp. Sala, O.E., Oesterheld, M., Le´on, R.J.C. and Soriano, A., 1986. Grazing effects upon plant community structure in subhumid grasslands of Argentina. Vegetatio, 67: 27–32. Schulze, E.-D. and Mooney, H.A., 1993. Ecosystem function and biodiversity. In: E.-D. Schulze and H.A. Mooney (Editors), Biodiversity and Ecosystem Function. Springer Verlag, Berlin, pp. 497–510. Smith, J.L., Papendick, R.I., Bezdicek, D.F. and Lynch, J.M., 1992. Soil organic matter dynamics and crop residue management. In: F. Blaine Metting Jr (Editor), Soil Microbial Ecology. Marcel Dekker, pp. 65–94. Soriano, A., 1971. Aspectos r´ıtmicos o c´ıclicos del dinamismo de la comunidad vegetal. In: R.H. Mej´ıa and J.A. Moguilevski (Editors), Recientes Adelantos en Biolog´ıa. pp. 441–445.

´ Claudio M. GHERSA and Rolando J.C. LEON Soriano, A., 1991. R´ıo de la Plata grasslands. In: R.T. Coupland (Editor), Natural Grasslands. Ecosystems of the World 8. Elsevier, Amsterdam, pp. 367–407. Soule, J.A., 1993. Tagetes minuta: a potential new herb from South America. In: J. Janick and J.E. Simon (Editors), New Crops. Wiley, New York, pp. 649–653. Stowe, L.G. and Teeri, J.A., 1978. The geographic distribution of C4 species of the dicotyledoneae in relation to climate. Am. Nat., 112: 609–623. Su´arez, S.A., 1997. Comunidad de Malezas en la Pampa Ondulada como ´ındice de Biodiversidad y de Rendimiento de los Cultivos. MSc Thesis, University of Buenos Aires, 80 pp. Su´arez, S.A., Le´on, R.J.C., Ghersa, C.M. and Burkart, S., 1995. Cambios ﬂor´ısticos en las comunidades de maleza del maiz relacionados con el deterioro del ambiente. In: Actas Primeras Jornadas Cient´ıﬁcas del Medio Ambiente. Asociaci´on de Universidades Grupo Montevideo, Montevideo (Uruguay). Swift, M.J. and Anderson, J.M., 1992. Biodiversity and ecosystem function in agricultural systems. In: E.-D. Schulze and H. Mooney (Editors), Biodiversity and Ecosystem Function. Springer, Berlin, pp. 15–42. Swift, M.J., Vandermeer, J.H., Ramakrishnan, P.S., Anderson, J.M., Ong, C.K. and Hawkins, B.A., 1996. Biodiversity and agroecosystem function. In: H.A. Mooney, J. Hall Cushman, E.A. Medina, O.E. Sala and E.-D. Schulze (Editors), Functional Roles of Biodiversity. A Global Perspective. Wiley, New York, pp. 261–298. Tecchi, R.A., 1983. Contenido de silicoﬁtolitos en suelos del sector sudoriental de la pampa ondulada. Cienc. Suelo, 1: 75–82. Tilman, D., 1982. Resource Competition and Community Structure. Princeton University Press, Princeton, New Jersey, 296 pp. Tivy, J., 1990. Agricultural Ecology. Longman Scientiﬁc and Technical, New York, 288 pp. van Andel, J., Bakker, J.P. and Grootjans, A.P., 1993. Mechanisms of vegetation succession: a review of concepts and perspectives. Acta Bot. Neerl., 42: 413–433. Vitousek, P.M. and Walker, L.R., 1987. Colonization, succession and resource availability: ecosystem level interactions, In: A.J. Gray, M.J. Crawley and P.J. Edwards (Editors), Colonization, Succession and Stability. Blackwell Scientiﬁc Publications, London, pp. 207– 224. White, P.S., 1979. Pattern, process, and natural disturbance in vegetation. Bot. Rev., 45: 229–299. Whittaker, R.H., 1975. Communities and Ecosystems, 2nd Edition. MacMillian, New York, 203 pp. Whittaker, R.H., 1977. Evolution of species diversity in land communities. Evol. Biol., 10: 1–67. Williamson, M., 1996. Biological Invasions. Chapman and Hall, London, 256 pp. Williamson, M.H., 1987. Are communities ever stable? In: A.J. Gray, M.J. Crawley and P.J. Edwards (Editors), Colonization, Succession and Stability. Blackwell Scientiﬁc Publications, London, pp. 353– 372. Wink, M., 1993. Production and application of phytochemicals from an agricultural perspective. In: T.A. van Beek and H. Breteler (Editors), Phytochemistry and Agriculture. Clarendon Press, London, pp. 171–213.

Chapter 21

PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND R.E. SOJKA

INTRODUCTION

Humanity’s presence on earth has forced the selective adoption of both anthropocentric and naturalistic perspectives of soil as an ecosystem component. From the anthropocentric perspective, soil is an ecosystem component used by humans for speciﬁc purposes (e.g., to grow forests and crops; support structures or roadways; and as a ﬁltration medium). The naturalistic perspective sees soil primarily as the natural foundation or backdrop for other ecological systems and processes, and philosophically excludes many soil-management technologies and scenarios, favoring only soil uses and management practices that derive from natural ecosystem processes. The naturalistic perspective is more willing to concede that soil, like other ecosystem elements, may at times respond to perturbations counter to human needs and aesthetics. The role of environmental managers and scientists is to know when and how ﬁrmly to embrace the validity of either or both outlooks. That requires an appreciation of the properties of ecosystem components, and how those properties affect a given management objective. Familiarity with fundamental soil properties is essential to understanding the physical aspects of soils of disturbed ground, regardless of the interpreter’s perspective. This chapter presents a summary of essential soilscience concepts necessary to begin understanding the interactive role of soil in a disturbed ecosystem. The emphasis is on soil physical properties and processes. However, soil is a biologically and chemically dynamic system with strong interactions, interdependencies, and feedback among all its compartments, phases, and functions. Thus, some fundamental chemical and biological concepts relevant to soil physical status are also brieﬂy outlined. The framework of fundamental

concepts is used to explain the role of soil physical status in several important kinds of land disturbance. A detailed analysis of all aspects of soil physical perturbation from all conceivable kinds of physical land disturbance is beyond the scope of the chapter and the expertise of the author. But application of principles to several key types of ecological disruption in which soil physical disturbance is important provide a conceptual framework that can be extended to other scenarios.

THREE-PHASE SOIL MODEL

The essential physical aspects of soil are often represented by a simple three-phase model (Fig. 21.1). The three phases are solid, liquid, and gas. The proportion, arrangement, and constitution of each phase dictates soil properties and functionality within a given ecosystem or for a given use. Typically, and perhaps surprisingly, half the volume of soil beneath one’s feet is composed of air and water.

Fig. 21.1. The three-phase conceptual model of soil, showing typical liquid, gas, and solid composition, including the distribution of mineral vs. organic solids in a productive soil from the United States “corn belt”.

503

504

R.E. SOJKA

Table 21.1 Limits of soil size separates for various classiﬁcation schemes System 1

Particle size range (mm) Very coarse sand

USDA-NRCS

2.0–1.0

Coarse sand

Medium sand

1.0–0.5

0.5–0.25

ISSS

2.0–0.2

DIN, BSI, MIT

2.0–0.6

ASTM

2.0–0.42

Corp, Bureau

4.76–2.0

Highway

2.0–0.42

0.6–0.2 2.0–0.42

Fine sand

Very ﬁne sand

0.25–0.1

0.1–0.05

Silt

Clay

0.05–0.002

<0.002

0.2–0.02

0.02–0.002

<0.002

0.2–0.06

0.06–0.002

<0.002

0.42–0.074

0.074–0.005

<0.005

0.075–0.002

<0.002

0.42–0.074 0.42–0.075

<0.074

1 System: USDA-NRCS, United States Department of Agriculture, Natural Resources Conservation Service; ISSS, International Soil Science Society; DIN, German Standards; BSI, British Standards Institute; MIT, Massachusetts Institute of Technology; ASTM, American Society for Testing and Materials; Corp, United States Army Corps of Engineers; Bureau, United States Department of Interior, Bureau of Reclamation; Highway, American Association of State Highway and Transportation Ofﬁcials.

Human use and natural processes alter the physical and chemical aspects of all three phases. While each phase may be altered somewhat independently, the effects are nearly always the result of interactions among the phases. The physical aspects of soil behavior on disturbed ground are affected foremost by changes in the arrangement and proportion of the phases, especially with respect to the amount and arrangement of the pores which hold and conduct soil gases and liquids. Chemical changes in one or more of the phases, however, can also bring about signiﬁcant physical effects. Solid phase Solid-phase properties are affected by the proportion of particles of various sizes (texture), their arrangement (structure), their mineralogy and organic-matter content, the composition of ions and other chemical constituents adsorbed on their surfaces or ﬁlling the interstices, and the degree of hydration of the system. Disturbance of any one of these aspects can affect a given system component singly, but usually also causes cascading effects within the three-phase model. Unlike many other soil properties, texture is regarded as nearly unchangeable in all but the most drastic of soil alterations (Soil Survey Staff, 1993). In the United States, there are several classiﬁcation schemes for soil textures (Table 21.1). These can sometimes come into conﬂict, because soils and landscapes are usually classiﬁed and mapped under one scheme, but, if not for farming, are often managed or manipulated using different standards. In the United States, for

example, the Natural Resources Conservation Service textural standards (Soil Survey Staff, 1993) are used to map soils, but one of several engineering standards (Table 21.1) might be used for engineering or construction purposes. Furthermore, when soils are mapped, the mapping unit texture is based on the surface diagnostic horizon (topsoil layer). The texture of underlying horizons may vary greatly. Table 21.1 presents the size range for soil separates, as classiﬁed by several systems. The ostensible immutability of soil texture derives from basing its assessment solely on the proportional composition of size separates of the mineral fraction. Texture analysis excludes organic material, which is oxidized before determination of the size separates, and is performed on the remaining mineral material which is ﬁrst entirely dispersed into individual (primary) particles – that is, the non-aggregated or non-structured mineral fraction (Gee and Bauder, 1986). Except as the result of catastrophic natural events or anthropogenic intervention, the proportion of these constituents in soil is very stable, because these soil mineral constituents change size (weather or accrete) very slowly – over many decades or longer (Lyles and Tatarko, 1986). Similarly, measurable in-situ vertical movement of soil separates from one horizon to another (with water and gravity) occurs only very slowly. This occurs through the loss of ﬁnely dispersed or dissolved solid material from one soil horizon (eluviation) and the deposition of the material in another horizon (illuviation). On a landscape basis, similar exchange of materials is only slightly faster, through the activity of biota such as worms, ants and termites (see discussion

PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND

below, p. 509). Changes in the proportion of mineral size separates can be accelerated by mass displacement, either in-situ (e.g., through tillage-associated mixing of the surface horizons) or by displacement across the landscape (e.g., by erosion and deposition, landslides, or anthropogenic earth-moving). The textural classiﬁcation scheme most commonly used in the United States is represented in the NRCS soil textural triangle (Fig. 21.2). Texture does not change by addition of organic matter, or through any manipulation of soil that does not add or remove a speciﬁc mineral size fraction. The perception of textural change, by many unfamiliar with the technical deﬁnition of texture, usually depends on change in soil structure, particularly aggregation (structural units, typically a few millimeters in size), and is often recognized as a difference in “tilth” or “friability”. Both of these terms are non-speciﬁc, but refer in a general sense to improved stability of soil structure, ease of soil penetration by roots, gases, or inﬁltrating ﬂuids, and a resistance to “slaking”, or rapid loss of structure upon wetting or mechanical disturbance. The presence of durable aggregates can make coarse, so-called light-textured (sandy) soil or ﬁne, so-called heavy-textured (clay) soil, feel and behave more like a medium-textured (loamy) soil, without actually changing the proportion of mineral size separates (Russell, 1976). The colloquial use of “light” vs. “heavy” as textural terms derives from the wet weight of soils, the feature commonly encountered by farmers in the ﬁeld. Since clays hold more water per unit volume, they are heavier when wet than are sands. Interestingly, the opposite is generally true when the soils are oven-dried to 105ºC to remove all water. The presence of some organic matter and clay, which act as particle binders and adhesives, favors formation of the kind of aggregates generally associated with tilth and ease of rooting. Aggregation and structure are more prevalent in medium-textured soils, such as loams (see Fig. 21.2), which have a balanced mixture of particle sizes, than in soils heavily dominated by a single size fraction, such as clays or sands. The thoroughly decomposed (stable) organic matter of soil is one of the most potent binding agents for soil aggregates. Aggregate stability tends to be highly correlated with organic matter content (Chaney and Swift, 1984; Soane, 1990). The desirability of organic matter and aggregation depend upon the intended soil use (see below, p. 511).

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Fig. 21.2. Soil textural triangle, showing limits of sand, silt, and clay size separates composing the various textural classes recognized by the United States Natural Resources Conservation Service.

Detailed soil structural classiﬁcation and terminology, though beyond the scope of this chapter, are explained in a variety of sources (Soil Survey Staff, 1975; Russell, 1976). The key consideration is that the ordered arrangement of primary soil particles into structural units alters the relationship of solid particles and interstices to one another. This spatial reorganization of the three-phase system affects mechanical behavior and strength, transport of ﬂuids and gases through the soil, and retention of water in the bulk soil proﬁle and within structural units. These properties in turn inﬂuence the behavior of soil organisms and oxidation-reduction chemistry. Mineralogy of the solid phase, particularly of the clay size fraction, signiﬁcantly inﬂuences soil physical characteristics and behavior, as well as greatly inﬂuencing the chemistry and biotic properties of the soil. Again, the subject of clay-mineral effects on soil physical properties is extensive and is only brieﬂy outlined in this chapter. Clay mineralogy is determined by the composition of secondary layer-silicate minerals, sometimes referred to as clay minerals or 1:1 and 2:1 layer silicates. In general, clay minerals are broadly grouped into classes determined by the ratio of silica layers (joined tetrahedral structures forming a thin mineral sheet) to alumina layers (joined octahedral structures forming a thin mineral sheet) in their crystalline structures. Isomorphic substitutions of elements displacing silicon or aluminum within the tetrahedra and octahedra result

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in a negative charge which varies among the mineral species. The 1:1 layer-silicates have a single tetrahedral layer and a single octahedral layer, and the degree of substitutions is usually low. Thus, the charges of these minerals are only slight. Kaolinite is a representative 1:1 clay species. Because the charge, quantiﬁed as cation exchange capacity (centimoles of charge per kilogram of solid), is low (3–15 cmoles kg−1 ), these minerals are relatively inactive in retaining cations (e.g., the plant nutrient ions Mg++ , K+ , or NH+4 ). Soils dominated by kaolinitic clays have low mineral fertility. These clays are also less effective in aggregate formation, and retain relatively stable physical conﬁguration regardless of hydration or cation composition on the exchange sites. The 2:1 layer silicates have two tetrahedral layers, oriented basal side to basal side, resting on top of a single alumina layer. Isomorphic substitutions occur more commonly, and thus cation exchange capacity is higher (up to 100 cmoles kg−1 ). Soils dominated by 2:1 clays tend to have greater mineral fertility than soils dominated by 1:1 clays. The 2:1 clays are also more active physically than the 1:1 clays, playing a larger role in joining soil primary particles to form aggregates (Hagin and Bodman, 1954). Because of the weak repulsion of adjacent 2:1 clay crystal laminae, water molecules easily invade the space between laminae. Thus, as hydration increases, 2:1 clays swell, earning them the common designation of “shrink–swell clays.” As the degree of isomorphic substitution increases and cation exchange capacity rises, 2:1 minerals are more inﬂuenced by hydration. As pure clays hydrate from an air-dry state, volume changes of a few percent for kaolinite occur, compared to 30–60% expansion for the expanding-lattice 2:1 clay mineral montmorillonite, depending upon the ions dominating exchange sites (Hillel, 1980). In soils which have mixed mineralogy and contain organic matter and other non-clay mineral solids, expansion ﬁgures of several percent for a given mass of soil are common if the soils are high in montmorillonitic clay. These changes can greatly affect construction of roads, foundations, or other structures, as a result of the large deep cracks that form when the soil is dry, and strong internal pressures when the soil is wet. The composition of cations adsorbed on clay surfaces and dissolved in soil water has physical consequences for soils with high clay content (Brooks et al., 1956; Auerswald et al., 1996). When exchange sites are

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dominated by sodium ions, which have large hydrated radii compared to divalent calcium or magnesium cations, the clay laminae become separated. This causes the ﬁne solids in soil, especially the clay material, to disperse, degrading or preventing development of stable aggregates (Velasco-Molina et al., 1971). These ﬁne dispersed solids (largely clay) are susceptible to movement and rearrangement with water, usually resulting in blockage of soil pores, which reduces inﬁltration and drainage (Shainberg and Singer, 1985; Shainberg et al., 1992). Pore blockage near the surface can accelerate runoff and lead to erosion (see p. 513). When the soil is wet, blocked pores in surface crusts have been shown to hamper soil aeration (see p. 512) by impeding diffusion of oxygen into internal pore spaces (Sale, 1964; Miller and Gifford, 1974). Organic matter is a very important component of the solid phase, having consequences with respect to physical and chemical properties greatly disproportionate to its relatively small proportion in the soil mass. Even minor changes in the small amounts of organic matter present in soils can result in signiﬁcant changes in soil properties. Soil organic-matter content is especially important for maintenance of soil structure, particularly soil aggregation, the crumb-like structure of soil that so greatly inﬂuences such soil properties as aeration, water inﬁltration, and resistance to erosion. These soil properties are exceedingly important at the scale of phenomena associated with tillage, perhaps the single most pervasive form of soil disruption on the planet (see pp. 510–511). Increased aggregate stability with increased soil organic-matter content has been noted essentially across soils of all textures and mineralogy by many researchers (Martin, 1945; Tisdall and Oades, 1982; Chaney and Swift, 1984; Burns and Davies, 1986) with few exceptions (Panayiotopoulos and Kostopoulou, 1989). Soil organic matter derived from decomposition of grass roots seems especially effective in promoting aggregation (Ekwue, 1990). Higher organic-matter content increases resistance to aggregate breakdown from a variety of forces, ranging from freezing and thawing (Lehrsch et al., 1991) to tillage (Tisdall et al., 1978). Furthermore, organic-matter turnover and aggregate stability are intimately tied to the microbial ecology of the upper soil proﬁle (Harris et al., 1964; Hepper, 1975; Tisdall et al., 1978; Dommergues et al., 1979; Burns and Davies, 1986). The speciﬁc organic fraction most potent in promoting and preserving soil aggregation has been broadly identiﬁed as a class

PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND

of polysaccharides of high molecular weight (Molope et al., 1985; Metzger et al., 1987; Roberson et al., 1991). Liquid phase The water contained in a soil greatly inﬂuences its physical properties and its ability to sustain plants, microbes, and soil mesofauna (e.g., ants, termites and worms). Soil water retention is quantiﬁed in terms of the amount of water per unit weight or per unit volume of soil that is held against a given free-energy gradient. This is usually expressed in terms of water content at a speciﬁc value of tension, “negative” pressure, suction, or matric potential. Plant-available water is sometimes deﬁned as the amount of water held between saturation (zero potential) and ~1.5 MPa of soil matric potential. In high-salinity soils, the osmotic component of soil water must be added to the matric potential to assess adequately a plant’s ability to utilize the water present. At zero potential soil is saturated, and water can drain freely in response to gravity. At ~1.5 MPa potential it would take an atmospheric pressure of 1.5 Mpa in a conﬁned system to drive water from the soil matrix. Water-retention characteristics are affected by soil texture (Ehlers et al., 1995) and structure (Taylor and Box, 1961). Coarse-textured soils (sands) hold relatively little water compared to ﬁne-textured soils (clays) (Richards, 1959). As mean soil pore size decreases, water retention generally increases (Donat, 1937). Compaction or aggregation thus affect water retention through their effects on pore-size distribution. The sufﬁciency of the water for the needs of soil biota and plants is dependent upon what fraction of the volume of water retained can be extracted from soil between about 0 and ~1.5 MPa of negative pressure. The sufﬁciency of this amount of water (often referred to as the available water holding capacity) within an ecosystem is further determined by the desiccating strength of the climate or microclimate and the ability of the organisms to regulate water loss. Soil disturbance often affects ecosystems by altering soil water-holding capacity, either by changing the water-retention characteristics of the soil, by decreasing soil depth (and hence water-storage volume), or by compacting the soil to strength levels that roots cannot penetrate, reducing rooting volume and de facto water storage. Water content affects several solid-phase properties. Soil strength (hardness or penetration resistance) is

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a function of water content. Soil strength decreases rapidly from a plateau value at desiccation, and reaches an asymptotic minimum value determined by solid-phase characteristics as soil becomes wet (Camp and Gill, 1969; Mirreh and Ketcheson, 1972; Ayers and Perumpral, 1982; Gerard et al., 1982; Campbell et al., 1988). As water content increases and mechanical strength decreases, soil is more easily deformed or compacted. Thus, compaction of soils, whether intentional or not, occurs more easily when trafﬁc or other applied stress is imposed while soil is wet (see p. 510). Texture also affects the relationship between soil water content and soil strength (Spivey et al., 1986). At medium textures, soils that have been compacted when wet often have a very low porosity. This is because under compression the various size fractions are arranged to almost completely ﬁll the interstices between soil particles (Campbell et al., 1988). Water content, rapidity of wetting, and length of time without disturbance affect the durability of soil structure (Blake and Gilman, 1970; Arya and Blake, 1972; Utomo and Dexter, 1981; Kemper and Rosenau, 1984). Rapid wetting of dry soil is highly destructive of exposed aggregates and surface structure, resulting in loss of surface porosity and formation of surface seals which impede inﬁltration and increase runoff (Segeren and Trout, 1991), contributing to erosion (see pp. 513– 515). The salinity of water that is in the soil or is being applied to soil can affect solid-phase relationships, depending on total salinity and the relative amount of sodium in the water (United States Salinity Laboratory Staff, 1954). Saline water low in sodium tends to preserve soil structure, whereas water high in sodium relative to other cations (especially if total salinity is low) is destructive to soil structure and increases soil erodibility (Le Bissonais and Singer, 1993; Lentz et al., 1996; see also p. 513). Gas phase The proportions of solid, liquid, and gas are dependent on the magnitude of porosity and the extent to which pores are ﬁlled with water, and determine the status of soil aeration. Soil aeration can be described in terms of capacity, intensity, or rate factors (Stolzy and Sojka, 1984). Capacity infers soil oxygen status from the relative volume of gas space in the three-phase system. In a number of soils, oxygen availability is

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adequate for plant growth when air-ﬁlled pore space or the concentration of oxygen equals or exceeds 10% (Wesseling and van Wijk, 1957; Anderson and Kemper, 1964; Grable and Siemer, 1968; Wesseling, 1974). The composition of the atmosphere near the earth’s surface is 21% oxygen, 78% nitrogen, and, as of 1993 (Keeling and Worf, 1994), about 0.035% carbon dioxide. The balance is composed of various trace gases. In soil air, oxygen depletion through respiration lowers the amount of oxygen present, and raises by an equivalent amount the content of carbon dioxide and trace organic gases such as methane and ethylene, which are byproducts of anaerobic respiration. Thus, oxygen and carbon dioxide levels of 10–12% in soil are common. Carbon dioxide levels can approach 20% and oxygen levels can fall to virtually 0% in extreme circumstances (Russell and Appleyard, 1915). Therefore, merely determining the volume of soil air gives an incomplete picture of aeration. Determining the intensity factor (concentration or partial pressures of gases) in addition to capacity is an improvement. However, chemical reactions in soil, and biological processes of soil micro- and mesobiota and the roots of higher plants, experience or utilize soil aeration as a rate factor. They depend on the rate at which oxygen can be exchanged at a speciﬁc microsite in the soil relative to the required rate of oxygen use (Letey and Stolzy, 1967). Soil air exchanges oxygen within the soil proﬁle and with the atmosphere through a variety of processes including surface turbulence, variation in barometric pressure and changes in soil temperature, and by physical displacement by and dissolution from inﬁltrating water. The most active mechanism of oxygen exchange, however, is by diffusion from the ambient atmosphere (Russell, 1952). Oxygen diffuses through air 104 times faster than through water (Greenwood, 1961) and only one-fourth as rapidly through dense protoplasm as through water (Krogh, 1919; Warburg and Kubowitz, 1929). Thus, the hydration of the soil system and the organisms active within it profoundly inﬂuence the system’s ability to supply oxygen at rates necessary for oxygen-dependent reactions, such as aerobic respiration, or to prevent reduction of compounds or elements such as iron, which can produce phytotoxins when the oxygen diffusion rate is low. MAJOR SOIL DISTURBANCE CATEGORIES

All forms of soil disturbance draw their consequences

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from the same collection of soil processes described as parts of the three-phase soil model. However, the systematics of each disturbance scenario, that is, the extent of such processes and their interaction with other processes differ from one kind of disturbance to another. Sometimes these systematics differ only incrementally and subtly, sometimes wholly and dramatically. Furthermore, the scale and intensity of certain kinds of disturbance can vary greatly, with obvious implications for evaluation or management of the particular disturbance. Thus, it is impossible to cover thoroughly the considerations with respect to soil physics for every conceivable category of disturbance. Based upon extent of phenomena and annual impact I have selected three speciﬁc categories of disturbance for more detailed consideration. I have called these categories Loosening and Compaction, Flooding and Erosion. Loosening and Compaction is almost entirely anthropogenic, particularly as the result of agricultural trafﬁc and tillage. Flooding is almost entirely nonanthropogenic. Erosion is an intimate combination of anthropogenic and natural disturbance factors. Other categories can certainly be identiﬁed. Some may be more visible to the general public (which is mainly urban and suburban), as is the case with construction disturbance (considered brieﬂy under Loosening and Compaction – p. 511). Some may be more intense and noticeable in their effects on the landscape within their contained areas of inﬂuence, as with mining. Yet, compared to Loosening and Compaction, Flooding, and Erosion, the global extent of impacts of these other categories on ecosystems is much less. Furthermore, many of these less extensive categories of soil physical disturbance are strongly analogous to the major categories mentioned. An understanding of their systematics can be derived from applications of principles from the three-phase soil model, and sometimes, for instance in the case of mining, with the overlay of potent impacts caused by changes in environmental chemistry. Taking the United States as an example, the total land area of the country is 917 063 560 ha. Paone et al. (1978) estimated that agriculture was by far the most extensive source of land disturbance in the United States, accounting for about 515×106 ha (56%) of the nation’s land area. About 191×106 ha were in cropland and 245×106 ha were in pasture, range, and grassland. An additional 80×106 ha were designated as agricultural land, but were composed of forested

PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND

land used for grazing. These ﬁgures compared to a total of only 1.5×106 ha (0.2%) of land disturbed for mining during the period from 1930 to 1971, half of which had been reclaimed by 1978. By way of further contrast, farm roads and farmsteads accounted for 4.6×106 ha (0.5%) and ungrazed forested land totalled 213×106 ha (23%). Recent estimates would indicate a 2% decline over the years in the combined non-forested area farmed and grazed. This was despite maintaining or improving the food supply of a large increase in population and expanding agricultural exports. The more recent area totals are 162×106 ha in all croplands excluding pasture (but including idle land), 266×106 ha in pasture, range, and grassland, and 262×106 ha in forested land (Hunst and Powers, 1993). The more recent estimates for forested lands include areas used for grazing, which prevents direct comparison with the earlier ﬁgures for forested and grazing lands. While agricultural areas in the United States are extensive, they are also remarkably productive, with much of the output going to export. The proportion of agriculture in the landscapes of the world is generally similar to that in the United States or greater, though the production efﬁciency is often much less (see Giampietro, Chapter 32, this volume). Production efﬁciency is improving rapidly in less developed countries, however, with the successful adoption of high-output agricultural technology. This latter point is especially important if viewed in terms of the role of high-output agriculture in preserving earth’s natural ecology. Waggoner (1996) studied the effect of improved agricultural technology on land use in India between 1966 and 1994. About 13 million hectares were devoted to the production of 11 million tons of wheat. In 1994 about 24 million hectares were cultivated to produce 57 million tons of wheat. Had the low technology of the sixties still been used, the land requirement for the same 57 million tons would have been 69 million hectares. In a similar analysis, Avery (1997) estimated that land spared from agricultural development worldwide since 1960 by adoption of advanced farming methods “is equal to the total land area of the United States, Europe, and Brazil.” There are just under 1.5×109 ha of land in arable crop production in the world (Higgins et al., 1988), not counting pasture, range, and grasslands. Most of this cropland receives tillage at least once a year, as well as extensive wheeled trafﬁc for other operations;

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hence the importance of loosening and compaction as a disturbance category. Loosening and compaction Loosening and compaction affect the arrangement of solids and pores, with secondary effects on soil properties that depend on these relationships. Some mixing of soil horizons during loosening events or operations may inﬂuence both physical and chemical properties (Campbell et al., 1988; Chapman, 1990; Sumner, 1995). Soil loosening and compaction rarely occur rapidly on a large scale through natural processes, but they occur quite commonly in agriculture, surface mining, and construction. A typical hectare of soil contains about 2.3×106 kg of soil in the surface 15 cm of depth. This entire mass can be inverted and mixed in a few hours by mechanized tillage. In nature, loosening of an equivalent mass by earthworms and burrowing animals occurs on landscape scales, but at rates that are typically apparent in their cumulative effects only over decades or longer. Individual species of earthworms can ingest from 10 t ha−1 yr−1 of soil (for the species Aporrectodea rosea) to 500 t ha−1 yr−1 (for the species Millsonia anomala). A group of species including M. anomala occurring together are capable of combined consumption rates of 1200 t ha−1 yr−1 (Lee and Smettem, 1995). If all ingested material were deposited on the soil surface, these ﬁgures would correspond with a complete turnover of the top 15 cm of soil in periods ranging from 230 to 1.9 years. However, incomplete soil displacement, inconsistent rates of ingestion in time and space (caused by variations of temperature, moisture substrate, etc.), and mixedspecies and variations in ﬁeld populations, reduce mean soil mixing rates, as measured by deposits on the surface. Lee and Smettem (1995) also summarized data that would suggest that the contribution to soil loosening by other mesofauna, such as termites and ants, is one to two orders of magnitude less than for earthworms. Mixing by earthworms and other burrowing mesoand macrofauna is far less intense than tillage, often does not penetrate as deeply, and is incapable of penetrating some hard subsoil zones that are readily disrupted by mechanized deep tillage. This is an important consideration in agriculture for management of persistent root-restrictive hardpans. Earthworm activity was shown unable to ameliorate hardpans effectively,

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even after decades and despite organic enrichment of surface horizons to increase earthworm populations (Horn, 1986). Earth slippage is one example of intense natural loosening. Trampling under the hooves of migratory herd animals is an example of natural compaction. However, since the degree of compaction is related to the load and extent of contact of the load-transferring surface, trampling does not result in the extensive or intense compaction often seen as a result of human activity. Soil compaction occurs primarily as the result of trafﬁc or other forces imposed from near the surface, or immediately below the depth of tillage; these forces are transmitted into the soil, causing compression well below the depth of direct wheel or implement contact (Horn, 1995). Persistent layers of compaction are often called pans (e.g., trafﬁc pans, tillage pans, or hardpans). Subsoil compaction pans can last a decade or more, even in the presence of annual freezing and thawing (Blake et al., 1976; Voorhees et al., 1978). In some soils, hard pans are formed in situ through natural forces of soil consolidation, usually under high rainfall, and sometimes as a result of illuviation of materials from overlying horizons ﬁlling pores and interstices in the layer that eventually becomes the genetic hardpan. Agricultural tillage is one of the most extensive forms of soil disturbance. Tillage is commonly performed to kill weeds, bury residues, reduce soil strength and/or ameliorate subsoil rooting restrictions, accelerate drying and warming, improve aeration and/or inﬁltration, incorporate fertilizer and/or pesticides, and provide seedbed preparation, levelling, and drainage. When the surface is tilled, there is an initial increase in the porosity of the surface and inﬁltration rates are higher. However, prolonged surface tillage disrupts the continuity between surface and subsurface pores in the vicinity of the “plow sole”, the subsoil layer supporting the weight of moving tillage implements (Douglas and Goss, 1987). The greater porosity and aeration following tillage temporarily accelerates microbial activity (Carter, 1991). This can be seen in the immediate increase in release of carbon dioxide (Reicosky and Lindstrom, 1993, 1995), and in the long-term decline of soil organic matter (Bauer and Black, 1981, 1983; Dalal and Mayer, 1986; Rasmussen and Collins, 1991; Ehlers and Claupein, 1994) and decrease in aggregate stability (Schønning and Rasmussen, 1989). Following many years of conventional tillage, surface soils tend to

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have higher bulk density after consolidation and trafﬁc than untilled soils (Bauer and Black, 1981; Ehlers and Claupein, 1994). This is caused by the loss of resistance to aggregate breakdown, compactibility, and compressibility accompanying lowered soil organic matter levels (Zhang and Hartge, 1995). Subsoiling (Sojka, 1995) is performed speciﬁcally to loosen soil at depths of ~0.30–0.45 m, that is, through the normal depth of crop rooting. Soil loosening (natural or imposed) is generally a transitory effect. The duration of the effect depends upon the organicmatter content and texture of the soil, the frequency and intensity of rainfall or irrigation, the amount of subsequent trafﬁc over the surface, and the water content of the soil at the time of the trafﬁc (Busscher et al., 1986; Busscher and Sojka, 1987; Sojka et al., 1990, 1991, 1997). The duration of loosening from agricultural subsoiling has been extensively studied. The various soil properties affected by loosening often persist for different lengths of time. Depending on soil properties, water content at the time of subsoiling, climate, and indicator crop, subsoiling effects have been found to persist from 1 to 5 years, but typically only 1–2 years (Lindner, 1974; Schulte, 1974; Swain, 1975; Bokerman and Graichen, 1981; Hartge, 1981; Threadgill 1982; Jager and Boersma, 1983; Kouwenhoven and Vulinck, 1983; Martinovic et al., 1983; Schulte-Karring, 1983; Busscher et al., 1986; Ide et al., 1987; Simmons and Cassel, 1989; Chapman, 1990). However, the effect generally only lasts for about one year in sandy soils, and two years on medium- to ﬁne-textured soils. The persistence of the disruption effect is shortened in environments with high rainfall. The effect can seldom be detected by measurement of bulk density for longer than one year, or by penetration resistance for longer than two years. Plant response to disruption is sometimes measurable for as long as ﬁve years in siltor clay-textured soils, and all measures of disruption tend to show longer duration of the effect if the soils were relatively dry at the time of disruption. For plants sensitive to compaction, growth of total phytomass and harvested yield (usually of seeds) are often highly correlated with the extent of soil compaction (Carter et al., 1965; Carter and Tavernetti, 1968; Carter, 1990). Sojka et al. (1991) showed a linear yield response of corn (Zea mays) to soil strength in a subsoiling study comparing disrupted and undisturbed soil proﬁles (Fig. 21.3).

PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND

Fig. 21.3. The dependence of yield of corn (Zea mays) on proﬁle soil strength, as measured with a standard ASAE cone penetrometer (both averaged over two years), in a two-dimensional grid across corn rows, showing effects of varying extent of subsoil disruption from three subsoil-loosening implements (from Sojka et al., 1991). Data are from two cropping systems: Conventional Tillage (Conv), soil surface-tilled bare of residue before planting; and Minimum Tillage (Min), soil surface-tilled to a shallow depth before planting, with a substantial amount of loose residue left unincorporated or partially incorporated on the soil surface. Three 45 cm deep, noninverting subsoiling implements were compared: a straight-shanked subsoiler (SS) with 13 cm wide ripping surfaces, a parabolic-shanked subsoiler (KE) with 7.5 cm wide ripping surfaces, and a subsoil lifting surface (termed a Paratill; PT) capable of extensive lateral soil disruption.

Elkins and Hendrick (1983) and Karlen et al. (1991) demonstrated that an entire mass of soil need not be disrupted to greatly improve plant growth on ﬁelds with root-restrictive subsoils. When small root-sized channels were made to penetrate restrictive layers in the subsoil, and the openings were quickly stabilized with thick, persistent plant roots, considerable yield improvement was realized. If plant establishment could be achieved with minimal disruption, the requirements for fuel and horsepower to ameliorate compaction could be much reduced. Busscher et al. (1988) demonstrated that there was a considerable range in the thoroughness and pattern of subsoil disruption among subsoiling implements commonly available. Some deep tillage is done more for chemical remediation than for the direct physical loosening. Sandoval et al. (1972) demonstrated that certain dispersed-clay sodic soils could be reclaimed using deep plowing tillage that mixed large amounts of calcium carbonate from lower horizons with highsodium soil of the affected upper horizons. The calcium reduced the percentage of exchangeable sodium in the soil, enhanced ﬂocculation and stabilized aggregation,

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thus improving the physical properties of the soil (Sandoval and Jacober, 1977; Sandoval, 1978). The type of soil structure desired by engineers for roadways or foundations is very different from that needed for plant growth (Zhang and Hartge, 1995). These two uses manipulate the three-phase soil model to achieve nearly opposite outcomes. Organic matter promotes and stabilizes aggregation and porosity while reducing soil compressibility (minimum void ratio achievable as stress increases) and compactibility (maximum achievable bulk density). Roadways and foundations require densely compacted soils to prevent consolidation and settling after construction. Where the local soils are high in swelling clays, compaction lowers porosity, inhibits water entry into the subsoil, and thereby reduces the ﬂuctuations in water content that disrupt foundations and roadways. For these reasons, construction procedures often remove topsoil to minimize the inﬂuence of organic matter, and intentionally work to reduce porosity, increase runoff, and reduce the ease of rooting in soils supporting the roadways or structures. Compaction effects associated with construction are often criticized because, until the construction is completed, these processes can promote erosion from an unprotected work site. New technologies (Roa, 1996) and environmental law have begun to reduce offsite sediment problems. Most construction operations now seek rapidly to revegetate work sites immediately upon completion of the construction, both for aesthetic reasons, and to protect the structures and roadways from being undermined by accelerated runoff and erosion from the low-permeability soils that have been created. Thus, in the revegetation phase a balance is struck to enhance surface-soil properties sufﬁciently to support sod, or other non-intrusive vegetative cover, without undermining the intensive earthwork that must remain undisturbed to support the structures. Flooding Perhaps the most extensive, frequent, and devastating form of natural land disturbance is ﬂooding (Kozlowski, 1984a), affecting tens of millions of hectares annually on a global basis. Flooding occurs in large regional events and in discrete isolated events, many of which are too small to monitor and estimate systematically. The areal extent of annual spring ﬂooding in the state of Mississippi (U.S.A.) alone is estimated at 1.6×106 ha (Kennedy, 1970). In the extensive ﬂood of

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1993 in the midwestern United States, it was estimated that 8×106 ha were damaged by ﬂoodwaters, and at one time 2.8×106 ha were under water. In addition to ﬂooding, Dudal (1976) estimated that 12% of the world’s soil resource has excess water on a sustained basis. Flooding can wreak its damage swiftly, particularly if ﬂows are large and energetic, through aeration, disease effects, and erosional and alluvial processes. But even placid water or rising water tables can inﬂict a severe toll in less than 48 hours (Stolzy and Sojka, 1984). Some ﬂooding is predictable and controllable, even if unavoidable. Some is unpredictable, and characteristically all the more devastating. In any case, ﬂooding is, more often than not, a primarily natural phenomenon – one of the most potent forces in natural ecosystems. Humans are particularly affected by ﬂooding because of their proclivity to live and grow crops in ﬂoodplains. This tendency is prompted by the ease of property development on ﬂat land, and proximity to rivers for water and transportation, as well as proximity to the land itself, which is usually highly productive farmland. Flooding damage to soil often goes unnoticed because of the degree of destruction above the soil surface, both to the works of humankind and to the natural landscape. Flooding can cause settling and disruption of earthworks, roadbeds, and foundations. The extreme reduction in soil strength, when coupled with high winds, often results in uprooting of trees, and the toppling of powerlines and other upright structures. Flooding can have direct or indirect effects on soil systems. Direct effects include soil cooling, interference with soil aeration, degradation of soil structure, accelerated consolidation, erosion, and leaching of nutrients. Indirect effects include proliferation and carriage of soil-borne plant pathogens (Stolzy and Sojka, 1984), and if soils are inundated long enough, soil reducing conditions will eventually lead to denitriﬁcation, and the production of phytotoxic chemicals (Ponnamperuma, 1984). Many ﬂoods occur in temperate climates in the spring, when rivers swell with snow-melt and cold spring rains. Soils in the spring are usually still cool, and ﬂoodwaters can further delay spring warming by raising the heat capacity, and lowering the thermal diffusivity during periods of reduced incoming radiation. Bonneau (1982) reported a reduction in the temperature of ﬂooded soil by 6ºC compared to well-drained soil.

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This is enough to hamper the absorption by roots of some nutrients, particularly phosphorus, which is highly responsive to soil and root temperature in some crops, such as corn (Zea mays). Uptake of phosphorus and other nutrients is directly inhibited by soil hypoxia; these effects were comprehensively reviewed by Glinski and Stepniewski (1985). One of the most consistent effects of soil hypoxia and ﬂooding is potassium deﬁciency. Uptake of the potassium ion by plant roots stops immediately when oxygen diffusion to the root zone is impaired. This deﬁciency is interesting because of its possible link to stomatal closure in a range of higher plants when soil oxygen diffusion rate drops below 20×10−8 g cm−1 min−1 (Sojka et al., 1975; Sojka and Stolzy, 1980; Sojka, 1985). As explained earlier, oxygen diffusion through water is much slower than in the gaseous phase. Thus, oxygenation of the root zone depends on mass ﬂow of water through the proﬁle. This mass ﬂow is often impeded by swelling of soil and the dispersion of ﬁne materials blocking soil pores (Wickham and Singh, 1978). Ponnamperuma (1984) made an insightful and concise analysis of the effects of ﬂoodwater in the soil proﬁle. Reviewing the work of Grable (1966), Greenland (1981), and Houng (1981), Ponnamperuma further noted that 1.5 m per day of water movement through the soil proﬁle is needed to meet root oxygen requirements and that at least 1 cm per day of water movement is needed to remove toxic products of reducing chemistry. However, movement of up to 3 cm per day is only capable of oxygenating the surface 1 cm of ﬂooded soil. Daily soil percolation rates are seldom enough to oxygenate more than a few centimeters of soil near the surface. So, if water stands for any length of time, plants begin to suffer stress or die from the combined effects of inadequate aeration and accumulation of toxic substances. Frequent ﬂooding alters the mix of trees and other plants on the landscape (Hook, 1984; Kozlowski, 1984b). Plants with good internal aeration, or those that can withstand prolonged shifts to anaerobic conditions and resultant alterations of metabolism and accumulation of toxins, are favored (Jackson and Drew, 1984; Kozlowski and Pallardy, 1984). Scott et al. (1989) found that the response of soybean (Glycine max) to short-term ﬂooding depended both on the duration of ﬂooding and the growth stage at which it occurred.

PHYSICAL ASPECTS OF SOILS OF DISTURBED GROUND

Virtually all forms of management for ﬂooding avoidance or control are costly and difﬁcult to implement. They include regional dam and waterway projects, land contouring and other forms of surface drainage leading to improvement in inﬁltration, installation of subsurface drainage, selection and breeding of plant species and varieties for ﬂooding tolerance or resistance, protection of large wetlands and enhancement of their ability to slow storm-water discharge, and limited techniques for improvement of soil aeration once ﬂooding has occurred, such as the use of peroxide fertilizers (Cannell and Jackson, 1981; Kozlowski, 1984a; Stolzy and Sojka, 1984; Sojka and Stolzy, 1988). Erosion The topic of erosion is covered in detail elsewhere in this volume (Pimentel and Harvey, Chapter 4, this volume), with an emphasis on negative impacts on agriculture. This section will focus on limited aspects of water-induced erosion. Wind- and water-induced erosion both result from the interaction of the soil, in whatever physical state it exists upon initiation of the erosion event, with energetic ﬂuids (wind and water). Soil losses from each have been documented in natural and managed environments, commonly ranging from nearly zero to extremes of hundreds of metric tons per hectare per year. There is also a growing recognition of “tillage erosion”, with rates dependent on slope and implements reaching 140 t ha−1 yr−1 in severely affected areas such as hill crests (Mech and Free, 1942; Lindstrom et al., 1990, 1992; Govers et al., 1994; Lobb et al., 1995). Tillage erosion results from the gradual, but systematic, mass movement of soil downslope during tillage of sloping land. A natural analogue of tillage erosion is mass displacement of soil on slopes by the hooves of grazing or migrating animals and by burrowing soil macrofauna [e.g., marmots (Marmota sp.) or foxes]. The extent of these natural activities may rival the extent, but not the accumulated impact, of tillage erosion. Wind- and water-driven soil erosion processes have been intensively investigated since early in the 20th century. The physical processes involved have been described in well-recognized statistically based models: USLE, the Universal Soil Loss Equation (Wischmeier and Smith, 1978), RUSLE, the Revised Universal Soil Loss Equation (Renard et al., 1994), and the Wind

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Erosion Equation (Woodruff and Siddoway, 1965). Recently, intensive efforts have been mounted to produce process-driven models: WEPPS, the Water Erosion Prediction Project (Laﬂen et al., 1991) and WEPS, the Wind Erosion Prediction System (Hagin, 1991). These new models strive to predict erosion from basic soil and landscape characterization data and ﬂuid dynamics, rather than solely through empirically derived calibrations for a particular soil. The uniqueness of irrigation-induced erosion and irrigation-driven erosion processes is an important reﬁnement in erosion prediction technology that has not yet been successfully undertaken (Trout, 1996, 1999; Sojka, 1997). Soil properties inﬂuence water erosion through several avenues. Water erosion cannot occur until runoff begins. Runoff occurs when water is added at the surface faster than it can inﬁltrate. Inﬁltration rate is determined by soil structure and the water content of the soil proﬁle. When a soil proﬁle is saturated, runoff will occur regardless of structure. If the soil or the inﬁltrating water is high in sodium, structure near the surface can rapidly degrade, sealing the surface against water inﬁltration, and rapidly promoting runoff and increasing erosion. Erosion can increase by as much as 50% if the eroding water is high in sodium (Le Bissonais and Singer, 1993; Lentz et al., 1996). Regardless of salinity, the amount of energy (in all its various manifestations) associated with water interacting with soil affects the structural integrity of the surface soil and its erodibility (Laﬂen et al., 1991). The ability of raindrops and ﬂowing water to cause erosion depends upon the energy of the falling drops and the shear force of the ﬂowing water to cause detachment. Erosion takes place more easily when ﬂow over the soil surface is fed by large raindrops. The high energy of large raindrops destroys surface soil structure and detaches soil in the splash process. Al-Durrah and Bradford (1981) showed that the mass of surface soil detached by raindrops was linearly correlated with the ratio of raindrop kinetic energy to soil shear strength. Detached and displaced soil then washes away in runoff and contributes to sealing of the surface pores, further increasing runoff volume and shear force (Mohammed and Kohl, 1987). Finer-textured soils can be dispersed into ﬁner particles and therefore generally form surface seals more easily than coarse-textured soil (Bradford and Huang, 1992). Runoff volume and velocity increase downslope, making the slope angle and slope length important

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soil properties with respect to erosion. For a given volume of runoff, surface ponding depth decreases with slope steepness (Liu, 1991; Bradford and Huang, 1996). Ponding depths greater than 2 mm can reduce splash-induced detachment (Palmer, 1963; Kirkby and Kirkby, 1974; Moss and Green, 1983). Slope and slope length tend to increase runoff, and therefore ﬂow shear force, resulting in increased erosion. Torri (1996) recently pointed out that many soil and landscape factors interact in governing the effects of slope and slope length on erosion. Simple estimates of how much erosion increases for a given increase in either slope or slope length are difﬁcult to make accurately without considering other soil factors. Renard et al. (1996) recently demonstrated the effectiveness of new tools for predicting erosion such as RUSLE and WEPP, because of their ability to integrate numerous soil properties. In considering slope, it is also important to consider landscape position. Soil disturbance from erosion on upper-slope-reaches will be primarily the result of soil loss. On foot slopes, erosion-caused disturbance effects will be from deposition of soil as slope decreases, reducing ﬂow velocity, shear strength, and carrying capacity of the runoff (Franzmeier, 1990). These soil conditions affecting erosion can be greatly mitigated by increased surface roughness or coverage of the soil surface with modest amounts of vegetation, plant litter, or crop residues. Less well known but equally effective are erosion mitigation by dominance of divalent cations in the soil or in the runoff water (impeding the clay dispersion that results in aggregate destruction and particle detachment), or by stabilization with surface-applied soil conditioners that enhance aggregate stability and cohesion among aggregates (Renard and Mausbach, 1990; Lentz et al., 1992, 1996; Sojka, 1997). In the manageable upper portion of the soil proﬁle, soil organic matter is once again a very important soil property indirectly affecting erosion through its impact on soil structure. Higher organic-matter content promotes and stabilizes aggregation and, because of the improved macro-porosity, increases inﬁltration (Boyle et al., 1989). Some erosion always occurs in natural systems as the topographies of land masses change through geomorphic forces such as uplift, vulcanization, and glaciation. “Natural” (pre-anthropogenic) erosion rates for the southeastern United States were estimated from geological data to range from 0.2 to 0.8 t ha−1 yr−1

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for the Piedmont and from 2.6 to 4.7 t ha−1 yr−1 for the mountainous areas (Schumm, 1963; Hack, 1965; Young, 1969). Natural erosion can be as imperceptible or as catastrophic as anthropogenic erosion. Erosion is also essential to the development of various natural landforms, such as river deltas, ﬂoodplains, and valleys. Cyclical erosion is vital to the nutrient enrichment of certain natural land-forms and ecosystems for the survival of terrestrial plant growth and the fauna they support, as well as for lacustrine, ﬂuvial, and marine ﬂora and fauna in various environments. Ecosystem balance can also be destroyed if overwhelmed by excess erosion. The activities of man in the 20th century have greatly obscured the degree to which erosion occurs as a natural process. Most of the study of the phenomenon of erosion has been from the anthropogenic perspective, particularly with regard to agriculture and construction. Erosion effects can be divided into effects on depositional areas and on eroded areas within the context of landscape processes (Daniels and Bubenzer, 1990; Franzmeier, 1990). Deposition increases soil depth, and can alter soil chemistry, water retention, and aeration. Some organisms may be buried by deposition, and others may be seeded or translocated. Deposition or erosion may alter landscape contours enough to affect microclimate through slope and aspect changes, and may alter the availability of water by changing the depth to the water table or by interception of runoff. Aside from the immediate negative impact of the burial of some species, enrichment from soil deposition can also beneﬁt the depositional microhabitat (Cassel and Fryrear, 1990). On eroded areas, soil loss can create rills or gullies, and expose subsoil horizons that often are less supportive of vegetative growth. Erosional soil loss can undermine root support for higher plants. Most analysis of erosion has focussed on its negative impacts on crop yield (Cassel and Fryrear, 1990; Hajek et al., 1990). In an agricultural setting, erosion carries away topsoil, which usually has greater long-term yield potential than the subsoil that is exposed (Fig. 21.4). Furthermore, inputs including fertilizer and pesticides are lost, decreasing short-term productivity, while also risking non-point pollution downstream of the erosional site. Because processes of soil formation result in roughly parallel horizontal layers (called horizons), erosion and deposition generally result in progressive alterations

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REFERENCES

Fig. 21.4. Relationship of crop yield (normalized as percent of maximum yield without erosion) of six irrigated crops to uneroded depth of the surface horizon, demonstrating increased loss of yield potential with increased severity of erosion (from Carter, 1993). The six crops represented are: sugar beet (Beta vulgaris), barley (Hordeum vulgare), alfalfa (Medicago sativa), dry beans (Phaseolus vulgaris), sweet corn (Zea mays), and wheat (Triticum aestivum).

in soil characteristics such as bulk density, clay content, and surface horizon thickness. However, these changes vary systematically with position on the landscape (Walker, 1966; Malo et al., 1974; Matzdorf et al., 1975), and vary somewhat unpredictably among individual soils.

CONCLUSIONS

The physical aspects of soils of disturbed ground are best examined from the effect of the disturbance on the gas, liquid, or solid phases of soils. Soil properties can greatly inﬂuence the extent to which a given disturbance force affects the soil. Some of the effects of soil disturbance can be managed by utilizing current understanding of physical and chemical principles. Agriculture is probably the largest single activity on the planet causing ecological disturbance of soils. Yet agricultural disturbance is a managed factor, whereas many other disturbance sources are not. The intensity of environmental impact depends on the skill of management and/or the intensity of environmental or anthropogenic disturbance. It is the manager’s role to utilize understanding of the soil system to mitigate negative environmental effects of management, regardless of the outcome being sought.

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Chapter 22

SOIL MICROORGANISMS Michael F. ALLEN, Edith B. ALLEN, Thomas A. ZINK, Sharon HARNEY, Lidia C. YOSHIDA, ¨ Concepci´on SIGUENZA, Fred EDWARDS, Cara HINKSON, Matthias RILLIG, David BAINBRIDGE, Christina DOLJANIN and Robert MacALLER

INTRODUCTION

Species diversity among soil microorganisms is greater than any other group of organisms at any terrestrial site. These organisms also play the greatest diversity of roles, from primary production through decomposition to nitrogen ﬁxation. After plants, these organisms are the largest biomass component. Despite these critical ecosystem roles, soil biology remains the least understood and least studied subject area within any type of ecosystem. With human perturbation of lands, loss of biodiversity and of the speciﬁc functions of these organisms in ecosystems could have devastating impacts both at the local and global scales. In order to understand the consequences of speciﬁc perturbations, and to put functioning ecosystems back together, it is essential to understand what these organisms are, what they do, and how and where they live. Then, one must understand how they disseminate, and how their migration patterns may be utilized to restore and manage disturbed lands. Studying disturbed areas also gives especially useful opportunities for studying microbial functioning. This is because there are measurable changes simultaneously in the composition, mass, and functioning of these organisms. In disturbed areas, additions of particular groups can be undertaken and responses measured. We have had the chance to work on catastrophic natural and human perturbations such as strip mines, areas of glacial retreat, volcanoes, and damage by off-road vehicles. In these sites, we have been able to document microbial re-invasion, to add speciﬁc microbes and to study the consequences of these invasions. These studies have demonstrated beyond a doubt that many of the most exciting aspects of ecology remain to

be discovered, and occur at scales barely imaginable to most ecologists. Below we describe ecosystem processes and the soil organisms catalyzing these processes, with the goal of outlining some principles for their management and some areas in need of future research. One aspect that must be considered is the evolutionary history of soil microbes, because the vast array of their roles is a reﬂection of the earth’s history. Disturbance, after all, is an ancient process, and the invasion of truly barren lands by soil organisms billions of years ago was one of the most important biological events after the initiation of life itself. Photosynthesis is the process that drives terrestrial life, and photosynthesis ﬁrst appeared in microbes. Soil cyanobacteria and algae are among the important components lost when arid soils are disturbed, which must be replaced to reconstruct a sustainable desert. Cyanobacteria (both free-living and symbiotically in lichens) living in soil in the tropics, deserts, and tundra ﬁx a large fraction of the nitrogen essential for plant photosynthesis. Asymbiotic and symbiotic bacteria provide most of the remainder. Phosphorus probably was the next major limitation to terrestrial invasion. Mycorrhizae and lichens, associations between fungi and either algae or primitive plants appear to have evolved in part to deal with this limitation (Malloch et al., 1980; Taylor et al., 1995). Understanding microbial functions and the limits to microbial functioning is critical to studying ecosystem responses to disturbance and recovery from disturbance. In this light, soil microbes become important tools for land managers in restoring severely disturbed habitats for all animals. Our goal in this chapter is to provide an overview of soil microbial activity in

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disturbed soils, and to provide some practical as well as theoretical perspectives for managing these crucial resources.

SOIL MICROORGANISMS – WHAT ARE THEY?

Soil organisms at any site, whether disturbed or undisturbed, are highly diverse both phylogenetically and functionally. These range from the multiple lines of “bacteria” that regulate many unusual processes, through Cyanobacteria and algae that actively photosynthesize in soil, to fungi, many of which are really macroorganisms with individual hyphae that stretch across several square meters of soil, and to a large array of invertebrates that feed on anything from bacteria to plants. Importantly, we really have as of yet only a marginal idea of how many species comprises soil organisms as a group (Table 22.1). Phylogenetically, soil microorganisms are extremely diverse. Based on molecular evidence, the “bacteria” have been divided into two kingdoms, the Eubacteria, or true bacteria, and the Archaebacteria, or prokaryotic organisms that live in highly specialized stressful environments. The eukaryotes have been placed in a third kingdom (Woese, 1987). The numbers of eubacterial species at any one site is hotly debated, but the debate seems to be converging on 4000– 10 000 species in one gram of soil, using a variety of molecular approaches (e.g., Tiedje et al., 1989; Torsvik et al., 1990; Paul and Clark, 1996). These ﬁndings may be repeated over more than one site, especially following a disturbance in which the conditions are rather uniform. The bacteria responsible for nitrogen oxidation appear to be very limited taxonomically (Mauret et al., 1996), and at any site there are generally only a limited suite of nitrogen ﬁxers. Both groups of bacteria are quite sensitive to soil disturbance and loss of speciﬁc plant groups. The Archaebacteria undertake rather specialized tasks. They include the methanogens, extreme halophiles, and extreme thermophiles. While not generally considered in soil dynamics, these may become important under toxic conditions created by waste disposal or other disturbances. Fungi constitute what is usually the second largest group (after plants) in biomass. They are generally viewed as microorganisms because individual hyphae tend to be only 2 to 15 mm in diameter. However, they often can extend over long distances. An “individual” fungus of Armillaria has been found to occupy over

M.F. ALLEN et al. Table 22.1 Estimated numbers of species 1 Species

Described

Estimated 2

Viruses

5000

perhaps 500 000

Bacteria

4000

400 000–3 million

Fungi

70 000

1.0–1.5 million

Protozoans

40 000

100 000–200 000

Algae

40 000

200 000–10 million

Plants

250 000

300 000–500 000

Vertebrates

45 000

50 000

Roundworms

15 000

500 000–1.0 million

Mollusks

70 000

200 000

Crustaceans

40 000

150 000

Spiders, mites Insects 1 2

75 000 950 000

750 000–1.0 million 8–100 million

From Systematics Agenda 2000 (1994). Estimated numbers of species remaining to be discovered.

1 ha, and an ectomycorrhizal fungus can encompass several trees within its mycelial matrix (see Allen et al., 1993). It is only in rare environments that all major fungal groups are not represented. However, the species diversity for a single site has never been measured. Overall data for fungal distribution are poor, particularly in the tropics. Hawksworth (1991) using a 1:6 ratio of plant species to fungal species (with some variations), suggested that, because of the high diversity of tropical trees and minimal survey work, only 4% of all fungal species have been identiﬁed. While this may overestimate the total number of unidentiﬁed fungal species (Allen et al., 1995; May, 1991) the signiﬁcance of this ignorance is overwhelming. Soil organisms form complex food webs to a degree rarely considered by theoretical ecologists. They range from extremely simple food webs with as few as three species (Freckman and Virginia, 1997) to webs with literally thousands of species (e.g., Moore et al., 1996). Indirectly, soil invertebrates inﬂuence fungal communities through comminution, channeling and mixing, and directly through grazing and distribution of spores (Visser, 1985). Despite documentation of the structure and functioning of soil food webs, there is little understanding of their structure in disturbed lands. Nevertheless, the organisms comprising these webs may be crucial for describing the functioning of ecosystem processes in disturbed lands (Whitford, 1988). The dominant groups of soil microinvertebrates

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are the Protozoa, nematodes, mites, and Collembola. These are the most common. Depending on the site, other interesting microfauna such as water bears (tardigrades) may be found. In a review on the diversity of soil arthropods in Canada, Behan-Pelletier (1993) reported that a soil sample from anywhere in North America will contain at least one unknown species. She estimated that only 25% of the Collembola and Oribatida, which have been most extensively studied, are known at the species level.

MICROBIAL ROLES

It is often assumed that soil microorganisms replace each other following a perturbation, so that the processes that they regulate will always occur. However, this idea derives from the notion that the “function” of soil microorganisms is to decompose organic matter, and that decomposition always happens. However, it has been known for a very long time that a great many organisms contribute to decomposition (e.g., Darwin, 1881) and that soil microorganisms in particular undertake a wide array of environmental functions (e.g., Winogradsky, 1890). To assess how effectively microbes replace one another, and thus how soils should be managed after disturbance, scientists and managers need an appreciation of the vast array of roles the microbes undertake. Carbon allocation to below-ground components may reach 85% of daily photoassimilates (Allen, 1991). The below-ground components of carbon allocation include root growth and turnover, root exudation and sloughing, and associations with microorganisms (Marschner, 1995). Soil microorganisms, in return, provide plants with increased nutrients and water (Allen, 1991; Marschner, 1995; Paul and Clark, 1996). There are many mechanisms and interacting processes through which soil organisms, including fungi, bacteria and invertebrates, facilitate the mobility and uptake of nutrients from organic matter and inorganically bound substrates. Within this general context, symbioses between microorganisms and plant roots may range in degree from mutualism, where both plant and microorganism beneﬁt, to parasitism, where the plant host is negatively affected and the microorganism receives most or all beneﬁts from the association (deBary 1887; Lewis, 1973; Smith et al., 1979). The nature of this symbiosis can, therefore, impact plant productivity and interactions among plants

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within a community (Allen, 1991). Thus, the belowground component of ecological systems contributes signiﬁcantly to ecosystem functioning. Elemental cycling Classically, bacteria and fungi constitute the soil microﬂora, the most abundant of the soil microorganisms. Most measurements of “microbial biomass” are largely composed of these organisms. These are concentrated in the rhizosphere, where the plant provides carbon substrates through root exudation and sloughing (Curl and Truelove, 1986; Richards, 1987). Bacterial populations include proteolytic, ammonifying, denitrifying, sugarfermenting, and cellulose-decomposing types (Curl and Truelove, 1986). Nitrogen-ﬁxing bacteria may be symbiotic with legumes (various genera of rhizobia); form actinorhizae (with Frankia), with a wide array of plant species; or exist as free-living bacteria in the rhizosphere (Steyn and Delwiche, 1970; Vlassak and Raju, 1977; Stewart, 1982; Ellis and Kummerow, 1988). Thus, bacteria play an important role in driving the nitrogen cycle through ammoniﬁcation, nitriﬁcation, denitriﬁcation, and nitrogen-ﬁxation. In addition to the nitrogen cycle, fungi and bacteria facilitate mobilization of organic phosphorus by phosphatase production (Helal, 1990) and through assimilation of inorganic phosphorus followed by release through grazing by the soil fauna (Curl and Truelove, 1986). Bacteria and fungi are also capable of solubilizing immobile phosphate from inorganically bound substrates by releasing organic acids and siderophores (Curl and Truelove, 1986; Cannon et al., 1995). Some bacterial and fungal species oxidize inorganic sulfur to produce sulfate in neutral and alkaline soils (Paul and Clark, 1996). In addition, bacteria may be involved in facilitating the diffusion of nutrients in the soil solution toward the root surface (Curl and Truelove, 1986). Both fungi and bacteria are involved in increasing availability of cations and anions of phosphorus, calcium, potassium, iron, zinc, aluminum and manganese, by the release of organic acids and speciﬁc iron-binding agents or siderophores, and changing the pH and oxygen partial pressure in the rhizosphere (Marschner, 1995). Cycling of nutrients between plants and soil is facilitated by mechanical or chemical decomposition of organic matter, and the death of immobilizing bacteria and fungi. Soil microfauna, (primarily nematodes, Protozoa, and Collembola) and other invertebrates (e.g., earthworms, insects, millipedes), contribute to

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decomposition and nutrient turnover rates primarily as secondary processors, by grazing on fungi and bacteria (Coleman and Crossley, 1996). In addition, these organisms increase decomposition rates by mechanically restructuring soil by making channels, carrying organic matter throughout the soil proﬁle, and breaking down large pieces of organic matter into smaller units, which increases the surface area for aeration and microbial breakdown (Kitchell et al., 1979; Richards, 1987; Paul and Clark, 1996). Despite the secondary role of soil fauna relative to soil microbes, greater complexity of the soil food web increases carbon utilization and nutrient release (Coleman et al., 1977; Anderson et al., 1981; Setala, 1995). This is especially evident in nitrogen mineralization, as bacteria and fungi break down fresh litter with carbon/nitrogen ratios ranging from 50 to over 1000 into microbial mass with carbon/nitrogen ratios of 10 or less. This results in the release of carbon but not nutrients. This microbial mass is then consumed and processed further by soil fauna (Paul and Clark, 1996), releasing nutrients as well as carbon dioxide. Thus, soil faunal activity increases with increasing activity of rhizosphere bacteria and fungi because of the dynamics associated with decomposition, as well as the availability of a food source for grazing (e.g., Paul and Clark, 1996). Mycorrhizae act to improve nutrient uptake by increasing the surface area of the host root system (Allen, 1996). This allows plants to access nutrient pools beyond the depletion zone around the root (Tinker et al., 1992; Marschner, 1995). Availability and uptake of phosphorus is known to increase signiﬁcantly in mycorrhizal plants as compared with non-mycorrhizal plants when phosphorus supply is low (Allen, 1991; Li et al., 1991; Marschner, 1995). Higher respiration rates of mycorrhizal roots as compared with those without mycorrhizae increases carbon dioxide production, thereby increasing solubility of calcium phosphate and the mobility of phosphorus (Knight et al., 1989). Acid and alkaline phosphatase activity by mycorrhizal roots increases the release and availability of phosphorus from organic phosphorus pools (Allen et al., 1981; Gourp and Pargney, 1991; Dinkelaker and Marschner, 1992; Tarafdar and Marschner, 1994). Oxalates produced by some plants, as well as mycorrhizae, also increase the release and uptake of phosphorus, by immobilizing aluminum, iron, and potassium, through chelation of these inorganically bound elements (Cromack et al., 1977; Cannon et al., 1995; Allen et al., 1996). Further, a wide variety of both

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fungi and bacteria then decompose the oxalates, thereby releasing the cations into the rhizosphere (Morris and Allen, 1994). Rhizosphere microorganisms are also known to mobilize potassium from crystal lattices of mica and feldspar, thus increasing its availability to plants (Voigt, 1965; Boyle et al., 1967; Dahlgren et al., 1994). Other nutrients made more available for plant uptake through mycorrhizal association include zinc and copper (Kothari et al., 1991a;b; Lambert and Weidensaul, 1991), magnesium, and sulfur (Marschner and Dell, 1994). Calcium transport by hyphae has been demonstrated, but calcium concentration in shoots is often lower in mycorrhizal plants (Kothari et al., 1990). Iron acquisition is facilitated by siderophore production in the fungi of ectomycorrhizae and ericoid mycorrhizae (Schuler and Haselwandter, 1988; Shaw et al., 1990; Crowley et al., 1992). Nitrogen is taken up by mycorrhizae as ammonium more readily than as nitrate, but many mycorrhizal fungi produce nitrate reductase (Ho and Trappe, 1975; Plassard et al., 1991). Nitrogen may be transported by arbuscular mycorrhizae or ectomycorrhizae in the form of arginine or glutamine associated with polyphosphates (Marschner, 1995). Some ectomycorrhizal species and those of ericoid roots access nitrogen from organic pools through the release of proteinases (Marschner, 1995). This capacity is ecologically important when nitrogen availability is poor due to soil leaching, gaseous loss, or competition with soil microorganisms (H¨ogberg, 1990; Vogt et al., 1991; Dahlgren et al., 1994). Mycorrhizal fungal hyphae may directly moderate plant uptake of heavy metals and reduce potential toxicity (Wilkins, 1991; Colpaert and van Assche, 1993; Marschner, 1995). When soils are acidic, certain metals that are bound and unavailable at a higher pH become available and potentially toxic. Ectomycorrhizae may be directly involved in moderating plant toxicity through sequestering heavy metals in the external mycelium or sheath, thus reducing the rhizosphere concentrations encountered by the root (Marschner, 1995). Heavy metals may bind to mucilage associated with the hyphae, be incorporated into fungal cell walls, or be stored in vacuoles containing phosphorus or nitrogen compounds (Kottke, 1992; Turnau et al., 1993). However, the capacity of mycorrhizae to moderate plant-uptake of heavy metals cannot be regarded as general, as mycorrhizae may also increase their uptake (Colpaert and van Assche, 1992). In contrast to direct moderation of heavy metal uptake, mycorrhizae were found to moderate manganese uptake

SOIL MICROORGANISMS

indirectly by increasing the activity of rhizosphere bacteria oxidizing manganese (Arines et al., 1989, 1992), decreasing the numbers of manganese-reducing bacteria (Kothari et al., 1991b), and increasing the root exudation of organic compounds that form complexes with manganese (Arines et al., 1989). In Pinus rigida, improved phosphorus nutrition led to indirect amelioration of aluminum toxicity (Cumming and Weinstein, 1990). Interactions between rhizosphere microorganisms Interactions between microorganisms may act additively or synergistically to increase beneﬁts of plant– microbe symbioses. For instance, plant associations both with nitrogen-ﬁxing bacteria and with mycorrhizae may act independently to increase plant productivity by increasing soil mineral nutrients and ﬁxednitrogen, resulting in greater photosynthesis. Increased photosynthesis in turn leads to greater photosynthate allocation below-ground, beneﬁting each microbe association indirectly (Cluett and Boucher, 1983; Carpenter and Allen, 1988). Nitrogen ﬁxation is known to increase in the presence of mycorrhizae as a result of improved phosphorus nutrition, which increases nodulation and nitrogenase activity, in addition to increased uptake of copper, molybdenum and zinc, which are involved in nitrogen-ﬁxation (Mosse et al., 1976; Smith et al., 1979; Rovira et al., 1983; Barea et al., 1989). Increased nitrogen ﬁxation has been observed in legumes (Hayman, 1987), actinorhizal plants (Gardner et al., 1984), and free-living associations (Pacovsky et al., 1985) when co-occurring with mycorrhizae. However, mycorrhizal infection decreases the development of root hairs (Harley and Smith, 1983), which could negatively affect rhizobial infection of root hairs for nodule development (Allen, 1991). In addition to interactions between nitrogen-ﬁxers and mycorrhizae, beneﬁts to plant productivity may result from other microbial interactions. For example, bacteria found in the mantle of ectomycorrhizae associated with Pinus radiata facilitated mycelial growth and mycorrhizal formation (Garbaye and Bowen, 1989). However, bacteria may also act as a deterrent to fungal root infection (Bowen and Theodorou, 1979). In addition, saprophytic fungi may produce antibiotics which reduce competition from bacteria and pathogenic fungi (Richards, 1987). Finally, plant roots may deposit exudates enabling non-pathogenic bacteria to outcompete pathogenic bacteria (Richards, 1987).

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Interactions between microbes and plants: the entire community Soil microorganisms reduce nutrient availability and plant uptake by competing with roots for nutrients. The rhizosphere is an ideal habitat for proliferation of bacterial populations, on account of the constant supply of carbon substrates for energy. Since roots take up nutrients from the rhizosphere, rhizosphere bacteria also have access to available nutrients for immobilization. This results in competition for nutrients between bacteria, fungi, and roots (Chapin, 1980; Paul and Juma, 1981; Rosswall, 1981; Vitousek, 1981; Duxbury et al., 1989; Jackson et al., 1989; Schimel et al., 1989). Bacteria and fungi are often nitrogenlimited, and increase in biomass, activity, and turnover rates with nitrogen fertilization (Liljeroth et al., 1990). Harte and Kinzig (1993) presented a model that incorporates the mutualistic–competitive dynamics within the plant–microorganism interaction into the context of population dynamics and ecological structure. They concluded that soil microorganisms utilize inorganic nutrients to maximize their biomass, and plants acquire the left-overs. Plants allocate 28 to 59% of their daily production of photosynthate to roots, 4 to 70% of which is deposited into the rhizosphere (Lynch and Whipps, 1990). Root exudation and secretions, and loss of root cells and tissues contribute plant-derived carbon to soil and soil microorganisms in a wide variety of compounds including organic acids, sugars, cellulose, lignin and proteins (Richards, 1987; Paul and Clark, 1996). The recycling of nutrients from plant litter through decomposition by the microbial and faunal communities in the soil releases nutrients that are taken up by plants directly or through mycorrhizal hyphae to maintain or increase plant photosynthesis and productivity. However, when soil nutrient availability is high, the role of mycorrhizae becomes less valuable. Supporting this relationship becomes less beneﬁcial to the plant since a signiﬁcant investment in photosynthate production goes toward supporting mycorrhizae and away from plant growth (Allen, 1991; Setala, 1995; Johnson et al., 1997). It would appear that in this case, mycorrhizae become parasitic on the plant. Thus, the plant–microbe association appears to be mutualistic during “ecological crunches” and benign to slightly parasitic at other times (Allen and Allen, 1986). Parasitism is typically deﬁned as an association between two organisms where one organism, the parasite,

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receives resources to the detriment of the host (deBary, 1887; Lewis, 1973). The continuum between parasitism and mutualism in the plant–soil–microbe system is governed by the presence or absence of abiotic and biotic factors that may serve to keep in check the potential parasitic nature of soil microorganisms. Adding soil fauna to a microcosm study decreased mycorrhizal infection and biomass but increased plant uptake of nitrogen and phosphorus and plant growth (Setala, 1995). The functional signiﬁcance of mycorrhizae was maintained even after soil fauna were allowed to graze upon mycorrhizal hyphae, and the advantages to the plant were greater than when soil fauna were not present. In natural systems, mycelia may connect root systems between several individual plants (Read et al., 1985). Thus, a one-to-one relationship, which is typically used to classify parasitism (deBary, 1887), does not necessarily occur between plant and fungus in the ﬁeld, but rather a network of exchange of carbon and nutrients occurs between plant and fungus over a larger scale. Determining differences in interactions across scales, therefore, provides the approach needed to observe the crucial role of the complex belowground component in ecological functioning of natural systems.

MICROBIAL RESPONSES TO DISTURBANCE

As discussed previously, soil disturbance can occur in many forms, including physical, chemical, hydrological, and biological disturbances. However, no matter the actual disturbance agent, all disturbance will alter in some way the availability of resources and the community structure within the disturbed soil. How a system responds is a consequence of the spatial scale and intensity of the alteration of these parameters. The various processes linked with decomposition and the subsequent recycling of nutrients are closely associated with both soil microﬂora (bacteria and fungi) and soil microfauna. The composition and function of both microﬂora and microfauna will inevitably be affected by any disturbance, possibly impacting one or all of the following factors: diversity, functional groups, microbial-mediated processes, and spatial patterns. If a disturbance is small in scale, overall species richness and diversity may be increased through the opening of small patches for colonization by species that would not normally occur without a disturbance. However, for most disturbance regimes, there is an initial decrease

M.F. ALLEN et al.

Fig. 22.1. Species increment curve for saprobic microfungi of a disturbed shrub-steppe (open diamonds) and its “matched” control site (open circles) (data from Allen and MacMahon, 1985).

in species diversity and community structure, and a decrease in nutrient cycling and resource use, which may result in changes to successional paths. In a comparative study on strip mine spoils in northern New Mexico, Elkins et al. (1984) found density and diversity of microarthropods to be greater in unmined control soil when compared to both untreated and amended disturbed soils. Using a dilution-plate method and a taxa increment curve, Allen and MacMahon (1985) studied a disturbed mine site, using 475 isolates from 25 sample points. The rate at which additional species were found continued to increase with increased sampling in the control. However, in the disturbed soil, the rate of increase declined (Fig. 22.1). As there is no leveling at the “control” site, it remains unknown how many microfungal species are present. In the disturbed soils, there were probably between 150 and 200 species of soil microfungi. Decreased density and diversity of microorganisms in disturbed sites were also found in several other such studies (Zak, 1992). In their study on the effects of chemical disturbance on microbial populations, Atlas et al. (1991) showed that the diversities of communities that were chemically disturbed were lower than those of undisturbed communities. This decrease in taxonomic diversity caused by disturbance was signiﬁcant, and led to the survival of relatively few populations compared to those of undisturbed communities. Paoletti and Bressan

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(1996) also studied the effects of disturbance on the diversity of soil invertebrates, and found that density and diversity decreased as a result of disturbances ranging from agricultural operations to the impact of radioactive fallout from the Chernobyl disaster. Although diversity has been used as a parameter for measuring ecosystem health, in the belief that higher diversity means a more stable ecosystem, taxonomic variability is only one indicator of a healthy system. Soil ecosystems function through the use of microbialmediated processes such as decomposition and nutrient cycling. It is known that lands disturbed by either anthropogenic or natural causes show decreases in taxonomic diversity, and high initial nutrient inputs due to release by microbial death. What is less well appreciated is that there are also changes in nutrient cycling due to altered rates of decomposition (either increases or decreases). These often fail to match production rates, which results in immobilization or mobilization of limiting nutrients, and perhaps to leaching losses (Carpenter and Allen, 1988; George et al., 1993). Fully functional cycles require a self-sustaining mineralization process resulting from the interaction between microﬂora and microfauna (Carpenter and Allen, 1988). The soil decomposition/mineralization processes can be disrupted by disturbance in several ways. The physical removal of the plant cover interrupts plant uptake of nutrients, mainly nitrogen, and thus initially increases nitrogen availability through increased decomposition and mineralization (Matson and Vitousek, 1981). If the area had originally been dominated by perennial shrub species selected for slow, low nutrient use, a ﬂush of nutrients could lead to invasion by annuals with a high demand for nutrients and a resultant change in habitat (Allen et al., 1996). Hyphae of mycorrhizal fungi are critical components, in that they spread farther out and into smaller soil pores than plant roots. Disturbance of these hyphae can occur through compaction of soil pores or physical destruction of the hyphae (Allen and MacMahon, 1985). A decrease in mycorrhizal infectivity resulting from disturbance has been noted in numerous studies (Jasper et al., 1989a,b; McGonigle et al., 1990; Jasper et al., 1991; Nadian et al., 1996). Another result of disturbance, which has mostly been overlooked, is the effect of disturbance on the spatial heterogeneity of soil microorganisms. In almost every case of disturbance, the microbial richness and system heterogeneity is destroyed. Allen and

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MacMahon (1985) suggested that the loss of spatial heterogeneity between the soil fungi, organic matter, and nutrient pools was a key factor affecting reclamation success on a strip-mined site in western Wyoming. Reclamation processes broadly dispersed fungal genera across the reclaimed site, leading to site homogeneity, while undisturbed sites had fungal taxa organized in islands of fertility around shrubs. This loss of spatial heterogeneity within the soil community affected both the species composition and structure of the aboveground system. The effects of disturbance on microorganisms and their importance in any reclamation process has been studied for many years. Microorganisms are key elements in all below-ground processes required for the existence of a self-sustaining ecosystem. Study of the effects of disturbance must not focus only on aboveground systems, but should determine the impact on the dynamics and functional status of the soil microbial community. Reclamation efforts must also incorporate the microﬂora–microfauna interactions, and consider how disturbance has disrupted these processes.

RE-INVASION

Soil microorganisms, because of their size and the dynamic habitat in which they live, will only survive if they are adapted to dispersal. Two dispersal strategies appear to be common: the ﬁrst, an investment of energy into producing large numbers of propagules with less specialized dispersal mechanisms; and the second, producing relatively smaller numbers of propagules with specialized dispersal mechanisms. Both abiotic and biotic vectors disperse soil microorganisms. The differences between the two sets of vectors are probably responsible for the two dispersal strategies mentioned. Dispersal by wind and water is spatially broad, and the majority of propagules dispersed will not encounter a favorable location. Dispersal by biotic vectors, on the other hand, tends to enhance deposition in safe sites. Along with these life-history strategies, most soil microorganisms have evolved mechanisms to help them cope with unfavorable conditions including spores, eggs, and cysts. In some cases, development of these structures and the life-cycle stage at the time of disturbance may be important factors in determining dispersal of the microorganisms and their recolonization of new sites. In the most active state, one response of soil

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organisms is to grow, walk, or crawl onto disturbed areas. Fungal hyphae expand into open habitats and along roots. Warner and Mosse (1980) estimated that the fungus Glomus mosseae could invade across an onion patch at a rate of about 10 cm yr−1 . Zachvatkin (1959) measured the speed of movement of the mite Glycyphagus michaeli at 2.5 mm sec−1 . However, Allen et al. (1993) and Siepel (1994) noted that growth or locomotion is only of minor importance compared to dispersal by abiotic and biotic factors and at best operates only on edges or islands of undisturbed habitat. For soil organisms it is much more efﬁcient to get a free ride either on wind, water, or another organism. Water is essential in all soil environments. Gravitational water, capillary action through soil pores, and surface tension can transport individual cells, spores, hyphal fragments, and the eggs of nematodes and microarthropods. Keen (1922) calculated that capillary action through pores in ﬁne silt could lift water 45 m. Subsurface soil bacteria are adapted to below-ground water movement. Bacterial adsorption onto soil particles is a result of the physical and chemical nature of the soil, water, and cell surface modiﬁcations. Dispersal in these bacteria is inﬂuenced by factors that promote or release cell settlement (Lindqvist and Bengtsson, 1991). Closer to the soil surface, runoff and erosion undoubtedly transport soil microorganisms. But the degree of effectiveness is uncertain and depends on the organism and type of propagule. Friese (1984) and Gemma and Koske (1990) noted that strands of mycorrhizal Ammophila were found that had been washed out to sea and redeposited back on beach fronts. The inoculum remained viable and could serve to reinoculate beach-front communities. On the other hand, Allen et al. (1984) could not ﬁnd spores in areas on Mount St. Helens where eroded material had been deposited. They suggested that there may be important differences in the survival of spores from that of root inoculum in materials immersed in water. Wind is probably the best known and most studied dispersal agent for soil organisms. Organisms dispersed in this manner must survive exposure to ultraviolet radiation, desiccation, and freezing. Jeppson et al. (1975) reported over 90% mortality for wind-dispersed microarthropods. Many soil fungi have evolved pigmentation and fortiﬁed cell walls to cope with ultraviolet radiation and desiccation. Wind dispersal can be divided into three crucial processes: entrainment, migration, and deposition.

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Entrainment Entrainment covers the processes whereby the propagules are “entrained” into large-scale wind-ﬂow patterns. Once entrained, the propagules are dispersed as long as the turbulence maintains them within the atmosphere (Allen et al., 1989). Many species of bacteria and fungi take advantage of entrainment by rainfall or droplet splash. Puffballs and earthstars (Lycoperdon, Geaster) and bird-nest fungi (Cyathus) are among the genera that release spores when struck by rain droplets. While increases in bacterial aerosols during rainfall have been shown (Graham et al., 1977), the net movement of bacteria in aerosols is downward (Constantinidou et al., 1990). This suggests that creation of aerosols by rainfall may not always be a major entrainment factor. Instead interaction with the soil surface is probably responsible for most entrainment. This entrainment can act on both distant and local scales. On a local scale, Allen (1987) found at Mount St. Helens that no spores of arbuscular mycorrhizal fungi were trapped in dense canopy sites or in the midst of the barren Pumice Plain, a long distance from the source area. However, at the margins of the blow-down areas, where there was virtually no plant canopy, but where the rodent Thomomys talpoides piled buried soil on the ash surface, Glomus spores were entrained and dispersed locally. Migration On a large scale, Warner et al. (1987) and Allen et al. (1989) demonstrated that arbuscular mycorrhizal fungi could disperse up to 2 km by wind, in the form of trapped spores and spore mimics. Allen et al. (1993) also observed that distribution patterns may be inﬂuenced by capacity for wind dispersal. A wide distribution pattern was observed for the smaller Glomus spores across the Great Basin of the United States. These spores have low terminal velocities, about 20 cm s−1 (Allen et al., 1993). In contrast, a rather narrow distribution pattern within this region was observed for many larger-spored species, with terminal velocities of 200 to 400 cm s−1 . Deposition Deposition studies are actually surprisingly rare. Although deposition is not well studied, any plant or other wind-break results in deposition by absorbing the wind

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energy and creating vertical wind speeds lower than the terminal velocity of the spores (Allen et al., 1993). There are many examples where a combination of biotic and abiotic factors are important to dispersal. Murie (1962) noted that Sciurus spp. made caches of sporocarps in the crotches of trees and allowed them to dry out. Any spores dropping from these would be entrained in the air mass and capable of dispersing for very long distances; mycorrhizal fungi are also commonly collected by Sciurus spp. (Allen, 1991). In a Mima mound created by Thomomys bottae in a coastal mesa site in the San Diego area (California), Cox (1990) estimated that deposition of soil in surface heaps amounted to 8.23 t ha−1 , and subsurface redeposition to 20.31 t ha−1 , so that total mining was 28.54 t ha−1 . For reference purposes, a soil system with a bulk density of 1.40 g cm−3 would contain 2800 t ha−1 in the surface 20 cm. Thus, an annual mining rate of 28 t ha−1 would imply a complete turnover of the surface soil system once per century. Allen (1991) suggested that disruptions to the soil surface by Thomomys spp. and Taxidea taxus are essential to the physical dispersal process by bringing the fungal inoculum to the soil surface, where surface winds can entrain that inoculum into the atmosphere for longdistance transport. It is reasonable to hypothesize that other soil microorganisms may be transported in this manner. Many soil microorganisms have evolved symbiotic relationships to facilitate dispersal. Animals move by sensing their surrounding environment. Szacki et al. (1993) noted that mammal movement is based on the entire landscape. By taking advantage of animal movement, many soil microorganisms are better able to ensure deposition into favorable locations. Two of the most common are mycophagy or ingestion of fungal material, and phoresy or active attachment by microarthropods to other animals (usually insects). Many rodents ingest fungi by consuming infected root material, and transport these fungi along their travel routes. Examples include Thomomys talpoides in subalpine and alpine meadows (Allen et al., 1992), Proechimys spp. in the Peruvian tropical forests (Emmons, 1982), Rattus spp. in the Australian eucalyptus forests (McGee and Baczocha, 1994), Sciurus spp. from pine (Pinus spp.) forests (Kotter and Farentinos, 1984; States, 1985), and Peromyscus maniculatus from shrub steppes (Warner et al., 1987). Large mammals

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may also serve as vectors. These include Alces americanus, Antilocapra americana, Cervus canadensis and Odocoileus hemionus (Allen, 1987; Warner et al., 1987; Cazares and Trappe, 1994). Fungal spores are not the only soil microorganisms capable of surviving passage through the digestive tract of animals. Nematodes and soil microarthropods including the Acari are capable of passing through the alimentary canal of organisms of various vertebrates (Chmielewski, 1970). Finally, it is worth mentioning that humans have intentionally and unintentionally transported soil microorganisms around the world, and are probably among the most active dispersal agents ever known. Intentional inoculation has been undertaken around the world and resulted in some fungi being dispersed into habitats where they would not be expected naturally. Examples include the inoculation of Pisolithus tinctorius into many disturbed forest types (e.g., Marx et al., 1984). Unintentional inoculation with exotic soil microorganisms threatens to alter many soil ecosystems signiﬁcantly, and may lead to confusion in future scientiﬁc attempts to understand soil biodiversity and ecology. Cameron (1972) described an incident, while in the ﬁeld, where an Antarctic study site was unintentionally inoculated when a spoiled can of human food was opened and Penicillium spores were spread by the wind. In the next sample collection this organism was detected. In sites that have been visited by humans, future scientiﬁc investigations must be aware of, and take into account, the possibility that exotic soil microorganisms may have already been introduced.

ESTABLISHMENT

Establishment is the stage following immigration of propagules of a new individual or species, in which the newcomer becomes a permanent member of the local community. It is thus a critical phase between dispersal and subsequent successional patterns. While plant establishment has been subject to intense ecological study, knowledge of the establishment of microbes after disturbance events is comparatively scant. The arrival of a living microbial propagule is not synonymous with a subsequent presence or functioning of that species in an ecosystem, since there are interacting factors that can limit establishment. We identiﬁed and will discuss the following factors that limit establishment: viability, abiotic and biotic factors of the environment

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to be colonized, and factors limiting establishment for symbiotic microbes. All these aspects interact to determine establishment success. Fungal propagules arriving at a new site may be alive, but can be non-viable, that is, incapable of germination under any circumstances (Cooke and Whipps, 1993). Viable propagules can be non-germinable, because internal (or external) conditions are unsuitable for the initiation of germination (Moore-Landecker, 1990). For example, mycostasis is frequently observed in oligotrophic soils, on account of the leaching of metabolites out of fungal spores and their rapid uptake by other soil microbiota. Spores and other propagules thus fail to germinate, and may become increasingly incapable of doing so (Lockwood, 1988). Conversely, loss of the dormancy period and premature germination in ascospores has been shown to lead to bacterial lysis of germ tubes when no appropriate substrate was available for fungal colonization (Wall and Lewis, 1980; Gochenaur, 1987). Another reason for lack of success of a propagule may be limited movement in soil of the propagule and its progeny. The ability to move in soil varies greatly among bacteria (e.g., Gannon et al., 1991) and among spores of fungi (Hepple, 1960). This may be particularly important for rhizosphere microbes that need to colonize and follow root tips (Curl and Truelove, 1986). The survival of potentially viable and germinable spores, once they have reached a site, is subject to a host of abiotic conditions. Bioremediation projects often require the introduction of microbes to polluted and disturbed ecosystems. Reports of and research into failure of inoculation attempts have provided a rich source for the evaluation of factors limiting establishment of bacteria. Among the abiotic factors responsible for unsuccessful establishment of bacteria are nutrient limitation, absence of growth factors, lack of a suitable carbon source for growth, pH, and salinity (Alexander, 1977). Cold, heat, salinity, heavy metals, waterlogging and drought have been shown to inhibit the establishment of mycorrhizal fungi (Harley and Smith, 1983). For ectomycorrhizae, the absence of organic matter after severe disturbance may limit establishment of the symbiosis despite adequate dispersal (Read, 1984; Allen et al., 1992). Another physical factor reducing survival of fungal propagules is freeze– thaw cycles, which caused introduced ascospores to be ruptured (Zlattner and Gochenaur, 1984). Some abiotic conditions exert their effect in a species-speciﬁc, or even a strain-speciﬁc manner.

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For example, in Cochliobolus sativus, dark-spored strains appear to show improved survival in a desert environment compared to white-spored strains, through protection against ultraviolet radiation (Durrell and Shields, 1960). Many abiotic factors may directly kill dispersed propagules, or limit growth of a newly introduced species. Limited growth may not necessarily eliminate a species after it arrives at a site, but this combined with other limitations could have serious effects. For example, protozoan grazing in combination with slowed growth has been shown to prevent establishment of bacteria in soil communities (e.g., Zaidi et al., 1989). A broad range of microbes are also capable of destroying fungal spores in a variety of habitats. There are numerous reports of destruction of sclerotia by bacteria, antibiotic-producing fungi, and mycoparasites (Whipps et al., 1988). Spores of arbuscular mycorrhizal fungi are also subject to parasitism by other microorganisms (Paulitz and Menge, 1986), and grazing by soil fauna (Ingham et al., 1986a,b; Allen et al., 1987). In a ﬁeld study in an oak–birch (Quercus–Betula) forest, experimentally introduced ascospores were shown to have been lost to grazing by soil arthropods (Gochenaur, 1987). Different life-history stages of fungi are differentially susceptible to loss, for example by bacterial lysis. Ascospores, chlamydospores, sclerotia, and conidia are more likely to survive than germ tubes or hyphae (Schreiber and Green, 1962; Old, 1967). Microbial obligate symbionts represent a special case in terms of reestablishment, since they require the presence of their symbiotic partner. Speciﬁcity varies greatly among mycobionts for a host plant, ranging from extremely high to relatively unspeciﬁc (Allen, 1991). The speciﬁcity of fungal pathogens for their plant hosts also has important consequences for the growth and establishment of fungal populations (Burdon, 1992). Establishment of arbuscular mycorrhizal fungi after disturbance has been studied intensively (Allen, 1988; Allen et al., 1984, 1989, 1992; E.B. Allen et al., 1987; Friese and Allen, 1991, 1993). Both symbionts must disperse independently and then encounter each other in a suitable microsite. A minimum propagule density, and an appropriate physicochemical environment is necessary for the formation of the mycorrhizal symbiosis (Nicholson, 1960). Additionally, factors inherent to phytobiont physiology, such as the nitrogen/phosphorus ratio, are important in establishing the symbiosis (Heijne et al., 1994). Spatial considerations are particularly

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important for the establishment of mycorrhizal fungi in semi-arid habitats, because mycorrhizal hyphae cannot grow long distances to ﬁnd a suitable host root, and plants are typically widely spaced (Allen and MacMahon, 1985). Genetic limitations on the fungal side may also play an important role in the establishment of symbionts. Many mycorrhizal fungi form sexually produced haploid spores and disperse over vast distances, and the establishment of a functional mycorrhizal symbiosis may be limited by the lack of another, compatible haploid spore necessary to induce heterokaryon formation (Wong et al., 1989; Allen et al., 1992). Another pathway to reestablishment is by the survival of residuals. Incomplete eradication of many soil organisms, including arbuscular mycorrhizal fungi, following disturbance has been found to greatly aid in reestablishment (Allen and Allen, 1980). Nests of harvester ants (Pogonomyrmex rugosus), which are devoid of any vegetation, have been shown to have a very high density of arbuscular-mycorrhizal inoculum after they were abandoned by the ants. Reestablishment proceeded from infected root pieces buried in the nest, and by means of hyphal spread (Friese and Allen, 1991, 1993). Following eruption of Mount St. Helens, an explosive volcano in the northwestern United States, arbuscular mycorrhizae and ectomycorrhizae rapidly reestablished in the blast zone and the zone of high ashfall from buried old soils (Allen et al., 1992). However, in the most severely disturbed zone, mycorrhizal reestablishment was considerably slower following the eruption.

MICROBIAL SUCCESSION

As discussed elsewhere in this volume, succession of plants and animals has been well studied; but less is known about succession of soil microorganisms after disturbance. Here we synthesize the major studies on succession of microorganisms, some of which are structural, and some functional in their approach. A theoretical framework for all of these studies begins with the early work of Odum (1969), who theorized that early-successional, disturbed systems are leaky, and that they become tighter in their recycling of nutrients as they mature. Vitousek and Reiners (1975) followed with their hypotheses that the increased biomass of organisms and build-up of organic matter is responsible for the increase and retention of nutrients in late seral

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systems. Just as important is that soil microorganisms change over time, and the functions they perform which increase leakiness in early succession and help to decrease system leakiness over time are critical (Fig. 22.2). Bare Soil plant invasion

Live Plant Mass microbial invasion Detritus recycled nutrients

Microbial Mass

Soil Organic Matter

Fig. 22.2. Nutrients are readily available in newly disturbed soils, but are rapidly immobilized in plant tissue. Only after microbial mass establishes can recycling match plant productivity and soil organic matter accumulate (Insam and Domsch, 1988; Allen, 1993).

Community changes during succession Most perturbation studies show a change in species composition of various groups of microorganisms over time. These changes can occur with very small, ongoing changes such as leaf-fall. Frankland (1966) demonstrated “succession” among fungi colonizing litter of fern petioles. Species of fungi changed over time on decomposing litter, and the species composition was dependent on the plant species that comprised the litter. Frankland (1992) designated this as “substratum succession” as opposed to the colonization of a site following disturbance. This provides for an unusually high diversity of organisms compared with a process like plant succession. There is a distinct seasonal phenology, as leaves are shed preferentially in distinct seasons such as autumn in temperate climates, or the beginning of the dry season in tropical areas. This is superimposed on the changes in substrate chemistry with decomposition, as the differing microorganisms utilize different substrates (Burges, 1958). Superimposed on this again is that each litter type has distinctive decomposers because of different leaf chemistry. Some litter types such as willows (Salix) and cottonwood (Populus) had succession initiated by “sugar” fungi using simple sugars, while in others such as pine litter, succession was initiated by lignicolous fungi because

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the heavy cell walls needed to be broken down ﬁrst, before the sugar fungi had access to a carbohydrate source (Allen, 1977). These differential degradation patterns occur until the litter is reduced to lignins and other complex compounds constituting the soil humus (Allen, 1977; Taylor et al., 1991). Virtually all of these interactions occur within a single year. Since a small perturbation like annual litter-fall initiates species turnover, greater perturbations would certainly be expected to do so. Brown (1958) reported that soil microfungi changed across a dune chronosequence in correlation with changing soil conditions. One year after the Chernobyl accident in 1986, the abundance of some taxa of soil fauna at a distance of 3 km was less than 1% of that at 70 km (Krivolutskii et al., 1989). Other taxa such as Thysanoptera could not be found at all, whereas some, such as beetle larvae, were apparently not affected. However, after 2.5 years most taxa had recovered to previous densities (Krivolutskii et al., 1989). Soil microfungi did not recover even 5 years after exposure to 3.7×107 Bq kg−1 close to the station (Zhdanova et al., 1994). Melanized fungi predominated in the highly irradiated sites at the initial sampling in 1986, and persisted into 1991. Zhdanova et al. concluded that these fungi may be better able to withstand the radiation than the nonmelanized fungi. Several of these melanized fungi are commonly found as airborne spores, including species of Alternaria, Drechslera, and Humicola (Li and Kendrick, 1995) which may recolonize after the death of the dominant soil fungi. Fire is a more common disturbance, which has been studied by soil ecologists. The composition and activity of both soil microorganisms and the organisms that feed on them change in response to ﬁre. Wicklow and colleagues [see Zak (1992) for a review] studied the changes in soil fungi following ﬁre over several years. Fire selectively favored some fungi (particularly ascomycetes), and these fungi persisted for a very long time. In more mature stands, these fungi were good indicators of past ﬁre activity (Wicklow and Whittingham, 1978). Despite the fact that intense heat at a point reduced fungal biomass, this disturbance resulted in an increased fungal diversity. Soil arthropod diversity was as high in raked as in burned plots, and both were higher than control plots (Lussenhop, 1976). This pattern was more pronounced 8 months than 1 month after burning. In contrast, ﬁre caused a reduction in diversity of cryptostigmatic mites, which took 15 years to recover from seven species to the

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original sixteen (Webb, 1994). Mite densities were also reduced by ﬁre by about 80%, and recovered to their original levels in eight months. Thus, ﬁre may cause an initial increase or a decrease in different groups of organisms depending upon their resource needs. Lussenhop (1976) suggested that inorganic nutrients and populations of saprophytes increased temporarily after the ﬁre, creating a food source for the arthropods. Webb’s (1994) mites were dependent upon the heavy litter layer; this was removed by the ﬁre, and the mite population declined afterward. Successional changes have also been observed in arbuscular mycorrhizal fungi. The classic work in this area was on sand-dune succession in England. Nicholson (1960) found that the newly formed dunes were colonized by a non-mycotrophic plant, Salsola kali, subsequently replaced by mycotrophic species. This pattern has been repeated at many sites in very different conditions [for a review, see Allen and Allen (1990)]. Further, Nicholson and Johnston (1979) found that the percentage of roots infected with mycorrhizal fungi increased in later-successional stable dunes, and the number of species of mycorrhizal fungi increased from one in the foredunes to several in the stabilized dunes. In coastal Baja California, the only plant to colonize foredunes, Abronia maritima, was non-mycorrhizal (Sig¨uenza et al., 1996). However, all later seral species were mycorrhizal. In the coastal dunes of Veracruz (Mexico), all colonizing plants were found to be mycorrhizal, but the degree of infection was greater in late-seral plants (Corkidi and Rinc´on, 1997). The later-seral species tended to have greater growth responses to mycorrhizae as demonstrated in greenhouse experiments, so the authors concluded that the late-seral plants were more dependent upon mycorrhizae than the early-seral species. Recolonization of arbuscular mycorrhizal fungi occurred patchily after the eruption of Mount St. Helens in 1980, and was mediated in part by wind but more importantly by animals (Allen, 1987, 1988). In some patches of volcanic tephra, infection levels of plants after ten years were as high as in areas unaffected by the eruption, while other patches still had low levels of infection (Allen et al., 1992). The infection levels of plants were dependent on whether animals had been able to arrive at a particular patch and inoculate the plants with their feces (Allen, 1987). Ectomycorrhizal conifers could not colonize the sterile tephra at all until they became inoculated, and this took a minimum of ﬁve years (Allen, 1988).

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Recolonization by arbuscular mycorrhizae also occurred slowly on mine spoil in Wyoming shrubgrassland (Waaland and Allen, 1987) and in mined alpine meadows in Montana (Allen et al., 1987), taking 5–7 years before root infection was similar in mined and undisturbed areas. However, in both sites the spore density had not achieved predisturbance levels even after 30 years. In early-seral soils, spores of only one or two species of mycorrhizal fungi were present; the number of species had increased to 6–8 after 5–7 years, whereas undisturbed soils had 12 or more species (Allen et al., 1987; Allen and Friese, 1992). Not all sites recover their microbial populations after perturbation. In a Swedish forest, the source of nitrogen fertilizer changed the species composition of microfungi (Arnebrant et al., 1990). Certain species of fungi such as Penicillium spinulosum increased with ammonium nitrate addition, while P. brevi-compactum increased with urea fertilization. Total biomass of saprophytic fungi increased after both sources of nitrogen fertilizer. These changes persisted for 13 years after fertilization, showing a slow recovery and perhaps a persistent change after fertilization. Changes in pH may have been responsible for some of the changes, as it is known that some of the fungal species have afﬁnities for soils of high or low pH. Functional changes during succession A number of researchers have tried to study the changes in soil microorganisms by examining microbial functioning rather than species composition. In Spain, after a ﬁre in a Pinus cembra forest, the ammonifying bacteria increased initially, but became more similar to the unburned treatments after one year (Acea and Carballas, 1996). Conversely, the nitriﬁers were lower in the burned plots initially, and higher in the burned plots after one year. This suggests a greater availability of ammonia and amino acids in the soil after one year, or that the nitriﬁers are slower to recolonize after a burn. Among the organisms responsible for breakdown of carbon compounds, the amylolytic organisms did not change initially, but were lower after one year, and the cellulolytic organisms were lower initially and still lower after one year. These results suggest that starch and cellulose sources were still low in the soil one year after the ﬁre. During succession on an Alaskan ﬂoodplain, the vegetation changes from alder (Alnus tenuifolia) to poplar (Populus balsamifera) at the same time that

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the system becomes nitrogen-limited (Schimel et al., 1996). The loss of a dominant shrub with nitrogenﬁxing root associates (in this case, A. tenuifolia) can explain the nitrogen limitation, but concurrent with these changes were a decrease in nitriﬁcation, and domination of the soil nitrogen pool by ammoniumcontaining compounds. Poplar produces high levels of tannins that inhibit decomposing organisms, which explains these observations. While there was no speciﬁc effect on the nitriﬁers, the lack of mineral ammonium caused by slowed decomposition would explain the reduced nitriﬁcation rates observed. Vitousek et al. (1989) observed that nitriﬁcation and nitrogen turnover are inhibited during the course of some successions, but increased during others. The increases were noted after abandonment from prolonged agriculture, where the intensity of disturbance may regulate nitriﬁcation over time. Other researchers have examined changes in microbial biomass and respiration as an index of activity during succession. For instance, in Bohemia, microbial biomass and soil respiration were higher in a Quercus pubescens forest and a stable meadow than in active and fallow agricultural ﬁelds (Santruckova, 1992). Conversely, microbial activity, a measure of turnover expressed as carbon dioxide production per unit of microbial biomass, was highest in the active ﬁeld. Higher microbial activity may be due to the agricultural practices in this locality and to larger ﬂuctuations of moisture and temperature. In two successional moraines in the Austrian Alps and the Canadian Rocky Mountains (0 to 200 yrs), the metabolic quotient [the ratio of microbial respiration to biomass, also called “microbial activity” (Santruckova, 1992)] decreased with time (Insam and Haselwandter, 1989). Insam and Haselwandter concurred with the hypothesis of Odum (1969) that the ratio of primary productivity to respiration (P/R) will become lower than unity later in succession. One of the stronger indicators of succession may be the ratio of microbial carbon to organic carbon. Insam and Domsch (1988) found that this ratio declined through time (to 50 yr) using chronosequences of recovering mined sites. They postulated that microbial mass increases rapidly in response to plant growth and inputs of labile carbon. The soil organic carbon only builds up slowly, as a greater amount of recalcitrant carbon begins to accumulate. On the other hand, Allen (1993) reported that, during a period of 5 years, the ratio increased, predominantly as a result of increasing

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microbial mass. This increase may depend, in part, on either the ecosystem under study, or, more likely, the means of treating the soil following disturbance. Most models assume that organic matter is lost with disturbance and that the microbes invade rapidly, utilizing whatever carbon input is available (Fig. 22.3A). On the other hand, most restoration practices replace topsoil with organic matter or add mulch. Further, it is known that the microbe groups most likely to establish and grow do not always invade rapidly. This results in a more complex pattern (Fig. 22.3B) where there is an initial large organic carbon pool and rapid growth (1– 10 yr) of microbes using that pool. Only thereafter do the dynamics postulated in Fig. 22.3A take over. This would result in a complex relation of microbial carbon to soil organic carbon (Fig. 22.3C), but would account for both sets of observations. The relationship of microbial structure to function during succession

Fig. 22.3. Changes in microbial carbon and soil organic carbon with time, with and without added topsoil and its organic constituents. (A) A primary succession in which no topsoil remains, showing microbial carbon (diamonds) and soil organic carbon (squares); (B) practices when the topsoil is retained, microbial carbon (diamonds) and soil organic carbon (squares); these two practices lead to quite different ratios of microbal carbon to soil organic carbon as shown in (C) ratio of microbial carbon to soil organic carbon, with topsoil (squares) and without topsoil (diamonds).

A more complex approach was taken by Ingham et al. (1985, 1986a,b, 1989) in relating soil food-web structure to functioning. One study compared meadow, grassland, and forest vegetation. They examined the relationship between the dominant soil organisms and forms of nitrogen (Ingham et al., 1989) and found that the meadow and grassland were dominated by bacteria, while the forest was dominated by fungi. The ratio of fungal to bacterial biomass was 0.1, 0.8, and 8.0, respectively, in the three vegetation types. Concomitant with these microbial ratios were the ratios of soil nitrate to ammonium nitrogen, which were higher in soils dominated by bacteria than where fungi were dominant. As nitrifying organisms, higher levels of nitrate would be expected in bacteria-dominated soils. In another study, Ingham et al. (1986a,b) examined microbial changes and nitrogen mineralizationimmobilization during the growing season. Bacteria and fungi decreased as their grazers increased and soil nitrogen increased, but as grazers decreased and decomposers increased, inorganic soil nitrogen became immobilized. Disturbance also promotes bacteria at the expense of fungi. For instance, tillage increases bacteria, but treatments that do not disrupt the topsoil tend to promote fungi (Moore, 1994). Agricultural soils have lower species diversity and less complex food webs than undisturbed natural ecosystems, but the relation between food-web complexity and nutrient stability still

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needs further exploration (Moore et al., 1996). Ingham et al. (1989) showed that removal of one trophic group, the bacterial-feeding nematodes, was followed by an increase in bacterial-feeding protozoans. This resulted in an increase in fungal biomass and increased soil inorganic nitrogen. This takes one back to the initial questions of how soil microorganisms may mediate the decrease in system leakiness during succession suggested by Odum (1969) and Vitousek and Reiners (1975). Bacteria have higher turnover rates and include the nitrifying functional group, so that higher concentrations of available nitrogen, especially nitrate, occur in bacteria-dominated soils. Nitrate is more readily leached than ammonium, which dominates in undisturbed or late-successional soils with higher fungal biomass. In addition, fungi have a hyphal structure that binds soil particles, which reduces erosion and increases soil stability (Miller and Jastrow, 1992). The hyphae can transport nutrients from patch to patch as they extend (Blair et al., 1992), and do not necessarily return nutrients to the soil at their death. Conversely, when single-cell bacteria die, they release their nutrients to the soil. This provides greater opportunity for freeing nutrients to the soil, which can then be leached. Coupled with more complex cell wall constituents and slower turnover, fungi retain their nutrients for longer time periods than bacteria. Thus, while Vitousek and Reiners (1975) are correct that increasing organic matter and plant biomass increase the nutrient retention of late-successional systems, the increase of soil fungi and a complex food web also contribute to the stability of nutrient cycles. One needs to ask whether the plants are changing during succession because of the microorganisms, or vice versa (Usher et al., 1982). Just as plant ecologists have been studying the mechanisms by which plants succeed each other, soil ecologists are interested in the mechanisms that control the population dynamics of microorganisms. Such attributes as rate of dispersal, reproduction, competitive ability, and resource needs will explain microbial succession as they do plant succession.

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This is particularly true for the complex yet littlestudied below-ground components and the important interactions between microorganisms and plant establishment. This makes it difﬁcult for land managers to make precise prescriptions for land recovery. The remainder of this chapter describes methods and approaches which utilize soil microorganisms to reduce the costs of more traditional reclamation strategies. First priority: site protection The adverse effects of disturbance should ﬁrst be minimized by limiting access or reducing the level of impact. Primarily this is to allow microbes associated with plant roots and organic matter to bind the soil, thereby reducing severe erosion. For construction activities this may mean back-blading temporary roads and access areas rather than traditional scraping. For agriculture it may mean switching from annual crops to perennial crops, or from clean tillage to conservation tillage. For forestry it may mean switching from tractor skidding to trailer skidding with wide-track equipment. Topsoil retention Topsoiling is the process of salvaging the topsoil prior to an anticipated disturbance and respreading it following the disturbance, or transferring it from a site that will be permanently covered (e.g., highway, housing tract) to another disturbed area (Bainbridge et al., 1996). Topsoiling can be an important source for microorganisms, nutrients, and plant propagules (Howard and Samuel, 1979; Allen and Allen, 1980). Its beneﬁts are relatively well established for restoration of ecosystems in moist to semi-arid climates (Bradshaw and Chadwick, 1980; McGinnes and Nicholas, 1980; Barth, 1984; Claassen and Zasoski, 1993), where the salvaged soil can restore biological activity to damaged surfaces, as well as provide a cover (cap) for very coarse or toxic subsoil materials that may have been left at the surface. Biological soil crusts

RECOMMENDATIONS FOR LAND MANAGERS

Although much has been learned about ecosystem structure and functioning in the last 100 years, much remains to be learned about virtually every aspect of ecosystem development and recovery from disturbance.

Cryptobiotic crusts can be very important in the functioning of soil ecosystems. Cryptobiotic crusts of algae and lichens reduce erosion (Fletcher and Martin, 1948; Marathe, 1972; Bailey et al., 1973) and may provide improved conditions for plant establishment

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by increasing inﬁltration and soil nitrogen, and by improving soil structure (St. Clair et al., 1984). Some efforts to encourage algal crust development have been successful (Ashley and Rushforth, 1984; St. Clair et al., 1986). The addition of a soil-crust slurry may be very helpful. Studies have shown this type of treatment may reduce the time for algal-crust recovery from 6–7 years to 6 months (Johansen et al., 1984). Recovery of these crusts can speed succession and revegetation. Water management Water is one of the most important factors limiting growth and recovery of damaged soil ecosystems. Changes in surface soils, loss of organic matter, and the removal of vegetation limit water inﬁltration and retention. Increasing surface roughness enhances the effectiveness of water inputs, and in addition reduces wind speed and facilitates the deposition of sand, seed, and inoculum. Often simply roughening the soil surface may be worthwhile. Deep ripping is often advantageous, and can dramatically improve inﬁltration and air exchange for both roots and microbes (Bainbridge and Virginia, 1990). Other techniques include pitting, imprinting (the creation of small pits), and contouring. Shaping the ground to concentrate available rainfall has been very effective for vegetation establishment in deserts (Evenari et al., 1982; McKell et al., 1979). Microcatchments have been used in North Africa since Roman times, and are now being used in many arid regions (Fidelibus and Bainbridge, 1994). A typical microcatchment might concentrate water from 30 m2 . Microcatchments can reduce salt concentrations at the planting spot, which can prove favorable for microorganisms as well as for plant establishment. If irrigation is feasible, survival and growth of plants and recovery of soil ecosystems can be enhanced. Once plants are established, irrigation can in many cases be tapered off and terminated. Deep or slow-release irrigation appears preferable. Deep-pipe irrigation uses an open vertical pipe to concentrate irrigation water in the deep root zone (Sawaf, 1980; Bainbridge and Virginia, 1990) and provides deep water and a carbon source for microorganisms. Buried clay-pot irrigation uses an unglazed, low-ﬁred clay pot ﬁlled with water to provide a steady supply of water to plants growing nearby, to seedlings, and to rhizosphere microorganisms (Bainbridge and Virginia, 1990).

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Soil management Adding soil amendments is often very important. Comparison with a less-disturbed reference site will show how much organic matter may be needed. Up to 75 t ha−1 have been added in some studies. Full restoration of soil organic matter may take repeated applications over a number of years. The organic matter should be incorporated by spading or tillage. The carbon:nitrogen ratio of organic matter can be adjusted to favor appropriate soil microorganisms. Material with higher carbon:nitrogen ratios is often desirable (e.g., Elkins et al., 1984; Zink and Allen, 1998). More recalcitrant materials such as bark chunks deteriorate more slowly and do not blow away as easily as lighter materials. Reestablishing fungal populations and the animals that graze them help to reestablish more stable cycling of soil nutrients. Low-fertility soils can sometimes be improved by trapping more fertile ﬁne particles carried by the wind. In a study in the Anza-Borrego Desert State Park in southern California, the soil trapped by a fence line was much more fertile than surrounding soils, with ten times more nitrate nitrogen and four times more phosphorus (Bainbridge, 1993). Pitting and ripping treatments also increase surface roughness and help to retain these nutrient-rich soils. On the other hand, markedly increased nitrogen and phosphorus in fertilizer can depress microbial symbionts and prevent natural inoculation (Allen, 1991). This can limit plant adaptation and long-term success. Surface mulches can provide a number of beneﬁts for plants and soil microorganisms. These include improved microclimate (more stable temperatures), increased soil moisture (reduced evaporation, increased inﬁltration, rainwater retention), better aeration, reduced erosion, and an early source of microbial energy. Straw can be used if it is crimped in or tackiﬁed. Rice straw is preferred as it is very durable and less likely to contain weeds. Bundles of rice straw set vertically into the soil have worked well in restoration efforts (Bainbridge, 1996). Natural ﬁber nets over longstemmed straw have been effective for erosion control. This method can also provide seeds if native grasses are used as the mulch. Other mulches include compost, bark, wood chips, ﬁber nets, sawdust, agricultural waste, and even inert materials such as gravel and coarse sand. Green manures can be used to hasten recovery of large areas with regular rainfall; typically, leguminous

SOIL MICROORGANISMS

plants are used to restore soil nitrogen and improve soil ecosystem health and complexity. Seeds are usually inoculated with appropriate symbionts, and seeding rates are high in order to ensure a full stand. The green manure is commonly tilled or spaded in before it has set seed. Plant management Plants and soil microorganisms are intimately related, and local seed collection should be given priority in many projects. Seeds from local stands are most appropriate because local genotypes are most likely to survive and successfully reseed (Lippitt et al., 1995). Just as important, mycorrhizal and other soil microbes are generally adapted to local plant populations (Weinbaum et al., 1996). However, seed production can be erratic, with a high proportion of non-viable seeds. Thus, locally adapted seeds of a particular species are often unavailable from wild stocks when needed. Seed quality can be assessed by non-destructive X-ray analysis, dissection, and germination tests (Lippitt and Bainbridge, 1993). For most restoration projects, it is desirable to harvest seeds from a diverse population; at least 15 and preferably 50 plants should be utilized. It is also important to assay the existing soil microorganisms and soil health. Even a crude understanding of soil fungi, soil microarthropods, and other microorganisms can be helpful in planning restoration strategies. Preliminary data indicate that soil microbial activity is the best indicator of restoration potential for a disturbed site. Soil ecosystem health is closely related to plant recovery and vegetation restoration often requires transplanting. Inoculation with suitable microorganisms may improve survival and growth of speciﬁc plants and thus allow for the restoration of microbes, particularly poorly dispersing species, to the site. This procedure is particularly suitable when vertical placement of speciﬁc components is desired. Fast, deep rooting may be achieved by early planting of nursery stock, or using deep containers to protect and encourage tap root development (Bainbridge et al., 1995). Using this approach allows for placement of appropriate symbiotic microorganisms to improve nutrient availability and plant performance (Aldon, 1978; Trappe, 1981; Read et al., 1985; Bloss, 1986; Virginia et al., 1986; Allen, 1988, 1991). Moreover, double inoculation of leguminous plants with both rhizobia and mycorrhizae can be undertaken if ﬁeld

537

populations of microsymbionts are severely depleted or absent (Carpenter and Allen, 1988) or if the spatial distribution is disturbed (Virginia et al., 1986; Allen, 1988). Improving the structure of organic matter may be even more important than inoculation. Providing surface compost, bark, or some other recalcitrant carbon source may improve fungal development. Reducing compaction, improving inﬁltration, and increasing soil moisture storage all make conditions more favorable for soil symbionts and for the organisms that transport them. If transplants are used, the surface can be stabilized with mulch, but the roots can be placed deep to prevent immobilization of critical nutrients by the fungi inhabiting the mulch. Islands of vegetation Concentrating resources to create resource islands may provide greater beneﬁts than less intensive treatments over a larger area (Allen and MacMahon, 1985; Allen, 1988). These islands can provide seed and inoculum for surrounding areas. These resource islands apparently play a major role in the development of many ecosystems. Transplanting clumps of shrubs into the center of barren areas is a low-cost method of promoting the formation of resource islands. General recommendations Restoring soil health is made difﬁcult by limited understanding of many key processes in the belowground compartment of ecosystems. Yet there is not a more critical part of the ecosystem, and neglecting to consider soil conditions is the most common cause of failure of sites to recover from disturbance. Developing information on a carefully chosen reference site, and careful planning, can improve results. A primary goal of restoration includes reestablishing the structure and function of the soil ecosystem. Activities such as deep ripping, ground shaping, compost, and mulch can aid recovery and initiate succession. Establishing islands of vegetation can act as nucleating agents to return needed soil organisms to the disturbed sites. CONCLUSIONS

Soil microorganisms are critical components of ecosystems and soil microbial communities are dramatically altered by disturbance. While some microorganisms

538

will always reinvade disturbed areas, there is no one-toone relationship between the pre-disturbance organisms and post-disturbance organisms. Many organisms such as large-spored species of mycorrhizal fungi fail to disperse or disperse slowly back onto a site. This results in a notably reduced biodiversity. Whether or not increasing biodiversity implies improved ecosystem functioning is highly debatable (Beare et al., 1995; Bruns, 1995; Huston, 1997). However, many of the ecosystem functions catalyzed by soil microbes are diminished by disturbance – particularly symbiotic activity. Further, the activity of the differing trophic groups may not be well matched. Speciﬁcally, plant productivity is often increased by fertilization, irrigation, and pesticide use, whereas the decomposition rates regulated by the altered populations of soil microbes are reduced. Active restoration of soil organisms and soil processes deserves as much attention as the engineering and aboveground planting involved in a recovery effort, and are essential for those practices to succeed.

ACKNOWLEDGEMENTS

This review was written with the support of the National Science Foundation Conservation and Restoration Biology Program, the California Department of Transportation and the United States Department of Energy, Program for Ecosystems Research.

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Chapter 23

RESPONSES OF CARBON AND NITROGEN CYCLES TO DISTURBANCE IN FORESTS AND RANGELANDS Dale W. JOHNSON and Bradley SCHULTZ

INTRODUCTION

Nutrient cycles are almost inevitably affected by ecosystem disturbance. The nature and magnitude of response to disturbance vary with the kind of disturbance, nutrient, and ecosystem type. The responses of the nitrogen cycle to disturbance are often most pronounced and have certainly received the most study. Nitrogen is unique among plant nutrients in that, with few exceptions (e.g., Dahlgren, 1994), its primary source is from the atmosphere rather than the soil; it does not accumulate in inorganic, exchangeable phases to a signiﬁcant degree, and it is intimately tied to organic carbon (Johnson, 1992). For these reasons, nitrogen is by far the most commonly limiting nutrient in terrestrial ecosystems; and the competition between soil heterotrophic organisms, plants, and nitriﬁers for soil available nitrogen is intense. Figure 23.1 depicts nitrogen cycling, with an emphasis on soil processes. Traditionally, it has been assumed that heterotrophs are the most effective competitors for nitrogen, plants second, and nitriﬁers least effective (reviewed by Johnson, 1992). Over the longer term, however, the situation is not so clear; and it is possible that both long-lived plants like trees and nitriﬁers may be more competitive than previously suspected. Recently, Stark and Hart (1997) challenged the assumption that nitriﬁers are poor competitors for nitrogen by showing through the use of 15 N isotope dilution that, although nitrate concentrations were very low in many forest soils, nitriﬁcation and microbial nitrate uptake rates were very high, seemingly dispelling the notion that nitriﬁers are poor competitors for soil nitrogen. Meanwhile, Kronzucker et al. (1997) showed that white spruce (Picea glauca) had a strong preference for ammonium in solution-culture studies, and hypothesized that this

is a reason for poor seedling regeneration in clearcuts where net nitrate production is high. As noted by Eviner and Chapin (1997), however, both Kronzucker et al. and Stark and Hart were studying only part of the system: plant roots were absent in the soil cores of Stark and Hart, whereas microbes were absent in the solution-culture studies of Kronzucker et al. (1997). Thus, serious questions remain as to the competitive abilities of plants and microbes for nitrogen in intact systems before any wholesale revision of nitrogen cycling models is warranted. Site disturbance can interrupt the competition for nitrogen in a number of ways. Harvesting, for example, causes increased nitrogen losses from forests both by direct removal in biomass and by increased leaching of NO−3 . Direct removal of nitrogen in biomass is often the dominant form of nutrient loss following harvest, especially whole-tree harvesting (e.g., Weetman and Webber, 1972; Boyle et al., 1973; Johnson and Todd, 1987; Mann et al., 1988; Hendrickson et al., 1989).

Fig. 23.1. Conceptual model of the nitrogen cycle. Width of arrows indicates approximate magnitude of ﬂuxes. (Modiﬁed from Johnson, 1992).

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Direct nutrient removal is rather easily characterized and therefore not subject to much scientiﬁc controversy; post-harvest nitrate leaching and runoff in stream waters has generated considerably greater concern and scientiﬁc investigation, especially after the studies on clear-cutting on the Hubbard Brook catchment in New Hampshire (Likens et al., 1969). In this study, a forested catchment was clear-cut and plant growth was prevented with herbicide for three years. During the period when herbicide was applied, stream-water runoff and NO−3 losses increased dramatically, resulting in considerable nitrogen loss from the ecosystem (499 kg ha−1 : Likens et al., 1978). Nitrate concentrations in stream-water also exceeded drinking-water standards over the period of herbicide application (Likens et al., 1969). The Hubbard Brook results created a furor over the practice of clear-cutting and set in motion a series of follow-up studies of harvesting in other forest types (Brown et al., 1973; Aubertin and Patric, 1974; Richardson and Lund, 1975; see also review by Vitousek and Melillo, 1979). Results from the latter studies showed that the losses of nitrate from the clearcut and herbicide treatment at Hubbard Brook were much larger than average losses from forests where herbicide was not applied (Likens et al., 1978; Vitousek and Melillo, 1979). In addition to stimulating several studies on the effects of forest harvesting on NO−3 losses, the harvesting studies at Hubbard Brook stimulated some basic research on the mechanisms of nitrogen retention and loss in forest soils under various disturbances, including stem girdling (Edwards and Ross-Todd, 1979; Johnson and Edwards, 1979) and trenching (Vitousek et al., 1979). The largest of these was a multisite comparison of nitrogen cycling after root trenching designed to prevent root uptake (Vitousek et al., 1979). Eight processes were identiﬁed which could delay or prevent nitrogen losses following disturbance: (1) nitrogen immobilization, (2) NH+4 ﬁxation in clays, (3) NH3 volatilization, (4) plant nitrogen uptake, (5) lag in nitriﬁcation, (6) denitriﬁcation, (7) lack of water, and (8) NO−3 sorption. The authors found that four of these processes – nitrogen uptake by vegetation, nitrogen immobilization, lags in nitriﬁcation, and lack of water – were most important in the sites they studied. With the recent emergence of the nitrogen-saturation issue (i.e., negative effects of excess inputs of nitrogen: Aber et al., 1989), there has been renewed interest in the effect of disturbance on NO−3 loss.

Dale W. JOHNSON and Bradley SCHULTZ

Rangelands are subject to many disturbances that alter nutrient cycling. Anthropogenic activities that almost always disturb rangelands include mineral exploration and mining, the construction of transportation and utility corridors, military training and testing, repeated travel by off-road vehicles, water developments and delivery systems, and infrastructure to manage livestock grazing (e.g., pasture fences, corrals). Each disturbance type results in the direct removal of vegetation, the death of root systems, and usually the compaction and/or removal of topsoil. The disruptions to nutrient cycling are obvious and almost always long-term, especially with high intensities of disturbance. Herbivores graze rangelands extensively but not always intensively. Grazing may alter vegetation composition and structure, and above-ground and below-ground productivity, but seldom eliminates the vegetative resource. The potential of herbivores to disturb nutrient cycles directly and indirectly, therefore, is less obvious but in need of a detailed synthesis. The rangelands section of this chapter (pp. 554–565) will focus on how herbivores, primarily large generalist grazers, directly and indirectly alter and disturb nutrient cycles. The concepts can be applied to virtually any type of disturbance. Herbivores inﬂuence and interact with nutrient cycles in numerous ways, including: excretion; enhanced nutrient retention; faster nutrient loss; alteration of nutrient supply and cycling rates; increased nutrient uptake by plants; facilitation of energy and nutrient ﬂow; and modiﬁcation of botanical composition and structure (Detling, 1988 and references therein). Grazing disturbs individual plants by removing and/or breaking plant parts; however, grazing may or may not be an ecological disturbance to rangeland systems. Most rangelands have evolved with some level of grazing, but grazing is an ecosystem disturbance only when site-speciﬁc defoliation and trampling thresholds are exceeded and new ecological conditions are superimposed upon the landscape (Mack and Thompson, 1982; Friedel, 1991; Laycock, 1991). Most research about nutrient cycling and grazing on native rangelands has addressed how grazers interact with the drivers (Table 23.1 lists important deﬁnitions) for the processes of organic matter (OM) accumulation and decomposition, not how grazers may adversely affect nutrient cycles. One key driver is the maintenance of inputs to the soil of organic carbon (OC), because organic carbon in the soil supplies the energy that drives microbial decomposition (Killham, 1994).

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS Table 23.1 Deﬁnitions of soil terms related to nutrient cycling in rangeland ecosystems Term

Deﬁnition

Driver

system component that initiates, promotes, or continues an ecological process

Organic carbon

carbon bound in live or dead plant and animal matter

Organic matter

live or dead plant and animal matter

Organic nitrogen

nitrogen bound in live or dead plant and animal matter

Soil organic carbon

carbon in live or dead plant or animal matter located in mineral soil; primary energy source that promotes microbial decomposition

Soil organic matter

live or dead plant and animal matter located in mineral soil

Grazing changes the spatial and temporal inputs of organic carbon (Briske, 1991), which inﬂuences rates of nutrient supply and cycling. Spatial and temporal changes in the drivers that promote the accumulation and decomposition of soil organic matter (SOM) will alter soil microbe populations, nutrient mineralization, and nutrient availability for plants. Other important drivers that promote microbial decomposition and nutrient cycling are moist soil and warm soil temperature. A full review of the effects of disturbance on nutrient cycling in terrestrial ecosystems is well beyond the scope of this chapter. Rather, this review will focus upon speciﬁc disturbances involving either the removal of plant biomass or reduced plant uptake – primarily harvesting in forests and herbivores in rangeland ecosystems. In the forest section, we will focus upon the literature published since 1980 (focusing upon harvesting and nitrate losses, which most of that literature concerns) and compare these results with the thinking that emerged from intensive studies during the 1970s. In particular, we will ask whether the conclusions drawn from those earlier studies and in reviews held up in the face of newer data, and whether there are additional factors affecting nitrogen retention which have been uncovered since that time. In the rangeland section, we will focus upon the effects of grazing by generalist ungulates. 1

547

EFFECTS OF DISTURBANCES IN FORESTS

Effects of forest harvesting A number of studies of the effects of harvesting on nutrient removal have been conducted over the last 15 years. These studies have included both removal of biomass, which nearly always has the largest harvesting effect on nutrient budgets1 (Johnson and Todd, 1987; Johnson et al., 1988; Mann et al., 1988; Hendrickson et al., 1989), and effects on nutrient leaching and water quality (especially NO−3 ). Many studies after 1980 on harvesting and nutrients focused on the effects of residue removal or whole-tree harvesting (removal of most aboveground biomass: WTH) as against conventional sawlog harvesting (removal of commercially important boles only: SAW) on nutrient leaching and water quality. Although several earlier studies noted the disproportionate increases in nutrient removal in biomass with WTH as compared to SAW (e.g., Weetman and Webber, 1972; White, 1974; Boyle et al., 1973), only recently have the effects upon leaching been considered. A full review of the literature on this subject is beyond the scope of this paper. This brief review will focus on leaching and water-quality effects of forest harvesting, with an emphasis on nitrate losses as a follow-up on the earlier syntheses by Vitousek et al. (1979) and Vitousek and Melillo (1979). The effects of harvesting on direct removal of nutrients will not be considered in this review; the interested reader is referred to papers by Johnson et al. (1988), Mann et al. (1988) and Federer et al. (1989) for useful summaries. Sawlog harvesting and nutrient losses As a result of the clear-cutting studies at Hubbard Brook by Likens et al. (1969), Hornbeck et al. (1986) compared the effects of cutting in strips versus blocks on hydrology and nutrient losses in stream water at Hubbard Brook. The strip-cutting treatments involved cutting strips 25 m wide across elevation contours in one catchment, leaving alternate strips uncut. During the ﬁrst 10 years since harvesting, the nitrogen losses from the strip-cut catchment (22 kg ha−1 ) were substantially less than in the block-cut catchment (59 kg ha−1 ). Strip-cutting resulted in lower losses of calcium and

Over the full rotation of a forest, nutrient leaching can rival nutrient removal in biomass; however, the effects of harvesting on leaching losses are usually quite short-lived and far less important over the long run than are those of direct nutrient removal.

548

potassium also, but not to the same degree as for nitrogen. In no case were the losses of nitrogen in stream water considered to be a threat to the long-term nutrient capital of these forests; nitrogen removals in biomass were more signiﬁcant. Jewett et al. (1995) reported the results of SAW on soil solution and stream water in two catchments with both coniferous forests [red spruce (Picea rubens) and balsam ﬁr (Abies balsamea)], and deciduous forests [beech (Fagus grandifolia) and sugar maple (Acer saccharum)] in New Brunswick, Canada. They found that harvesting (approximately 90% of total forest on each catchment) increased concentrations of nitrate in soil solution, as expected, with higher concentrations in hardwood than in conifer stands. Stream-water nitrate increased after harvesting also, but the increases were somewhat smaller and more prolonged than those in the soil solution. There were also increases in the leaching losses of calcium, magnesium, phosphorus and potassium. In no case were the losses considered signiﬁcant for either water quality or the long-term productivity of the site. Several studies in Europe, including the United Kingdom, showed that SAW harvesting caused increases in nitrate in the soil solution and/or stream water but of much lesser magnitude than in the original Hubbard Brook studies. In addition, many of them showed that the reduction of canopy scavenging (dry deposition of ions and subsequent leaf wash-off) upon harvesting caused signiﬁcant changes in the chemistry of the soil solution and/or the stream water. Adamson et al. (1987) studied the effects of SAW (slash left on site) on stream-water chemistry in a plantation of Sitka spruce (Picea sitchensis) in Kershop Forest, Cumbria, United Kingdom. They found that harvesting caused increased water yield and increased concentrations and total export of nitrate, ortho-phosphate and potassium. Concentration and ﬂuxes of ions with a predominantly − + atmospheric origin (SO2− 4 , Cl , and Na ) decreased after harvesting due to reduced canopy scavenging of these ions by forest canopies. None of the observed changes were of sufﬁcient magnitude to cause concern for either water quality or site nutrient status. Wiklander et al. (1991) studied the effects of blowdown and removal of fallen trees on stream-water chemistry in a catchment in southern Sweden. This site is subjected to relatively high inputs of nitrogen through atmospheric deposition (10 kg N ha−1 yr−1 in bulk precipitation), creating concerns over the potential contamination of stream and ground waters with nitrate

Dale W. JOHNSON and Bradley SCHULTZ

after harvesting. Over a period of approximately seven years, about one-third of the forest vegetation in the catchment blew down [primarily Norway spruce (Picea abies)] and was removed. The blow-down sites were replanted within one to two years of blowdown. The type of removal was not speciﬁed, but we assume it to be conventional harvest (slash left on site). The blow-down–clear-cut disturbance caused a pattern of nitrate loss typical of other sites: an increase in stream-water nitrate, with nitrogen losses increasing to 18 kg N ha−1 yr−1 in years 3–4 followed by a decrease to approximately 3 kg N ha−1 yr−1 in years 7– 8. The initial increase was attributed to increased nitrogen mineralization and nitriﬁcation after harvest, as is typically found, and the subsequent decrease was attributed to increased uptake by regenerating vegetation. Wiklander et al. noted that reduced dry deposition of nitrogen to forest canopies after blowdown and harvesting was also a signiﬁcant factor in the effects of disturbance on nitrogen budgets at this site. Stevens and Hornung (1988) studied the effects of conventional harvesting on nitrate leaching from a Sitka spruce/Norway spruce plantation in Beddgelert, Wales. They found the expected increases in inorganic nitrogen (mostly NO−3 ) concentrations in soil solutions from the C horizons, but decreases in the organic (O) horizons and no change in the E (eluviated) horizon or the Bs horizon (where oxides of iron and aluminum from the E horizon have been deposited). They attributed the concentration decreases in the O horizons to increased water ﬂux, and the increases in the C horizon to reduced root uptake. Leaching of nitrate increased from approximately 10 to 70 kg ha−1 yr−1 after harvesting and was attributed to active nitriﬁcation (despite the acidic soil) and reduction in nitrogen uptake. Reynolds et al. (1995) summarized the results of studies on the effects of conventional clear-cut harvesting on stream chemistry in Plynlimon and Beddgelert, Wales. At Plynlimon, an entire catchment of Sitka spruce was clear-cut, whereas in Beddgelert, two catchments of Sitka spruce were partially clear-cut (62 and 28% on an area basis). In each case, streamwater concentrations of nitrate and potassium increased rapidly after harvesting and remained elevated for periods that varied with nutrient and location. At Plynlimon, stream-water nitrate reached a maximum of 3.2 mg N °−1 after one year and declined to control levels (0.5 mg N °−1 ) after ﬁve years. At Beddgelert,

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

concentrations of nitrate in stream water remained elevated compared to the control for three years, and then declined to control levels. The responses in nitrate were attributed to increased nitrogen mineralization and nitriﬁcation in the soil after harvesting, and in each case the increases in nitrate caused temporary increases in acidity. As in the previous studies by Adamson et al. (1987), clear-cutting resulted in reduced stream-water − + concentrations of SO2− 4 , Cl , and Na due to reduced canopy scavenging at Beddgelert and, to a lesser extent, at Plynlimon. None of the increases in stream-water nitrate were seen as posing serious threats to water quality. In a particularly interesting study relative to the nitrogen-saturation issue, Nohrstedt et al. (1994) and Ring (1995) reported the results of a clear-cutting study in a stand of Scots pine (Pinus sylvestris) in central Sweden. The stands had been previously fertilized with ammonium nitrate over a 20-year period. The total additions of nitrogen to the various plots were 0, 360, 720, 1080, 1440 and 1800 kg N ha−1 . Preharvest inventories showed that the content of carbon and nitrogen in the forest ﬂoor (organic or O horizon) of the treatment showing the greatest growth response (1440 kg N ha−1 ) were about double those of the unfertilized plots. Fertilization ceased in 1981, and the plots were subjected to SAW harvest in 1987. The slash was carefully placed upon the plots of origin after harvesting. Prior to harvesting, organic nitrogen was the dominant form of nitrogen in soil solution in all treatments (Nohrstedt et al., 1994). During the ﬁrst two years after harvesting (when precipitation was abnormally low), concentrations of all forms of nitrogen in the soil solution decreased compared to preharvest conditions. During the third year, nitrate concentrations increased in the most heavily fertilized plots. Denitriﬁcation was also measured and was very low (<0.1 kg ha−1 yr−1 ) in all cases. In a more detailed and longer-term study at the same site, Ring (1995) calculated rates of nitrogen leaching for the ﬁrst ﬁve years after harvesting as follows: (1) 2–3 kg N ha−1 yr−1 in the uncut reference plots, the clear-cut, unfertilized plots, and the plots receiving 360 kg N ha−1 yr−1 ; (2) 4–5 kg N ha−1 yr−1 in the plots receiving 720 and 1080 kg N ha−1 yr−1 ; (3) 14 kg N ha−1 yr−1 in the plots with 1440 kg N ha−1 yr−1 ; and (4) 37 kg N ha−1 yr−1 in the plots with 1800 kg N ha−1 yr−1 . The duration of the elevated nitrogen leaching from these plots is not yet known, but typically is on the order of 6–10 years (Wiklander, 1983).

549

The studies of Nohrstedt et al. (1994) and Ring (1995) show how tightly accumulated nitrogen is retained in forest systems which have been fertilized. There is certainly an increase in nitrate leaching following harvesting in the more fertile (i.e., the more heavily fertilized) sites, but the nitrogen losses are only a small fraction of what has been previously accumulated within the system. It does not appear that these systems will move toward some baseline, nitrogen-deﬁcient condition within the time frames that humans normally consider, but rather will remain “permanently” enriched. Thus, it appears that the fertilizer nitrogen previously accumulated in these systems was very stable and not subject to signiﬁcant release following even a major disturbance such as clear-cutting. It is quite possible that fertilizer nitrogen in these studies was abiotically ﬁxed by soil organic matter (e.g., Foster et al., 1985; Axelsson and Berg, 1988). In any event, these sites had clearly not become nitrogen-saturated even after massive doses of fertilizer nitrogen. Whole-tree harvesting and nutrient losses Early studies by Weetman and Webber (1972) and Boyle et al. (1973) showed that disproportionate nutrient removals in biomass with whole-tree harvesting (WTH) would cause long-term degradation of nutrient status of the site. Many subsequent papers and symposia addressed this issue (e.g., Leaf, 1979). In this section, we will focus on the effects of WTH on leaching losses, where there have been some interesting and often counterintuitive results bearing upon the general issue of disturbance and nutrient cycling. In particular, several studies have found that WTH results in less nitrate leaching than SAW, in contrast to earlier predictions based upon the knowledge base up to 1980 (e.g., Vitousek, 1981; Aber et al., 1978). The factors responsible for this interesting turn of events include plant uptake, which is often reduced simply by the interference of the slash with vegetation establishment in most cases, and mineralization, which is facilitated by the more favorable microclimate under slash. Interestingly, nitrogen release from the slash itself does not seem to be a signiﬁcant factor. The following paragraphs review in more detail the various studies where such results have been obtained. Vitousek and Matson (1985) examined the effects of harvest intensity and site preparation on availability of soil nitrogen and losses of nitrogen from a plantation

550

of loblolly pine (Pinus taeda) in the Piedmont of North Carolina. The main treatments included both SAW and WTH. Within these treatments, two additional site-preparation treatments were imposed: (1) lowintensity [roller-drum chopping: (RDC), and (2) highintensity (shear, or stump removal by bulldozer, piling of residues, and disking: (SPD)]. In addition to this, the SPD plots received herbicide to control hardwood sprouting. Vitousek and Matson noted that nitrate concentrations in the soil solution were not well correlated with net nitrogen mineralization or nitriﬁcation, but only with the size of the soil nitrate pool in late summer prior to the onset of leaching in autumn. The effects of herbicide treatments in increasing the nitrate concentration in the soil solution were quite clear, however, especially in the SPD plots. Thus, the data on nitrate leaching alone would suggest that uptake by vegetation was a major factor controlling nitrate losses. However, the tracer studies showed that microbial immobilization within the soil was the most important process controlling nitrogen availability and losses in these systems. In addition, the removal of residues with high C/N ratios in the SPD treatment reduced nitrogen immobilization and increased the availability of nitrogen for loss by leaching. Hendrickson et al. (1989) tested two alternative hypotheses on the effects of whole-tree harvesting on leaching losses of nitrogen that were posed in the late 1970s. Aber et al. (1978) and Vitousek (1981) had predicted greater leaching of nitrate with WTH than with SAW, because of reduced nitrogen immobilization in decomposing slash with WTH. On the other hand, Wells and Jorgensen (1979) noted that slash had a signiﬁcant component of material with a low C/N ratio (needles, ﬁne twigs) which could cause higher leaching of nitrogen with SAW compared to WTH, where such materials would be absent. In order to test these alternative hypotheses, Hendrickson et al. (1989) studied the effects of two harvesting intensities on nutrient budgets and nitrate losses in mixed hardwood–conifer forests in Ontario, Canada. In the SAW treatment, all aboveground woody biomass with a diameter at breast height (dbh) > 9 cm was removed, and in the WTH treatment all woody biomass greater than 1.3 m in height was removed. The harvests were conducted after leaf-fall. They found that the hypothesis of Wells and Jorgensen (1979) was supported: nitrate in the soil solution was elevated in the SAW treatment only, and was not affected by the WTH treatment. Even in the SAW

Dale W. JOHNSON and Bradley SCHULTZ

treatment, however, the leaching losses amounted to only 1.6 kg ha−1 yr−1 compared to 0.3 kg ha−1 yr−1 under WTH and 0.3 kg ha−1 yr−1 in the control treatment. Several other studies of the effects of WTH on nitrate leaching produced similar results. In a follow-up to earlier studies of conventional harvesting on a Sitka spruce plantation in Beddgelert forest in Wales, Stevens and Hornung (1990) studied the effects of WTH versus SAW harvesting on nutrient losses. They monitored concentrations of inorganic nitrogen in the soil solution for two years before and for four years after harvesting. In both treatments, harvesting resulted in increases in nitrogen concentrations (again mostly nitrate) for up to 14 months. After this time, the concentrations dropped steadily in the WTH treatment, approaching detection limits (0.1 mg N °−1 ) four years after harvesting. In contrast, concentrations of inorganic nitrogen in the soil solution under the SAW treatment remained elevated (1–4 mg N °−1 ) throughout the duration of the study. Leaching of inorganic nitrogen from slash in the SAW treatment was not the cause of the prolonged elevation of nitrogen leaching in the SAW treatment: concentrations of inorganic nitrogen in throughfall beneath slash were lower than in bulk precipitation and could not account for the observed differences in soilsolution concentrations. The differences in nitrogen leaching with treatment were attributed to reduced vegetative cover in the SAW treatment and to a lack of a nitrogen source in the slash under the WTH treatment. Regeneration biomass was 50 to 100% greater in the WTH than in the SAW treatment 2– 3 years after harvesting, apparently as a result of simple physical interference with regeneration. Similarly, the nitrogen content of vegetation in the WTH treatment was 40 to 200% greater than in the SAW treatment. Nitrogen uptake by regenerating biomass more than accounted for the differences in nitrogen leaching. Budget calculations indicated that the additional removal of nitrogen with WTH had no deleterious effect on site nitrogen status. Nitrogen removal in biomass with WTH was 428 kg ha−1 , compared to 128 kg ha−1 with SAW. Given the large pools of soil nitrogen (9400 kg ha−1 ) and relatively high nitrogen deposition rates (10–14 kg ha−1 yr−1 ), no long-term deleterious effects of WTH on nitrogen status of the site were envisioned. Rosen and Lundmark-Thelin (1987) studied the effects of slash piles in a clear-cut site in central Sweden on the chemical composition of throughfall and soil solution in the O, E, and B horizons. This study was

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

undertaken because of the effects of modern harvesting techniques on slash distribution. With manual removal of limbs, slash is left relatively evenly distributed over the site, whereas with modern mechanical equipment slash is left in piles covering about 10–15% of the harvested area. Rosen and Lundmark-Thelin found reduced concentrations of inorganic nitrogen in throughfall beneath slash piles as compared to precipitation but increased concentrations of organic nitrogen. In soil solutions beneath slash piles, on the other hand, they found signiﬁcantly increased concentrations of all forms of nitrogen in solution (NH+4 , NO−3 , and organic N) as compared to soil solutions taken from open areas. This corresponded to greatly increased concentrations of exchangeable NH+4 in the humus beneath slash piles. The positive effects of slash on availability of soil nitrogen and leaching of soil-solution nitrogen were attributed to more favorable moisture conditions. Rosen and LundmarkThelin also noted that vegetation establishment was inhibited in slash piles, and thus uptake of nitrogen by vegetation would also be reduced, as noted by Hendrickson et al. (1989) and Stevens and Hornung (1990). However, because of the type of lysimeter used (ﬁlled-in PVC pipes, isolated from roots), the direct effects of vegetation on the differences observed were not measurable. In a later study of the effects of slash on soil leaching, Staaf and Olsson (1994) measured the effects of SAW, WTH, and complete tree harvest (CTH), which includes the removal of all aboveground biomass together with stumps. They found that WTH caused lower concentrations of NO−3 , NH+4 , and K+ in soil solutions below the rooting zone than SAW. The CTH treatment caused a large increase in NH+4 concentrations during the ﬁrst year after harvesting, and this was followed by an increase in NO−3 and reduced pH as the NH+4 initially released underwent nitriﬁcation. Staaf and Olsson (1994) suggested that WTH with minimal soil disturbance could be used to reduce the negative effects of nitrogen saturation in regions with high nitrogen deposition – a stark contrast indeed to concerns over nitrogen depletion with WTH in less polluted regions (e.g., Boyle et al., 1973). They noted that conventional (SAW) harvesting, on the other hand, seemed to lead to greater NO−3 leaching and waterquality problems. Perhaps the biggest surprises in the effects of harvesting on nutrient leaching came with the studies of Dale Cole and his students on red alder (Alnus rubra)

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in the northwestern United States. Van Miegroet and Cole (1984) had previously shown that the presence of red alder, a nitrogen-ﬁxing species, caused very large amounts of nitrate to be leached compared to Douglas-ﬁr (Pseudotsuga menziesii). This raised the question of how red alder systems would respond to harvesting. In the ﬁrst of two studies to address this question, Bigger and Cole (1983) reported the results of studies on the effects of harvesting intensity on nutrient budgets in red alder and Douglas-ﬁr stands at Pack Forest in western Washington. Previous studies in Douglas-ﬁr forests of low site quality had shown only slight increases in leaching of nitrate following harvesting (Cole and Gessel, 1965). Adjacent stands of these species in sites of both high and low fertility were harvested at three intensities: (1) SAW, (2) WTH, and (3) WTH plus forest-ﬂoor removal. They found that increased intensity of harvesting caused disproportionate increases in nutrient removal in the biomass, as is well known from other studies. They found slight increases in leaching of nitrate after harvesting in the Douglasﬁr plots, as in previous studies in low sites (Cole and Gessel, 1965). Surprisingly, however, they found large reductions in leaching of nitrate following harvesting in red alder, in sites both of low and high quality. The reduced leaching of nitrate following harvesting in red alder resulted in reduced leaching of base cations also, which substantially offset the increased export of nutrients in biomass. Van Miegroet et al. (1992) reported the results of a study at another site at Cedar River, Washington, where adjacent stands of red alder and Douglas-ﬁr were harvested and replanted with Douglas-ﬁr and red alder on both sites. Thus, there were conversions of alder to alder, alder to ﬁr, ﬁr to ﬁr, and ﬁr to alder. As in the previous studies at Pack Forest (Bigger and Cole, 1983), harvesting caused dramatic reductions in leaching of nitrate in the former red alder site over the ﬁrst three years. Replanting with red alder did not inﬂuence the nitrogen dynamics of either the former red alder or the former Douglas-ﬁr plots over this time period. The authors found no differences in nitrogen mineralization, moisture, or temperature that could account for the observed differences in nitrate leaching. Thus, the studies by Van Miegroet et al. (1992) provide no evidence for microbial immobilization as a cause of the decrease in nitrate leaching after harvesting in red alder; and plant uptake (which is greatly reduced immediately after harvesting) certainly

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cannot be a factor. Increased denitriﬁcation is, as yet, an unexplored possibility, but seems unlikely as a major factor in the reduction in nitrate leaching from the forest ﬂoor and upper soil horizons. Nitrogen ﬁxation itself is almost certainly reduced after harvesting in red alder, and, although not mentioned by either Bigger and Cole (1983) or Van Miegroet et al. (1992), this reduction seems a likely cause of the reduced nitrate leaching. This in turn raises an interesting question as to the route which excess ﬁxed nitrogen takes in these systems. Is it taken up by the red alder and cycled to the forest ﬂoor, where mineralized nitrogen exceeds biological demand and is nitriﬁed? Or does the nitrogen cycle operate as in non-ﬁxing forests, the excess ﬁxed nitrogen being released directly from the nodules to be nitriﬁed in situ? Mann et al. (1988) summarized the results of SAW and WTH studies in a variety of sites in the eastern and northwestern United States. These sites included a mixed oak forest (with Acer rubrum, Carya spp., Quercus alba, Q. prinus and other Quercus species) near Oak Ridge, Tennessee (results previously described in detail by Johnson and Todd, 1987); a mixed deciduous forest at Coweeta, North Carolina (with Acer rubrum, Liriodendron tulipifera L. and Quercus spp.); a loblolly pine forest at Clemson, South Carolina; a slash pine (Pinus elliottii) forest at Bradford, Florida; the red alder and Douglas-ﬁr forests described above (Bigger and Cole, 1983); a central hardwood forest in Cockaponset, Connecticut (with Acer rubrum, Betula lenta, Carya spp. and Quercus spp.); a northern spruce–ﬁr stand in Chesuncook, Maine (with Abies balsamea and Picea rubens); and a northern hardwood forest at Mt. Success, New Hampshire (with Acer saccharinum, Betula alleghaniensis and Fagus grandifolia) [results from the latter three sites are described in detail by Hornbeck et al. (1990)]. Most of these sites were nitrogendeﬁcient; and the comparisons with sites in the United Kingdom and southern Sweden, which are richer in nitrogen and have high deposition rates, are of interest. In most cases, harvesting produced increases in nitrate concentration in either soil solution or stream water (with the notable exception of the red alder stands, as described above). In most cases, there were only minor (<20%) differences in nitrate leaching between the SAW and WTH treatments, and both were very low (ranging across all sites from 0.2 to 6.3 kg N ha−1 yr−1 over the 2–4 years sampling period). In the slashpine forest at Bradford, Florida, the leaching of nitrate was greater with WTH than with SAW, but both

Dale W. JOHNSON and Bradley SCHULTZ

were very low (< 0.1 kg N ha−1 yr−1 ). Similarly, in the Douglas-ﬁr forests of low site quality at Pack Forest, Washington, the leaching of nitrate was greater with WTH than with SAW (1.8 and < 0.1 kg N ha−1 yr−1 , respectively). However, the reverse was true in the Douglas-ﬁr forests at this site of higher site quality (NO−3 leaching in the WTH and SAW treatments were 0.1 and 2.2 kg N ha−1 yr−1 , respectively). Reasons for these differences were not speciﬁed, but could have included differences in nitrogen uptake by regeneration and nitrogen release from slash, as noted in the European studies. The effects of other disturbances on forests There have been far fewer studies on the effects of disturbances other than harvest on nutrient losses from forest ecosystems. Among the disturbances that can be included in this miscellaneous category are insect outbreak, addition of substances with a high C/N ratio, and temperature ﬂuctuations. Swank et al. (1981) studied the effects of an outbreak of the fall cankerworm (Alsophila pometaria) on stream-water nitrate in mixed hardwood forests at Coweeta. In contrast to earlier studies on the effects of insect outbreaks at Hubbard Brook, New Hampshire (Bormann and Likens, 1979), Swank et al. found clear signals in stream-water nitrate which coincided very closely with levels of infestation. Stream-water nitrate was elevated in infested catchments, and the authors were able to track the timing of infestation and increases in stream-water nitrate. They attributed these results to increased rates of nitrogen cycling by premature litter-fall and frass return to the forest ﬂoor. Studies on the effects of the gypsy moth (Porthetria dispar) now in progress also show increased rates of nitrogen cycling and losses as a result of insect infestation (G.M. Lovett, pers. comm.). Edwards and Ross-Todd (1979) reported the results of a study where plots of tulip poplar (Liriodendron tulipifera) were girdled to examine the effects of belowground carbon allocation on nutrient cycling and retention processes. (By removing the phloem, girdling cut off transport of carbohydrates to roots, whereas nutrient and water uptake was allowed to proceed in the xylem.) In one treatment, sprouts were allowed to regrow after girdling and in another treatment sprouts were removed. Girdling did not result in immediate root death or the cessation of aboveground cycling: litter-fall in the girdled plots two years after treatment

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

remained at 70–80% of levels in the control-plots, and diameter growth was 50–70% of control-plot rates. By far the most signiﬁcant effect of girdling was in leaching of soil nitrate, which increased from 0.15 kg N ha−1 yr−1 in control plots to 25 kg N ha−1 yr−1 in the girdled plot where sprouts were removed and to 9 kg N ha−1 yr−1 where the sprouts were allowed to grow. Differences in uptake could account for only a portion of the increased losses, so Johnson and Edwards (1979) conducted detailed studies on the dynamics of soil nitrogen in the same treatments. They found that girdling had stimulated both nitrogen mineralization and nitriﬁcation rates. There was no evidence of nitriﬁcation inhibitors from either litter or root extracts; nitriﬁcation seemed to be controlled by ammonium supply. Additions of sucrose to subplots within the girdled plot caused large reductions in nitrogen mineralization and nitrate leaching, whereas nitrogen fertilization in control plots caused increases in both. Johnson and Edwards attributed the positive effects of girdling on nitrogen mineralization and nitrate leaching to the death of ﬁne roots and mycorrhizae; in retrospect, it seems that these responses may have also been due to carbohydrate exudation following girdling, and the resulting lower heterotrophic nitrogen demand.

Summary of forest disturbances In summarizing the results of studies since 1980 on the effects of harvesting and other disturbances on nitrate leaching, it is useful to compare the lists of major factors controlling nitrogen losses given by the various authors with those known of and discussed previously. Table 23.2 lists the major factors to which nitrogen retention and losses have been attributed in the various studies discussed above. The list of factors include those of Vitousek et al. (1979), slightly modiﬁed, plus some additional factors which were not on their list but which have proven to be important: canopy scavenging, nitrogen ﬁxation, and increased cycling. It is readily apparent that soil microbial mineralization and immobilization is the most commonly invoked mechanism: it appears on every list. Nitriﬁcation in and of itself is invoked as an important factor in only about half of the studies reviewed; however, it is clear that nitrogen mineralization/immobilization would have little effect upon nitrate leaching unless also followed by nitriﬁcation. Plant uptake is invoked as an important factor in about half of the studies reviewed and is

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often explicitly identiﬁed as being more important than any other factor (e.g., in the European wholetree versus sawlog harvesting studies). In other cases, plant uptake is identiﬁed as being a rather unimportant factor (e.g., in the red alder harvesting studies, the study of harvest following fertilization in Sweden, and in the girdling study). Denitriﬁcation is either ignored, or, when measured, found to be unimportant. Increased water ﬂux is often a factor affecting nitrate losses after harvesting, but it was not identiﬁed as being a major factor in these studies. In one case (the harvesting and site-preparation studies in North Carolina), water ﬂux was invoked as a cause for delay in nitrate leaching. Nitrate adsorption was not invoked as a factor in any of the studies reviewed. Canopy scavenging was identiﬁed as a major factor affecting ecosystem nitrogen budgets, and nitrate leaching was recognized as important in several of the European studies but not in those in North America. Site fertility clearly had an effect on nitrate leaching in the fertilizer harvest studies in Sweden, but not in the red-alder harvesting studies. When it was measured rather than speculated upon, nitrogen leaching from slash was never identiﬁed as an important factor affecting nitrate leaching. Nitrogen ﬁxation was not identiﬁed as an important factor in any of the studies reviewed, except as the reason for high rates of nitrate leaching prior to harvesting in the red alder studies. We have entered reduced nitrogen ﬁxation as a possible reason for the reduced nitrate leaching following harvesting in red alder (with a question mark), for the reasons discussed above. Increased cycling rate was invoked as an important factor in two of the studies that did not involve harvesting (insect defoliation and girdling). Clearly, increased cycling rate would not be an important factor in the initial stages of response after harvesting. In summary, then, the studies reviewed above concluded that microbial mineralization/immobilization was nearly always an important factor in nitrate leaching and that plant uptake, nitriﬁcation, and canopy scavenging were the most frequently important after this. Vitousek et al. (1979) identiﬁed plant uptake, nitrogen immobilization, lags in nitriﬁcation, and lack of water as the four most frequently important factors affecting nitrate leaching in their series of trenched-plot experiments. In a later analysis of the same data set combined with laboratory incubations, Vitousek et al. (1982) identiﬁed nitrogen mineralization/immobilization and lags in nitriﬁcation as the two most important factors. In a later paper reviewing

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Dale W. JOHNSON and Bradley SCHULTZ

Table 23.2 Factors thought to be most important in controlling nitrate losses in various forest disturbance studies Reference

Type of manipulation 1

Forest type

Major factors 2

Adamson et al. (1987)

SAW

Picea sitchensis

1,4,5,9

Bigger and Cole (1983)

SAW, WTH, litter removal

Alnus rubra, Pseudotsuga menziesii

(1),(4),(10),12?

Edwards and Ross-Todd (1979); Johnson and Edwards (1979)

girdling

Liriodendron tulipifera

1,(4),5,13

Hendrickson et al. (1989)

SAW, WTH

mixed conifer, hardwood

1,4

Hornbeck et al. (1986)

SAW, strip cutting

northern hardwood

1,4

Jewett et al. (1995)

SAW

mixed conifer, hardwood

1,4,5

Mann et al. (1988)

SAW, WTH

various species

1,4

Nohrstedt et al. (1994); Ring (1995)

fertilization, SAW

Pinus sylvestris

1,(4),5,(6),(9),(10)

Reynolds et al. (1995)

SAW

Picea sitchensis

1,4,5,9

Rosen and Lundmark-Thelin (1987)

SAW, slash piling

Pinus sylvestris, Picea abies

1,4,(11)

Staaf and Olsson (1994)

SAW, WTH

Picea abies

1,4

Stevens and Hornung (1988)

SAW

Picea sitchensis

1,4,5

Stevens and Hornung (1990)

SAW, WTH

Picea sitchensis

1,4

Swank et al. (1981)

insect defoliation

mixed deciduous

1,12

Van Miegroet et al. (1992)

SAW, species replacement

Pseudotsuga menziesii, Alnus rubra

(1),(7),12?

Vitousek and Matson (1985)

SAW, WTH, site preparation, herbicide

Pinus taeda

1,5,7

Wiklander et al. (1991)

blow-down, SAW

Picea abies

1,4,5,9

1

Abbreviations: SAW, conventional sawlog harvesting, slash left on site; WTH, whole-tree harvesting, slash removed. 1, Microbial mineralization/immobilization; 2, ammonium ﬁxation in clays; 3, ammonia volatilization; 4, plant uptake; 5, nitriﬁcation; 6, denitriﬁcation; 7, water ﬂux; 8, nitrate adsorption; 9, canopy scavenging; 10, site fertility; 11, leaching from slash; 12, nitrogen ﬁxation; 13, increased cycling rate; (parentheses indicate that the factor in question was measured and considered not to be important).

2

nitrogen budgets of a variety of undisturbed forest ecosystems, Johnson (1992) identiﬁed plant uptake as the most important factor affecting nitrate leaching, and speculated that, over the long term, forest vegetation could actually out-compete heterotrophs for soil nitrogen, resulting in “mining” of soil nitrogen by plants. This speculation arose from the fact that net nitrogen increment in vegetation often exceeded atmospheric nitrogen inputs by a considerable margin. Free-living nitrogen ﬁxation was not seriously considered as a way to account for this imbalance, as subsequent studies clearly showed. Bormann et al. (1993) found rates of nitrogen accumulation well above atmospheric deposition rates in nitrogen-poor soils (sand) under Pinus spp. over a ten-year period, suggesting that free-living nitrogen ﬁxation was an important nitrogen input. Free-living nitrogen ﬁxation certainly deserves more consideration in all studies of nitrogen balance. Johnson (1992) also pointed out the possibility that

abiotic reactions between nitrogen and soil organic matter could account for a signiﬁcant proportion of nitrogen retention in soils. This factor was not listed in Table 23.2 because none of the investigators considered it; however, it remains an unexplored possibility.

EFFECTS OF DISTURBANCE ON RANGELANDS

This section will focus on spatial and temporal variation in nutrient cycling on native landscapes mainly grazed by generalist ungulates. Carbon and nitrogen are the primary elements discussed. Other nutrients have different initial sources (geologic versus biotic), but all life-sustaining nutrients eventually are supplied and recycled through the soil organic matter, similarly to nitrogen (Killham, 1994). Much of the discussion about how grazing affects and disturbs inputs of soil organic matter and nitrogen cycling is

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

transferable to other essential nutrients. Reference to anthropogenically maintained pasture (e.g., plowing, seeding, fertilization) is avoided because the inﬂuences of pasture maintenance probably override the effects of grazing, and the two are difﬁcult to separate. To illustrate how grazers interact with and may disturb nutrient cycles, our discussion will initially emphasize three topics: (1) inputs of organic carbon to the soil from roots, above-ground live and dead plant material, and animal excreta; (2) how grazers may alter inputs of organic carbon to the soil; and (3) several ecosystem level processes that integrate grazing, inputs of organic carbon to the soil, and carbon–microbe interactions. Linkages with the hydrologic cycle and cryptobiotic crusts are addressed because grazing-induced disturbance to them can affect important indirect links with nutrient cycling (West, 1990; Thurrow, 1991; Johansen, 1993). Carbon inputs and grazing inﬂuences Inputs of organic carbon to the soil are from six sources: roots (exudates, and root turnover); standing live, standing dead, and surface (i.e., litter) plant material; animal excreta; microbial biomass; cryptobiotic crusts; and decomposing fauna. Grazers may affect inputs of organic carbon from the ﬁrst ﬁve sources. Most organic carbon inputs, however, are from roots; standing live, standing dead, and surface plant litter; and animal excreta. Our discussion will therefore focus on them. Roots constitute the largest input of organic carbon, and intense frequent grazing may alter total root biomass and/or its distribution in the soil proﬁle. Surface litter and standing plant material, live or dead, also contribute large amounts of organic carbon to the soil (Coleman et al., 1983). Heavily grazed locations, however, may have substantial inputs from excreta and only small inputs from plant litter. Roots Grazing disturbs (alters) nutrient cycles by changing the total and/or the distribution of root biomass (organic carbon inputs) in the soil. Total root biomass is important because sufﬁcient amounts of labile organic carbon must be present to support microbial decomposition. Root distribution is important because the organic carbon must be near the soil surface, where microbial populations are largest. An understanding of how excessive grazing disturbs nutrient cycles requires knowledge about how grazing-induced plant succession affects root biomass and its distribution.

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Individual plants concentrate their root biomass at species-speciﬁc depths (Abbott et al., 1991; Lee and Lauenroth, 1994); however, two patterns are evident. First, most biomass is shallow (Bartos and Sims, 1974; Svejcar and Christiansen, 1987; Ganskopp, 1988; Manning et al., 1989; Rodriguez et al., 1996); and declines with depth (Svejcar and Christiansen, 1987; Manning et al., 1989; Rodriguez et al., 1996). Second, increased aridity generally results in most roots being concentrated at shallow depths, except when shrubs are the dominant life-form (Rodriguez et al., 1995), particularly if the climatic pattern promotes deep percolation of soil moisture. Root biomass for individual plants declines with increased defoliation (Branson, 1956; Richards, 1984). Total root biomass at the community level often decreases (Padney and Singh, 1992; Zhang and Romo, 1994; Hofstede and Rossenaar, 1995; Rodriguez et al., 1995) but not always (Bartos and Sims, 1974). Also, intense grazing usually redistributes root biomass, leading to greater biomass at shallow depths (Smoliak et al., 1972; Bartos and Sims, 1974; Rodriguez et al., 1995; Chaieb et al., 1996). Changes in root biomass, and/or its distribution, probably are greatest where shallow- and deep-rooted species, or woody and herbaceous species, replace each other. Standing plant material, live and dead, and surface litter Above-ground plant matter has three components: standing live, standing dead, and surface litter. Grazing inﬂuences the input of organic matter from each component to the soil, the ratio among components, and each component’s contribution to soil organic matter. Rates of nutrient cycling generally increase when transfer rates between components increase, or the transfer process bypasses one or more components (Botkin et al., 1981; Risser and Parton, 1982; Ruess et al., 1983; Seagle et al., 1992). The transfer of organic matter among components results from the ingestion, digestion, and excretion of plant material, and the trampling and breakage of vegetative material. During digestion, organic matter is ground into smaller particles, with many nutrients extracted, used in metabolic processes, and excreted at concentrations greater than those in the standing dead and surface litter (Botkin et al., 1981; Ruess and McNaughton, 1987). Ingested organic matter bypasses the typical microbial decomposition process, with mineralized nutrients returned to the soil, both directly and quickly (Botkin et al., 1981).

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Ungrazed grasslands have substantially more biomass from standing dead material and surface litter than do grazed grasslands (McNaughton, 1979; Naeth et al., 1991), which lengthens the time required to incorporate the organic matter into the soil (Ruess and Seagle, 1994). High levels of trampling quickly transfer large amounts of standing dead material to the pool of surface litter. Additional trampling further reduces the particle size of the litter. Rapid conversion from standing dead material to small, particulate soil organic matter increases decomposition and cycling rates for at least three reasons. First, soil moisture levels beneath the soil surface generally are higher than at the soil surface, thus, soil moisture remains higher for longer periods. High soil moisture increases microbial decomposition. Second, soil microbes have direct access to soil organic matter, facilitating microbial breakdown. Third, particles of organic matter incorporated into the soil typically are smaller than surface litter, and small particles decompose more quickly than large particles (given the same environmental conditions). Animal excreta Ingested organic carbon is returned directly to the soil in animal excreta. This causes organic carbon that normally would pass through three organic-matter components (described above) to pass through two or fewer components and return to the soil within days, not weeks, months or years (Botkin et al., 1981; Ruess and McNaughton, 1984). Also, organic carbon (and nutrients) in the fecal matter often is leached into the soil more readily than nutrients in undigested plants (Ruess and McNaughton, 1987). This probably occurs because digestion reduces the size of the plant particles, facilitating the transport of the remaining labile nutrients to the soil. Small plant particles also are more rapidly incorporated into the soil, facilitating microbial breakdown. For ungrazed locations, a relatively large stable pool of non-labile organic carbon (and nutrients) will remain as standing dead material and surface litter, with small ﬂuxes into and through the soil. Ecosystem-level processes Understanding how grazers increase or decrease nutrient ﬂow through grazed ecosystems requires understanding of how grazers alter the spatial and temporal drivers for organic matter decomposition. A comparison of spatial and temporal variation in two

Dale W. JOHNSON and Bradley SCHULTZ

types of grazed ecosystems best illustrates landscapelevel variation for the drivers of organic-matter decomposition among ecosystems. The ecosystems discussed are true grasslands in the Serengeti–Mara Savanna in equatorial Africa, the North American Great Plains, and the shrub-steppe rangelands in the Intermountain West of the United States. The African and American grasslands evolved with widespread, intensive grazing, regional migration patterns across large relatively level landscapes, and growing-season precipitation (Gerresheim, 1974; Maddock, 1979; McNaughton, 1985). The sagebrush steppe, however, has been without bulk generalist grazers for more than 10 000 years, has limited vertical migration between valley ﬂoors and adjacent high mountains, and receives most of its precipitation during the winter when plants are dormant (Mack and Thompson, 1982; West, 1983). Other rangeland systems exist, but we are limiting our discussion to the Serengeti and Great Plains grasslands, and the shrub-steppe region for two reasons. First, these systems have important and clear contrasts in grazing history, climate, vegetation structure, and topography; and these contrasts illustrate how several ecosystem components drive the nutrient-cycling process and how disturbance to these components can affect nutrient cycling. Second, these rangeland systems have large tracts of native rangeland ecosystems that are little impacted by external human activities (e.g., increased nitrogen deposition from heavily polluted urban areas). This helps to ensure that the research results discussed are the result of grazing, and not of other external inﬂuences. Grazing, plant succession, and root biomass Grasslands that evolved with abundant generalist grazers and that are grazed at moderate intensities exhibit little or no decline in soil organic carbon when grazed (Milchunas and Lauenroth, 1993; Dormaar et al., 1994; Frank et al., 1995; Manley et al., 1995). Roots are the largest input of soil organic-carbon (Coleman, 1976); therefore, the maintenance of levels of soil organic carbon with grazing suggests that organic-carbon inputs from roots do not decline. The critical factor appears to be the process of grazinginduced succession, which results in dominance by one or more grazing-tolerant species, which invest large amounts of organic carbon below ground at shallow depths. For example, in the North American Great Plains, tall-grass species show decreased abundance and individual root biomass when grazing intensity

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

increases (Svejcar and Christiansen, 1987; Milchunas et al., 1988). With increased grazing intensity, the abundance of blue grama (Bouteloua gracilis) and other short grasses typically increases (Smoliak et al., 1972; Lauenroth et al., 1994). Blue grama also invests relatively more photosynthate per plant below ground than the grazing-intolerant tall species and concentrates its roots within 30 cm of the soil surface (Frank et al., 1995). The shallow rooting depth is important because infrequent and small precipitation inputs during the warm growing season (Lauenroth et al., 1994) limit inﬁltration. Thus, inputs of organic carbon and moisture occur together spatially and temporally, facilitating microbial activity (Killham, 1994). A shallow, dense root system (i.e., greater biomass) with a high-quality substrate (i.e., low lignin content and high nitrogen content) probably enhances microbial decomposition and nutrient mineralization in grazed locations. For ecosystems that evolved without extensive grazing by large generalist ungulates (e.g., the shrub-steppe region of North America’s Intermountain West), heavy grazing during the growing season results in the decline of root biomass for individual grazed plants (Branson, 1985; Ganskopp, 1988). The long-term, communitylevel response differs between true grasslands and shrub-grass rangelands, largely because differential herbivory among shrubs and grasses results in less palatable, deep-rooted shrubs replacing shallow-rooted grasses (Ellison, 1960; Young et al., 1976; Sturges, 1977; Branson, 1985; Miller et al., 1994). Replacement of grass species by shrubs promotes the immobilization of nutrients and slows the cycling process. Five mechanisms may be responsible. First, root biomass is distributed across greater depths (Sturges, 1977), potentially reducing total concentrations of organic matter (and hence nutrients) at shallow soil depths (Rodriguez et al., 1995). Also, microbial biomass typically declines with soil depth (Burke et al., 1989; Bolton et al., 1993); therefore, the microbial population and its energy source may become partially separated in space. Essentially, the balance between inputs of organic carbon and microbial biomass changes. Second, woody roots generally have a higher lignin content than herbaceous roots, which lowers substrate quality. Third, live and dead woody stems remain standing (above ground) longer than herbaceous culms, reducing their incorporation into the soil organic matter. This slows the rate of transfer of organic matter and nutrients through the cycling process. Fourth, many shrubs are evergreen, and thus retain their leaves longer

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than herbaceous species. Fifth, generalist grassland herbivores prefer grasses to shrubs (Hanley, 1982; Hanley and Hanley, 1982); thus, shrub leaves and stems remain as standing organic matter for long periods (discussed above, p. 556). Replacement of shrubs by grasses also alters the spatial distribution of above- and below-ground carbon inputs. Litter accumulates under shrub canopies (Rickard et al., 1973; Charley and West, 1977; Barth and Klemmedson, 1978; Doescher et al., 1984), reducing levels of organic matter in the interspaces. Beneath bare sites organic matter decomposition and nutrient cycling decline because inputs of organic carbon and nutrients are less (Hook et al., 1994). Under the shrubs, turnover rates often are low because of high lignin content in the roots and litter. The total amount of organic matter decomposed and nutrients mineralized eventually may be substantial, and may approach levels in grasslands (Bolton et al., 1993), but only after large amounts of litter accumulate. Poor substrate quality (i.e., small amounts of labile carbon and nutrients) increases the amount of organic matter required to mineralize similar amounts of labile organic carbon and nutrients. Carbon–microbe interactions Serengeti and North American Great Plains grasslands: Microbial populations (and hence biomass) ﬂourish when inputs of organic carbon and nitrogen are high, and the soil is moist and warm (Killham, 1994). Grazing may modify soil moisture and temperature (Tomanek, 1969; Whitman, 1971), potentially upsetting balances of nutrient mineralization and cycling that optimize microbial biomass (e.g., the C:N ratio). The C:N ratio in the soil largely regulates the availability of nitrogen and other nutrients. Numerous studies in the North American Great Plains (Detling et al., 1979; Polley and Detling, 1989) have found that defoliation generally increases the proportion of carbon allocated to shoots and leaves while decreasing the proportion allocated to roots, reducing C:N ratios in the soil. This decreases root and microbial biomass (Ruess and McNaughton, 1987), but not necessarily net mineralization (Ruess and McNaughton, 1987) and availability of nutrients for plant uptake (Ruess and McNaughton, 1984; Holland and Detling, 1990; Ruess and Seagle, 1994). A high C:N ratio results in high rates of nitrogen immobilization by soil fauna. Large inputs of carbon allow small microbial populations to expand rapidly and sequester most recently mineralized

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nutrients (Killham, 1994). Low C:N ratios result in less nitrogen being incorporated into the microbial biomass, low immobilization potential (Holland and Detling, 1990), and increased levels of soil nutrients (Holland and Detling, 1990; Seagle et al., 1992). Landscape patterns strongly inﬂuence spatial variability for organic carbon inputs and C:N ratios (Ruess and Seagle, 1994). High growing-season precipitation in the tall grasslands of the Serengeti permits high inputs of organic carbon, which sustain a large microbial biomass. The microbes mineralize large amounts of organic nitrogen. Nitrogen availability for plant uptake, however, is low because recently mineralized nitrogen is rapidly incorporated into the microbial biomass. The absence of grazing (disturbance) during the growing (rainy) season allows annual precipitation, rather than grazing levels, to inﬂuence maximum above- and below-ground primary production. Total nitrogen in the system is ﬁnite and primary production high; therefore, nitrogen is diluted across large biomass accumulations, resulting in low tissue concentrations. The limited return of organic carbon and organic nitrogen to the soil when microbial biomass and plant growth are high (i.e., in the rainy season) limits total nutrient mineralization and permits the microbes to incorporate most nutrients into their biomass. Repeated grazing of the tall grasses during the growing season would be an ecosystem disturbance, because it departs from the evolved pattern. Such grazing probably would result in succession by shorter, grazing-tolerant grasses and the rapid return of organic matter to the soil in the form of animal excreta, and litter from trampling damage. Soil and plant levels of nitrogen may increase because the amount of organic matter returned to the soil has quickly increased and during plant growth the nutrients are distributed across less biomass. Soil and plant nitrogen, however, may decrease if the average annual precipitation is high enough to increase leaching losses. Short-grass grasslands in the Serengeti (and probably other short-grass rangeland systems) have less rainfall and smaller total inputs of organic carbon than tallgrass grasslands. Short-grass grasslands, unlike the tall-grass grasslands, have high fertility despite high grazing intensities (Ruess and McNaughton, 1987; Ruess and Seagle, 1994). Intense grazing results in plant succession with grazing-tolerant species (McNaughton, 1979) supplying ample amounts of organic carbon to support microbial activity but insufﬁcient to permit large increases in microbial populations

Dale W. JOHNSON and Bradley SCHULTZ

that would sequester most nutrients (inferred from Holland and Detling, 1990). Plants may either increase their above-ground net primary production (ANPP) or increase nitrogen concentrations in their leaves and shoots. Increased production seems likely if nutrient availability, rather than water or defoliation levels, limits growth. Higher nitrogen concentrations are likely if water availability or defoliation levels limit productivity, because the nitrogen then becomes concentrated across a smaller amount of plant biomass. This contrast between grazed short-grass grasslands, and tall-grass grasslands (Table 23.3) suggests that short-grass grasslands are disturbed less by grazing and have greater fertility with grazing, because of numerous landscape-level differences. Lower precipitation accompanied by high grazing intensities reduces total carbon inputs and promotes succession by species that are tolerant to both drought and grazing. Lower inputs of carbon result in microbial populations becoming limited by carbon or water, not nutrients. Carbonlimited microbial populations may mineralize less total nutrients, but have high net mineralization rates because of lower immobilization potential. Lower productivity also reduces total nitrogen assimilation in stems and leaves, but allows for greater nitrogen concentrations. Limited inﬁltration in arid areas also reduces or eliminates nutrient losses from leaching. Finally, ﬁner textured soil at some locations may reduce leaching through cation retention. Sagebrush steppe: In the sagebrush steppe, direct simultaneous measurements of interactions among organic carbon inputs, microbial dynamics, nutrient mineralization, and grazing have not been made. Most research has addressed nutrient cycling in ungrazed locations, or the grazing component of the process has not been addressed and discussed. Inferences for grazing-mediated interactions are possible, to some extent, because landscape-level variations in the drivers of the decomposition and mineralization processes (i.e., carbon inputs, soil moisture and temperature levels) are relatively well known (Burke, 1989), as are factors that inﬂuence animal distribution (Heady and Childs, 1994). This ecosystem type generally has high winter snowfall (West, 1983), strong winds (Sturges, 1986), and substantial topographic variability (West, 1983), which results in substantial redistribution of precipitation (Burke, 1989). This creates spatial and temporal differences in soil moisture across distances much

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Table 23.3 Comparison of important components of the nutrient cycle and their response with and without grazing in true grasslands and shrub-steppe rangelands Nutrient cycling component

Response with grazing

Response without grazing

True grassland with growing-season precipitation Plant succession

short-statured species with high tolerance for defoliation favored

taller-statured species with less tolerance for grazing favored

Organic-matter inputs to soil

rapid; quick conversion of standing live and dead material to excreta bypasses normal decomposition process

slow; vegetation remains as standing live and dead material for longer periods; slow transfer to soil where microbial decomposition can occur

Roots

decreased rooting depth; decreased biomass per plant; high biomass at shallow depth

increased rooting depth; increased biomass per plant; biomass more evenly distributed throughout soil proﬁle

Standing live material

decreases

increases

Standing dead material

decreases

increases

Surface litter

decreases

increases

Animal excreta

increases

decreases

Microbial biomass

less than if ungrazed; concentrated near soil surface under plants; active throughout summer after precipitation events

more than if grazed; concentrated near soil surface; active throughout summer after precipitation events

Mineralized nitrogen

increases

declines

Shrub-steppe rangeland with precipitation concentrated during period of winter dormancy Plant succession

shrubs increase and grasses decline; grass abundance declines in interspaces between shrubs

shrubs are few and grasses abundant; grasses abundant in interspaces between shrubs

Organic matter inputs to soil

rapid if site remains primarily grassland; generally slow if perennial shrubs dominate site

slower than grazed sites because animal excreta are absent; rate can depend on percent composition of shrubs and grasses; quicker when grasses are most abundant

Roots

total biomass may remain high, but increase in shrubs alters distribution, with more biomass at greater depths; for grasses individual and total biomass declines and biomass at shallow depths is low

shrub roots are few; most root biomass is from grasses and occurs at shallow depths

Standing live material

can increase as less-palatable shrubs increase; shrubs have large amounts of less labile organic carbon

high, but consists primarily of grasses with large amounts of labile organic carbon

Standing dead material

may or may not be higher, but becomes concentrated in woody stems

high, but is concentrated in herbaceous culms

Surface litter

may be high or low, depending on age of shrubs, but is concentrated in woody material

high, but is primarily herbaceous culms

Animal excreta

high

low

Microbial biomass

high under shrubs; low in interspaces; active only in spring when soil is moist and soil temperature is warm

high under grasses and shrubs; lower in interspaces, but still higher than barren interspaces typical of grazed sites dominated by shrubs; active only in spring when soil is moist and soil temperature is warm

Mineralized nitrogen

can be high under shrubs, but only after large amount of poor-quality litter accumulates; very low in interspaces

probably higher than grazed locations; higher under than between vegetated clumps, but difference is much less than in grazed sites

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shorter than in the relatively level Serengeti and North American Great Plains. Locations that accumulate winter snow retain high soil moisture levels longer than windblown sites, but no location has moist soil throughout the potential growing season (Burke, 1989). Plant communities on landscapes that receive less effective precipitation and more solar radiation, and have shallow or coarse soil, have lower inputs of organic carbon, less soil moisture, and warmer soil temperatures. This reduces the carbon and nitrogen in total and microbial biomass, the length of the period of active mineralization, and total mineralization (Burke, 1989; Burke et al., 1989). Within-community spatial and temporal heterogeneity also are great. At the shrub–intershrub spatial scale, the pattern of organic-matter accumulation and nutrient pools follows the pattern observed for landscape variability. Mesic locations under plants have greater accumulations than xeric locations between plants (Burke et al., 1989; Bolton et al., 1993). Grazing levels that reduce organic carbon inputs and/or soil moisture, when soil temperatures are warm, may be expected to reduce nutrient pools and mineralization levels, which in turn would reduce soil fertility. Most likely, this results when repeated differential defoliation (between shrubs and grasses) during the growing season promotes an increase in shrubs and a decrease in perennial grasses. Shrubs have roots, stems, and leaves giving substrate of lower quality and less root biomass at shallow depths; therefore, inputs of labile organic carbon should decline. Less organic carbon combined with a short period of active mineralization (Burke et al., 1989) may be expected to reduce total nutrient mineralization. Excessive grazing also reduces soil moisture because inﬁltration declines (discussed below, pp. 561–562). Less moisture would further reduce the already short period of active mineralization. Less input of organic carbon, or a change to lower-quality substrate, when combined with low soil moisture, probably have the greatest negative inﬂuence on mineralization rates. Table 23.3 contrasts important components of the nutrient cycle for grazed and ungrazed locations in the shrub-steppe. Facilitation of energy and nutrient ﬂow The redistribution of nutrients in animal excreta generally has a positive feedback for productivity (Day and Detling, 1990; Haynes and Williams, 1993; and references therein). Dung patches may enhance plant growth in the area covered by a factor of as much as six.

Dale W. JOHNSON and Bradley SCHULTZ

The facilitation of energy and nutrient ﬂow among grazers is the result of at least three mechanisms. First, live growth that would become standing dead material and/or surface litter is consumed; and the nutrients are returned to the soil quickly, in animal excreta. This facilitates rapid reuse by plants and their consumers. Second, large-bulk grazers can remove large amounts of coarse low-quality forage, increasing the accessibility of shorter plants with high nutrient quality to specialist grazers (McNaughton, 1976). Third, grazed plants often have higher nutrient concentrations than ungrazed plants (Day and Detling, 1990; McNaughton, 1992; Gauthier et al., 1995). This phenomenon results from interactions among carbon, nitrogen, soil microbes, and herbivores (Holland and Detling, 1990). Grazing decreases inputs of organic carbon from the roots and from above-ground net primary productivity. Decreased inputs of organic carbon reduce microbial biomass (due to carbon limitation) and immobilization potential (Holland and Detling, 1990). Carbon-limited microbial populations have net positive mineralization rates, increasing soil concentrations of nitrogen. Increased amounts of mineralized nitrogen, when accompanied by growing-season precipitation, probably result in increased nitrogen uptake. Increased nitrogen uptake, but decreased above-ground net primary productivity per plant (in grazed sites), leads to higher nitrogen concentrations in stems and leaves. Hydrologic cycle, grazing, and nutrient cycling Nutrient cycling may be indirectly affected by herbivores through alteration of the hydrologic cycle (Archer and Smeins, 1991). Research about the effects of grazing on rangeland hydrology has primarily investigated how grazing affects: (1) interception of precipitation; (2) surface detention; (3) aggregate stability; and (4) inﬁltration (Thurrow, 1991). Little if any research has directly addressed links between these four areas and nutrient cycling. The need for synthetic research is obvious. Accelerated erosion removes nutrients from one landscape and deposits them in another, and is discussed elsewhere (see Pimentel, Chapter 4, this volume). Our discussion will incorporate knowledge about how moisture inﬂuences decomposition of organic matter and nutrient cycling, and we will suggest probable interactions between hydrology, grazing, and nutrient cycling. Efﬁcient decomposition and nutrient cycling requires inputs of organic carbon and moist soil (Burke, 1989;

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

Killham, 1994). From a hierarchical perspective, annual precipitation initially controls the input of organic carbon. Secondary control is due to the season, intensity, and frequency of grazing (see Ruess and Seagle, 1994). Increased grazing generally decreases aboveand below-ground primary production, and generally reduces microbial biomass. Grazing levels that spatially or temporally reduce soil moisture, and/or alter soil physical conditions so that organic-carbon inputs, microbial biomass, or microbial activity are reduced, probably reduce the amount and rate of organic matter decomposition and nutrient mineralization. The ﬂow of water across and subsequent retention in grazed landscapes is inﬂuenced by at least 11 factors (reviewed by Thurrow, 1991). These include: (1) foliar plant cover; (2) percent surface litter cover; (3) the size distribution of plant litter; (4) the distribution and orientation of plant litter across the landscape; (5) the number, size and orientation of surface depressions; (6) the orientation of livestock trails (with respect to slope); (7) soil aggregate structure; (8) surface texture; (9) soil depth; (10) slope; and (11) rainfall intensity (not impact). Grazers directly or indirectly inﬂuence each of these, except slope and rainfall intensity. A change in one or more of these features may increase or decrease water retention (and soil moisture levels). The speciﬁc results will depend on site-speciﬁc stocking levels (Willatt and Pullar, 1983) Reductions in foliar or litter cover below site-speciﬁc thresholds may increase the force of raindrop impact, increasing the dispersion of silt and clay particles (Blackburn, 1975; Thurrow et al., 1986). Nutrient-rich clay particles or small colloids can be moved off-site by subsequent overland ﬂow or eolian transport more quickly than large aggregates (Graetz and Tongway, 1986). Substantial nutrient loss eventually reduces site fertility with negative feedbacks for inputs of organic matter, from roots and litter-fall. The size and orientation of plant litter and surface depressions inﬂuence inﬁltration by altering patterns of overland ﬂow and surface detention. Grazing management that promotes inﬁltration in arid regions may increase the maximum water content of the soil (compared with improperly grazed locations), and/or result in higher levels of soil moisture for a longer period in the growing season. At a minimum this should maintain inputs of organic carbon from roots, and may increase total primary production. Less inﬁltration, depending on how severe and prolonged, may reduce the period of microbial activity (see Burke,

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1989; Burke et al., 1989), organic-carbon inputs, or both. Lower inputs of organic carbon are possible from reduced root growth of existing species, or from a change to more xeric species, which may alter the distribution of root biomass. A change from grass to shrubs would probably alter the spatial distribution of roots (Lee and Lauenroth, 1994), and/or reduce substrate quality. Well-developed livestock trails (i.e., barren or nearly so) probably slow nutrient cycling, and may reduce nutrient levels. Trails running up and down a slope increase the transport rate of water downslope, increasing erosion potential and decreasing inﬁltration. Increased erosion moves nutrients off-site, and as rills and gullies enlarge, nutrient loss increases substantially in three dimensions (depth, width, and length). Decreased inﬁltration probably shortens the period of active mineralization, which can reduce nutrient availability. Livestock trails perpendicular to a slope may be less prone to erosion. These trails, however, often have reduced vegetative growth, which reduces inputs of organic carbon, in turn probably slowing nutrient cycling. Also, if depressions develop, the downslope movement of water may be interrupted, reducing inﬁltration immediately downslope from the trail. Less inﬁltration could shorten the period of mineralization. The extent to which downslope inﬁltration is impeded by livestock trails undoubtedly depends on regional and/or site-speciﬁc precipitation patterns. Locations with relatively heavy winter precipitation, particularly in the form of slow-melting snow, probably have much better inﬁltration and retention downslope (from livestock trails) than do locations subject to infrequent convection storms during the warm summer months. At the landscape level, locations with low trail density probably are affected less than locations with high trail density. Soil aggregate structure and stability determine pore size and porosity (Kemper and Rosenau, 1986), which inﬂuence inﬁltration rates, particularly in wet soil (Thurrow, 1991). Excessive trampling reduces aggregate structure, aggregate stability, and inﬁltration (Warren et al., 1986) by increasing clay dispersion, which plugs soil pores. Trampling also increases compaction (Willatt and Pullar, 1983), which decreases pore size, soil depth, and water holding capacity. Grazing wet soil results in the highest compaction, which may reduce the inputs of organic carbon inputs from root growth (Lowery and Schuler, 1991), effectively reducing microbial decomposition. Trampling of dry

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soil may result in segregation of particle sizes (Graetz and Tongway, 1986), with ﬁner (smaller) particles remaining on the soil surface, and thus facilitating off-site transport (Thurrow, 1991). Particle segregation may also result in a non-permeable, ﬁne-earth fraction covering the soil surface (Lynch and Bragg, 1985) or development of a poorly permeable compacted layer with reduced root growth, lower soil moisture, and less soil oxygen (Thurrow, 1991). Lower soil moisture and inputs of organic carbon probably would reduce microbial biomass and activity levels, and net nutrient mineralization. Smaller soil pores associated with reduced aggregate structure reduce total inﬁltration, inﬁltration rates, and soil volume (Allison, 1973). All three conditions reduce the soil’s water holding capacity and thus the moisture available to support inputs of organic carbon and microbial decomposition (particularly microbial decomposition later in the growing season). Cryptobiotic crusts, grazing, and nutrient cycling Cryptobiotic crusts are surface mats of algae, fungi, lichens, mosses, and bacteria that mix and bind soil particles and chemical exudates (West, 1990; Johansen, 1993). They may form extensive mats under and between plants on relatively undisturbed arid rangelands (Johansen, 1993). The species mixture is speciﬁc to region and site and depends upon local soil conditions; however, these crusts develop best in arid shrublands with cool spring temperatures and moist soil (Johansen, 1993). Direct contact with the soil suggests that cryptobiotic crusts have important roles in nutrient cycling and that disturbing them can alter or break nutrient cycles. West (1990) critically reviewed the literature about cryptobiotic crusts and urged ecologists to conduct rigorously controlled experiments before accepting the proposition that these crusts are important in arid ecosystems globally. Johansen’s (1993) review also documented additional research needs. Despite strong contentions about the functional value of cryptobiotic crusts in arid rangelands, our discussion initially will assume that cryptobiotic crusts in arid ecosystems, to some degree, increase nutrient retention on-site and/or promote rapid nutrient cycling. We will also address problems with the current paradigm that welldeveloped cryptobiotic crusts always beneﬁt rangeland ecosystems. Cryptobiotic crusts may have important ecological functions, or they may be benign landscape

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features that become well developed with long return intervals for disturbances (e.g., lightning-caused ﬁre) that eliminate or reduce their cover. Most likely, their importance will be speciﬁc to site and crust type. Cryptobiotic crusts increase concentrations of soil organic carbon at and just below the soil surface (Fletcher and Martin, 1948; Shubert and Starks, 1978; Beymer and Klopatek, 1992). Because trampling by grazers reduces the cover of cryptobiotic crusts, it has been reasonably inferred that grazing reduces crust biomass and annual inputs of organic carbon to the soil (Marble and Harper, 1989; Beymer and Klopatek, 1992). Annual inputs of organic carbon from cryptobiotic crusts have rarely been measured (West, 1990; Johansen, 1993), but Jeffries et al. (1993a,b) estimated them at 6 and 23 kg ha−1 , respectively, in grazed and ungrazed blackbrush (Coleogyne ramosissima) communities. These values obviously are much less than above- and below-ground inputs of organic carbon from vascular plants. Inputs of organic carbon from cryptobiotic crusts probably have little value in cycling nitrogen, and their loss due to grazing probably has little effect on the inputs of organic carbon in the nutrient cycle. Additional links between the presence of cryptobiotic crusts and nutrient cycling include increased nitrogen from ﬁxation, nutrient retention, increased soil moisture, and increased seedling establishment. Each process, with or without cryptobiotic crusts, directly or indirectly interacts with grazing and nutrient cycling. The cyanobacteria (i.e., blue-green algae) are the primary nitrogen ﬁxers within the cryptobiotic crusts (West and Skujins, 1977; Rychert et al., 1978). Measured ﬁxation rates between 2 and 41 kg N ha−1 yr−1 have been recorded, but these may be overestimates (see West, 1990 and references therein). The functional beneﬁt of nitrogen ﬁxation by cryptobiotic crusts has been questioned, because only one ﬁeld study (West, 1990; Evans and Ehleringer, 1993) has linked nitrogen ﬁxed by cryptobiotic crusts with nitrogen in vascular plant tissues. This suggests that nitrogen inputs from cryptobiotic crusts are not necessary to maintain many vascular plant communities. Harper and Marble (1988) found most nitrogen ﬁxed by cryptobiotic crusts was lost by denitriﬁcation, resulting in minimal net gains. Other work found that water, not nitrogen primarily limits plant growth in arid shrublands (Romney et al., 1978). Additional nitrogen was only beneﬁcial with supplemental water. Fletcher and Martin (1948) found nitrogen levels four times higher in surface crusts than

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

in soils below them. This suggests that there may be small ﬂuxes of nitrogen from surface crusts to the root zone. Collectively, these data suggest that grazing levels which reduce nitrogen ﬁxing cryptobiotic crusts probably do not adversely affect nitrogen inputs to the vascular plant community, given current climatic patterns. One condition not addressed is the ecological importance of small surface accumulations of nitrogen through many decades or centuries. Short-term changes in nitrogen levels are relatively well understood. Nitrogen cycling associated with infrequent, intense disturbances, however, is not. Rough, intact crust surfaces may enhance nutrient concentrations (primarily calcium, magnesium, manganese, nitrogen, phosphorus and potassium) by trapping nutrient-rich eolian particles (Kleiner and Harper, 1972, 1977). It has, however, not been demonstrated that these nutrients are transported from crust surfaces to plant roots and shoots. Since nitrogen apparently accumulates in crusts, with minimal ﬂuxes to the root zone, it is probable that other nutrients accumulate on crust surfaces, but remain largely unavailable to vascular plants. Functionally beneﬁcial interactions between cryptobiotic crusts, nutrient accumulation, and nutrient cycling may be transport-limited, particularly in low-rainfall locations. If so, grazing that disturbs, reduces, or eliminates rough-surfaced crusts should not greatly reduce nutrient inputs to the soil. Observations that cryptobiotic crusts increase soil moisture (Gifford, 1972; Loope and Gifford, 1972) have been countered by quantitative studies ﬁnding a decrease (Graetz and Tongway, 1986; Harper and Marble, 1988; Verrecchia et al., 1995). Increased soil moisture, if present, should enhance inputs of organic carbon (from increased growth of vascular plants), microbial biomass, and/or the duration of microbial activity (Burke, 1989). Each of these results should increase total nutrient mineralization. Lower soil moisture should reduce nutrient mineralization. The presence of cryptobiotic crusts may decrease soil moisture through three mechanisms. First, trichomes from cyanobacteria can plug soil pores (Verrecchia et al., 1995). Second, ﬁne silt and clay trapped by rough crust surfaces swell when moist, plugging soil pores (Verrecchia et al., 1995). Third, lichen and moss crusts are often darker than the soil they cover, which increases soil temperature and evaporation (Harper and Marble, 1988). If cryptobiotic crusts reduce soil moisture, arguments that inputs of organic carbon from the crusts beneﬁt nutrient cycling may be incorrect.

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Most organic carbon inputs from cryptobiotic crusts are located at the soil surface or within several millimeters of it, not in the shallow root zone where microbial biomass and decomposition of organic matter are greatest (Bolton et al., 1993). Surface soils in arid locations with high solar insolation and cold wet winters generally are drier than subsurface soils. This is particularly true when soil temperatures are warm enough to permit high levels of microbial activity. Subsurface soil moisture and surface organic carbon inputs from cryptobiotic crusts may be spatially separated at times when soil temperatures are optimum for microbial activity. Interactions among grazing, cryptobiotic crusts, seedling establishment, and nutrient cycling have not been well documented (St. Clair et al., 1984; West, 1990). The most probable link is for plant succession to take different courses with and without cryptobiotic crusts and for the succession to change total root biomass and/or its distribution. An interaction between grazing and the crust which maintains high root biomass (particularly of herbaceous roots), shallow root depth (regardless of crust presence), and moist warm soils, will probably maintain relatively high cycling rates and an available supply of nutrients. The recovery of cryptobiotic crusts from grazing is speciﬁc to the crust and the landscape (Anderson et al., 1982; Johansen et al., 1984; Johansen and St. Clair, 1986; Cole, 1990). Algal mats return to pre-disturbance levels of percent cover in one to ﬁve years, while lichen and moss crusts may require 20 to 100 years (see Belnap, 1993 and references therein). Belnap (1993) and Belnap et al. (1994), however, found that levels of algal chlorophyll and nitrogenase activity recover much slower than either algal or crust cover. Their disturbance treatments, however (e.g., scalping, tracked military vehicles), often were much more intense than with properly managed grazing. Yet stated recovery rates may be misleading, as they usually report the time required for cover by the cryptobiotic crusts to reach predisturbance values, not predisturbance functional levels which may not be one and the same. Similar ecological function may occur at high and low values for cover, chlorophyll, or nitrogenase. Two hypothetical examples illustrate this point. First, nitrogen ﬁxed by a 20% cover of cyanolichen crust may substantially exceed the nitrogen incorporated into the vascularplant community. Restoration or maintenance of a 20% cyanolichen cover for purposes of nitrogen ﬁxation would be unnecessary if some lesser value would

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maintain functional equivalence. Second, mosses and fungi do not ﬁx nitrogen. Maintaining their cover values at high levels would be irrelevant to supporting nitrogen inputs. The importance of recovery or maintenance of cryptobiotic crusts, following disturbance by grazing, depends on the functional importance of the cryptobiotic-crust components. The functional importance of cryptobiotic crusts is not well known, and may vary spatially and temporally across different landscapes and ecosystems. One important point to remember, perhaps, is that many shrub-steppe sites, prior to the introduction of livestock grazing, are believed to have burned every 17 to 100 years (Burkhardt and Tisdale, 1976; Wright and Bailey, 1982; Miller et al., 1994). Fire can sharply reduce cover by cryptobiotic crusts (Johansen et al., 1984) and presumably their functional values. If the ﬁre frequencies stated are true, “fully intact” and “well developed” cryptobiotic crusts may have been rare and suggestions by Evans and Ehleringer (1993) and Belnap et al. (1994) that such cryptobiotic crusts are necessary to optimize nutrient cycles and/or maintain plant communities are misleading. Historic ﬁre frequencies and the time required to develop “fully” intact and “fully” functional cryptobiotic crusts would suggest that well-developed crusts are indeed artifacts of lengthened disturbance intervals. Factors that inﬂuence the recovery of cryptobiotic crusts following grazing-induced declines are: (1) the size of the area affected (Belnap, 1993); (2) the proximity of established crusts (Belnap, 1993); (3) soil moisture and temperature (Marble and Harper, 1989); (4) soil fertility and texture (Anderson et al., 1982); and (5) the length of the period between grazing events (Johansen and St. Clair, 1986; Marble and Harper, 1989). Large areas require longer recovery periods, particularly when crusts are entirely eliminated, because the spores necessary for recolonization must then come from adjacent crusts. Erratic precipitation or high soil temperatures (e.g., southern Great Basin) also slow recovery. Infertile soils have slow recovery because nutrients are insufﬁcient to support rapid growth by both the cryptobiotic crusts and the vascular plants. Fine-textured soils apparently enhance recovery because moisture capacity and nutrient availability are higher. Long periods without grazing permit better crust development, particularly for lichens and mosses.

Dale W. JOHNSON and Bradley SCHULTZ

Summary of rangeland disturbances High inputs of organic carbon (particularly from a labile substrate), moist soil, and warm soil temperatures drive decomposition of organic matter and nutrient cycling on native rangeland ecosystems. Carbon inputs are primarily from roots and plant litter. A grazing regime differing from that in which an ecosystem evolved may change inputs of organic matter at the community level. Redistribution of organic-carbon inputs occurs because grazing changes species composition to one dominated by grazing-tolerant or grazingresistant species. Grasses tolerant of grazing typically provide greater inputs of organic carbon at shallow depths than grasses not tolerant of grazing. Shrubs may also replace grasses sensitive to grazing. This redistributes root biomass to deeper soil horizons, and concentrates surface litter under shrub canopies. Increased inputs of organic matter from shrubs also reduce the quality of the substrate (i.e., lower concentrations of labile nutrients) resulting in lower turnover rates, particularly in arid climates, where limited soil moisture reduces the period of microbial activity. Increased amounts of labile organic carbon, without a concomitant increase in mineralized nitrogen, result in soil microbe populations becoming nitrogen-limited with most nitrogen sequestered in their biomass and unavailable for plant uptake and return to the grazers. Grazers increase nutrient cycling when they quickly transfer organic matter from above-ground primary production to organic matter. This can occur through the digestion process or by trampling that converts standing organic matter to surface plant litter which can be readily incorporated into the soil. Annual precipitation varies widely among landscapes and can be strongly redistributed within landscapes. This variation results in wide spatial and temporal variation for soil moisture levels, and hence in inputs of organic carbon and microbial activity. Locations that predominately receive their precipitation during cold winters have shorter periods of microbial activity. Inputs of labile organic carbon, moist soil, and warm soil temperatures only coincide for brief periods early in the growing season. Grazing that reduces inputs of labile organic carbon (e.g., from herbaceous species) and/or increases inputs of organic carbon from poor-quality substrates (e.g., shrubs) results in slower turnover rates and fewer available nutrients. Excessive trampling or defoliation by grazers decreases soil aggregate stability and surface plant litter.

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS

Good aggregate structure and litter cover slow the movement of water across the soil surface, increasing inﬁltration, and decreasing runoff and wind erosion. Increased inﬁltration may increase soil moisture levels, increasing microbial activity. Decreased runoff reduces nutrient transport off-site, maintaining or increasing soil fertility. Cryptobiotic crusts, because of their intimate association with the soil, are thought to beneﬁt nutrient cycling, but this belief has not been well substantiated by rigorous experimental research (but see Evans and Ehleringer, 1993). These crusts are easily reduced by grazing ungulates; therefore, reduced nitrogen input is possible. Fixed nitrogen and eolian-deposited nutrients may not be transported to the root zone or incorporated into vascular plant tissues. Observations that cryptobiotic crusts enhance soil moisture, which should beneﬁt nutrient mineralization, have not been supported by experimental research. Crusts with rough surfaces appear to trap eolian-deposited clay and silt particles; however, they may become non-permeable barriers upon wetting, because precipitation results in the eolian deposited material swelling and plugging soil pores. Trichomes from cyanobacteria in the crusts also swell, further limiting soil porosity. Cryptobiotic crusts may be present in undisturbed landscapes because of long intervals between disturbances that reduce their cover. Their functional value may be substantially less than is commonly believed.

DIFFERENCES AND SIMILARITIES BETWEEN FOREST AND RANGELAND DISTURBANCES

Nitrogen is the most commonly limiting nutrient in both forest and rangeland ecosystems and thus factors affecting nitrogen cycling are of critical importance. In forest ecosystems, disturbance most often leads to increases in nitrate leaching, whereas leaching is minimal in most rangelands and denitriﬁcation is probably the most important (and most difﬁcult to measure) pathway of nitrogen loss. In both desert and rangeland systems, there is competition among heterotrophs, roots, and nitriﬁers for ammonium ions. In both cases, it appears that heterotrophs are the most effective competitors, followed by plant roots and nitrifying bacteria, in that order. The two primary disturbances considered here – harvesting and grazing – have in common the disruption of nutrient cycles and the removal of nutrients in

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biomass. The two types of disturbance differ markedly with respect to how rapidly the vegetation recovers from biomass removal and the substantial amount of nutrients (60–90%) which grazers return to the soil in their excreta. The two disturbances differ with respect to frequency and the rapidity with which vegetation recovers from the disturbance. However, in each case the primary effect of disturbance is to cause an increase in the rate of nitrogen cycling and an attendant increase in the rate of nitrogen loss from the system. The magnitudes of these losses are much better known for forest than for range ecosystems, but in both cases the interactions of microbes and vegetation are critical to the intensity and duration of loss. ACKNOWLEDGMENTS

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Dale W. JOHNSON and Bradley SCHULTZ Polley, H.W. and Detling, J.W., 1989. Defoliation, nitrogen, and competition: effects on plant growth and nitrogen nutrition. Ecology, 70: 721–727. Reynolds, B., Stevens, P.A., Hughes, S., Parkinson, J.A. and Weatherly, N.S., 1995. Stream chemistry impacts of conifer harvesting in Welsh catchments. Water Air Soil Pollut., 79: 147–170. Richards, J.H., 1984. Root growth response to defoliation in two Agropyron bunchgrasses: ﬁeld observations with an improved root periscope. Oecologia, 64: 21–25. Richardson, C.J. and Lund, J.A., 1975. Effect of clearcutting on nutrient losses in aspen forests on three soil types in Michigan. In: F.G. Howell, J.B. Gentry and M.H. Smith (Editors), Mineral Cycling in Southeastern Ecosystems. ERDA Symposium Series CONF 740513, pp. 673–686. Rickard, W.H., Cline, J.F. and Gilbert, R.O., 1973. Soil beneath shrub halophytes and its inﬂuence upon the growth of cheatgrass. Northwest Sci., 47: 213–217. Ring, E., 1995. Nitrogen leaching before and after clear-felling of fertilised experimental plots in a Pinus sylvestris stand. For. Ecol. Manage., 72: 151–166. Risser, P.G. and Parton, W.J., 1982. Ecosystem analysis of the tallgrass prairie: nitrogen cycle. Ecology, 63: 1342–1351. Rodriguez, M.A., Brown, V.K. and Gomez-Sal, A., 1995. The vertical distribution of below-ground biomass in grassland communities in relation to grazing regime and habitat characteristics. J. Vegetation Sci., 6: 63–72. Rodriguez, M.A., Alvarez, J. and Gomez-Sal, A., 1996. Vertical distribution of below-ground biomass in intensively grazed mesic grasslands. J. Vegetation Sci., 7: 137–142. Romney, E.M., Wallace, A. and Hunter, R.B., 1978. Plant response to nitrogen fertilization in the northern Mojave Desert and its relationship to water manipulation. In: N.E. West and J. Skujins (Editors), Nitrogen in Desert Ecosystems. Dowden, Hutchinson, and Ross, Stroudsburg, Pennsylvania, pp. 20–30. Rosen, K. and Lundmark-Thelin, A., 1987. Increased nitrogen leaching under piles of slash – a consequence of modern forest harvesting techniques. Scand. J. For. Res., 2: 21–29. Ruess, R.W. and McNaughton, S.J., 1984. Urea as a promotive coupler of plant–herbivore interactions. Oecologia, 63: 331–337. Ruess, R.W. and McNaughton, S.J., 1987. Grazing and the dynamics of nutrient and energy regulated microbial processes in the Serengeti grasslands. Oikos, 49: 101–110. Ruess, R.W. and Seagle, S.W., 1994. Landscape patterns in soil microbial processes in the Serengeti National Park, Tanzania. Ecology, 75: 892–904. Ruess, R.W., McNaughton, S.J. and Coughenour, M.B., 1983. The effects of clipping, nitrogen source, and nitrogen concentration on the growth response and nitrogen uptake of an East African sedge. Oecologia, 59: 253–261. Rychert, R., Skujins, J., Sorensen, D. and Porcella, D., 1978. Nitrogen ﬁxation by lichens and free-living microorganisms in deserts. In: N.E. West and J. Skujins (Editors), Nitrogen in Desert Ecosystems. Dowden, Hutchinson, and Ross, Stroudsburg, Pennsylvania, pp. 20–30. Seagle, S.W., McNaughton, S.J. and Ruess, R.W., 1992. Simulated effects of grazing on soil nitrogen and mineralization in contrasting Serengeti grasslands. Ecology, 73: 1105–1123. Shubert, L.E. and Starks, T.L., 1978. Algal succession on orphaned

CARBON AND NITROGEN CYCLES IN FORESTS AND RANGELANDS coal mine spoils. In: M.K. Wali (Editor), Ecology and Coal Resource Development. Pergamon Press, New York, pp. 661–669. Smoliak, R., Dormaar, J.F. and Johnston, A., 1972. Long-term grazing effects on Stipa–Bouteloua prairie soils. J. Range Manage., 25: 246–250. St. Clair, L.L., Webb, B.L., Johansen, J.R. and Nebeker, G.T., 1984. Cryptogamic soil crusts: enhancement of seedling establishment in disturbed and undisturbed areas. Reclamation Revegetation Res., 3: 129–136. Staaf, H. and Olsson, B.A., 1994. Effects of slash removal and stump harvesting on soil water chemistry in a clearcutting in SW Sweden. Scand. J. For. Res., 9: 305–310. Stark, J.M. and Hart, S.C., 1997. High rates of nitriﬁcation and nitrate turnover in undisturbed coniferous forests. Nature, 385: 61–64. Stevens, P.A. and Hornung, M., 1988. Nitrate leaching from felled Sitka spruce plantation in Beddgelert Forest, North Wales. Soil Use Manage., 4: 3–9. Stevens, P.A. and Hornung, M., 1990. Effect of harvest intensity and ground ﬂora establishment on inorganic nitrogen leaching from a Sitka spruce plantation in north Wales, UK. Biogeochemistry, 10: 53–65. Sturges, D.L., 1977. Soil water withdrawal and root characteristics of big sagebrush (Artemisia tridentata) in the United States. Am. Midl. Nat., 98: 257–274. Sturges, D.L., 1986. Responses of vegetation and ground cover to spraying a high elevation big sagebrush catchment with 2,4-D. J. Range Manage., 39: 141–146. Svejcar, T. and Christiansen, S., 1987. The inﬂuence of grazing pressure on rooting dynamics of Caucasian bluestem. J. Range Manage., 40: 224–227. Swank, W.T., Waide, J.B., Crossley, D.A. and Todd, R.L., 1981. Insect defoliation enhances nitrate export from forest ecosystems. Oecologia, 51: 297–299. Thurrow, T.L., 1991. Hydrology and erosion. In: R.K. Heitschmidt and J.W. Stuth (Editors), Grazing Management. An Ecological Perspective. Timber Press, Portland, Oregon, pp. 141–160. Thurrow, T.L., Blackburn, W.H. and Taylor Jr., C.A., 1986. Hydrologic characteristics of vegetation types as affected by livestock grazing systems, Edwards Plateau, Texas. J. Range Manage., 39: 505–509. Tomanek, G.W., 1969. Dynamics of mulch layer in grassland ecosystems. In: R.L. Dix and R.G. Beidleman (Editors), The Grassland Ecosystem. Science Series 2. Range Science Department, Colorado State University, Fort Collins, Colorado, pp. 225–240. Van Miegroet, H. and Cole, D.W., 1984. The impact of nitriﬁcation on soil acidiﬁcation and cation leaching in a red alder forest. J. Environ. Qual., 13: 585–590. Van Miegroet, H., Homann, P.S. and Cole, D.W., 1992. Soil nitrogen dynamics following harvesting and conversion of red alder and Douglas-ﬁr stands. Soil Sci. Soc. Amer. J., 56: 1311–1318. Verrecchia, E., Yair, A., Kidron, G.J. and Verrecchia, K., 1995. Physical properties of the psammophile cryptogamic crust and their consequences to the water regime of sandy soils, north-western Negev Desert, Israel. J. Arid Environ., 29: 427–437. Vitousek, P.M., 1981. Clearcutting and the nitrogen cycle. Ecol. Bull. (Stockholm), 33: 631–642.

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Chapter 24

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS Ian K. BRADBURY

INTRODUCTION

A disturbance is deﬁned here as the comparatively sudden loss, or the onset of loss, of a signiﬁcant quantity of biomass over an “ecologically signiﬁcant” area. For convenience, a minimum area is set at around 30 m2 . However, responses to local disturbance may aggregate to have a major inﬂuence on the ecology of landscapes. The term “intensity” is used here when referring to the agent of disturbance (e.g., ﬁre temperature, windspeed, number of grazers), while “severity” is used when referring to the degree of impact of the disturbance event on the vegetation or on the soil. Intuitively, one can envisage both positive and negative effects of disturbance on primary productivity. In some situations, severe disturbance may lead to a deterioration in site conditions, with a consequent decline in productivity. In other situations, certain types and intensities of disturbance may result in an increase in productivity, for example by encouraging the growth of younger tissues or by stimulating nutrient cycles. The long history of ﬁre use by humans in vegetation is a testimony to the stimulating effects of burning on productivity in some situations. The relationship between disturbance and primary productivity is unlikely to be simple. First, each disturbance agent elicits its own characteristic suite of ecological responses within a particular type of community. Second, the effect of each type of disturbance in a particular community is inﬂuenced by the timing and intensity of the disturbance, the disturbance history of the site, and the physiological state of the constituent

plants. Nonetheless, it is reasonable to anticipate some patterns in response of ecological processes which may be expressed as alterations in primary productivity. The key components of plant production are net photosynthetic rate (usually expressed per unit leaf area), leaf area index (ratio of leaf area per unit area of ground), ratio of photosynthetic to non-photosynthetic tissue (usually represented by the leaf area ratio), and leaf area duration1 (Mooney and Gulman, 1983; Lambers and Poorter, 1992). By inﬂuencing the components of production differentially it is highly probable that a disturbance event will change the rate of dry matter accumulation in the short term. Disturbance is also likely to alter the availability of resources at a site – for example, by modifying microclimate or stimulating nutrient mineralization. As primary productivity is largely controlled by resource availability (Mooney and Gulman, 1983), a change in primary productivity may be a normal consequence of disturbance. Because of the formidable problems involved in accurately determining primary productivity, it is necessary to evaluate very carefully the methodology employed when primary productivity data are presented. This is particularly so when comparative data are presented for the same locality, as for example, in comparing undisturbed and disturbed vegetation stands or comparing pre- and post-disturbance stands on the same site. If two stands are spatially separated, an observed difference in productivity could be due to site characteristics, whereas if measurements are made on the same site in different years differences in productivity could be due to the weather. These problems are magniﬁed when comparisons are being

1

Leaf area ratio, the ratio of total leaf area to whole-plant dry weight; leaf area duration, the integral of the area beneath the curve relating leaf area index to time (a measure of ‘leaﬁness’ over time).

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made between vegetation of different life-forms, as may well be the case when pre- and post-disturbance situations are being studied. Care is also needed when attempting to characterize the disturbance event (Sousa, 1984). From a single disturbance it is impossible to derive a relationship between the intensity of the disturbance and a response variable of interest, and it cannot be assumed that the relationship will be monotonic. Normally, trends in response to disturbance intensity have been established by comparing responses to disturbance in broadly similar circumstances. Controlled experimental studies are not without their problems either. Artiﬁcially defoliating vegetation, for example, while permitting a range of disturbance intensities to be examined, may not faithfully reproduce grazing (Jameson, 1963). This review focuses on the response of primary productivity to two key ecological factors, ﬁre and animal activity. The productivity response to other agents of disturbance is only sparsely represented in the literature. Throughout, the intention is to highlight those mechanisms, both biotic and abiotic, by which ﬁre and grazing may inﬂuence primary productivity. Case studies are selected from a range of community types and a diversity of geographical provenances.

FIRE AND PRODUCTIVITY

This section begins with a brief consideration of postﬁre productivity in different categories of vegetation. This is followed by an examination of the mechanisms by which ﬁre may inﬂuence productivity, both in the immediate aftermath of a burn and over longer timescales. Forests The ecological effects of ﬁre within a forest are determined largely by the vertical distribution of the combusted material. At one extreme, surface ﬁres may temporarily alter vegetation structure near the ground but have little effect on tree growth. At the other extreme, a crown ﬁre may leave a site devoid of living vegetation. Characteristic post-ﬁre ecological changes vary between forest types. In some situations, the dominant tree species may be replaced, from seed or vegetatively, by the same species. In other situations, decades may elapse before the pre-ﬁre complement of species is re-established.In the former

Ian K. BRADBURY

case, explanations for productivity differences would involve a consideration of stand age and site conditions, while in the latter case differences in productivity would also be inﬂuenced by species composition and vegetation structure. Because annual net primary productivity (ANPP) is not constant over the growth cycle of a tree stand (Peet, 1981), the timing of a stand-replacing burn inﬂuences the productivity response. The pattern of decline in annual net primary productivity of tree stands in later growth stages has been attributed principally to nutrient limitation (Sprugel, 1985), but there is no consensus on this issue (Peet, 1992). Major ﬁres in old moribund tree stands are therefore likely to raise annual net primary productivity in the short term, partly because of the greater photosynthetic capacity of the vegetation that re-establishes and partly because of increased nutrient availability. Other possible effects on site conditions are dealt with later. Fire is a major, and frequently-occurring, disturbance agent throughout much of Australia’s forested area. Typically, forests of the seasonally-dry and subhumid zones are relatively open with a well-developed herb and/or shrub layer. Most of the ﬁve hundred or so species of Eucalyptus s.l. in Australian forests sprout readily after ﬁre damage, while the remainder are equipped to survive ﬁre, for example by massive seed germination (Gill, 1981). Christensen et al. (1981) commented on the “rejuvenating” effect of ﬁre in Australian open forest and described the post-ﬁre response of vegetation as “obvious and often spectacular”. They attributed this response primarily to the release of nutrients from highly ﬂammable litter into soils which are inherently infertile. Whelan (1995) suggested that suppression of competition and the proliferation of shoots with relatively high photosynthetic capacity may contribute to enhanced post-ﬁre growth. The relationship between enhanced growth and changed site characteristics is only correlative: controlled and replicated experimentation is required to establish causation (Whelan, 1995). In North America, the frequency of ﬁre and its effects vary considerably between regions and forest types (Ahlgren and Ahlgren, 1960; Wright and Bailey, 1982). Much of the evidence for ﬁre effects on primary productivity remains circumstantial. In general, however, the longer the interval between a major ﬁre and the closure of the tree canopy, the lower the overall productivity. For some species, such as Pinus banksiana (jack pine) and Pinus resinosa (red

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS

pine), successful seedling establishment depends on the removal of the humus layer, by ﬁre or tree uprooting, and exposure of mineral soil. Under these conditions vigorous growth occurs in the ﬁrst year after burning (Ahlgren, 1974). In contrast, if an organic layer remains after ﬁre, pine seedlings, with their slow vertical root growth, will dry out during the summer months, and colonization by coniferous trees will be delayed (Ahlgren, 1974). Burnt sites have signiﬁcantly higher maximum temperatures and longer growing seasons than comparable unburnt sites, which will potentially raise primary productivity. Shrublands Many shrublands periodically experience ﬁre. In addition to lightning-induced burns, ﬁre may be started accidentally (Le Hou´erou, 1981) or used as a management tool (Gimingham, 1972). In ﬁre-prone communities, the survival of most shrub species is attributable to rapid resprouting from shoot bases or underground organs, although very severe ﬁre increases the importance of regeneration from seed. Some shrubs, however, including most Artemisia (sagebrush) species of the western United States, may be severely damaged by ﬁre (Wright and Bailey, 1982). Specht (1969) compared post-ﬁre growth of shrubby sclerophyllous vegetation in southern France (Quercus coccifera garrigue), California (Adenostoma fasciculatum – Ceanothus chaparral) and Australia (Eucalyptus incrassata – Melaleuca uncinata mallee). In all these communities, post-ﬁre regeneration occurs principally from lignotubers or crown burls. In each case, photosynthetic tissue dominated production during the recovery phase, although the proportional contribution of non-photosynthetic tissue increased progressively. Underground biomass remained relatively constant between ﬁres, but maximum root turnover occurred immediately after ﬁre. Annual shoot productivity in each community was greatest in the ﬁrst post-ﬁre year, and declined progressively thereafter. Specht (1966) measured annual shoot production in a Eucalyptus incrassata–Melaleuca uncinata shrubland in South Australia over a 12-year interval following a ﬁre. Here, return times for severe burns vary between 15 and 100 years. Shoot growth was at its highest in the ﬁrst post-ﬁre year but had declined considerably by the twelfth year. In British heathlands, annual shoot production of the dominant shrub Calluna vulgaris increased rapidly following prescribed burning, but

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after 25–30 years plants began to degenerate (Miller, 1979; Hobbs and Gimingham, 1987). In chaparral in California (U.S.A.), return times for wildﬁre are typically between 30 and 50 years (Rundel, Chapter 10, this volume), although the natural ﬁre regime has been altered by human activities for at least the last few hundred years (Hanes, 1981). Relatively old (30–50 yr) stands of chaparral appear moribund and poorly productive. After a ﬁre, around half of the shrubs present resprout from lignotubers, often within a few weeks. Vigorous shoot growth is a normal post-ﬁre response (Hanes, 1981), although this may be at least partly due to the translocation of stored assimilates from below-ground tissues. A proliferation of seedling establishment also typically occurs after a ﬁre in chaparral, prompted largely by seed scariﬁcation and the removal of germination inhibitors (Christensen and Muller, 1975a,b). Enhanced nutrient availability may also play a role in vegetation recovery. Several studies have reported enhanced nitrogen and phosphorus concentrations in chaparral soils after ﬁre (Hanes, 1981). The relationship between site nutrient status and productivity was examined experimentally by Godron et al. (1981) in garrigue in southern France dominated by Quercus coccifera. Fertilization treatments, which mimicked post-ﬁre nutrient enrichment, resulted in a ﬁve-fold increase in total annual shoot production compared to non-fertilized controls, most of the increased production being accounted for by grass species. Grasslands and other herbaceous vegetation Fire is an important ecological factor in many of the world’s grasslands. An international program on productivity in tropical and subtropical grasslands in the 1980s has provided some reliable data on ﬁre effects (Long et al., 1992). In these studies special attention was paid to underground tissues and the loss of dry matter between harvests. At a location in southern Thailand, an accidental ﬁre occurred during winter, which is a dry interval in this wet-monsoonal climate. Shoot regrowth was rapid, and biomass was similar to that on unburnt plots after ﬁve months (Kamnalrut and Evenson, 1992). However, after a second ﬁre thirteen months later, biomass values remained well below those on unburnt plots. The second ﬁre also caused a marked decline in underground biomass, presumably due to the translocation of carbohydrate

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to shoots. These results demonstrate the dangers of making generalizations from a single disturbance event. In a study of saline grassland near Mexico City dominated by Distichlis spicata (Garcia-Moya and Castro, 1992), a January ﬁre delayed spring growth, but thereafter recovery was rapid, and peak values in the following two seasons were only slightly less than preﬁre values. However, below-ground biomass declined markedly after the ﬁre – again presumably because of the translocation of stored assimilates to developing shoots. Hummock grasses, often with an overstorey of woody vegetation, cover approximately one-third of Australia (Suijdendorp, 1981). Over this vast area, annual rainfall is generally less than 400 mm, and lightning-induced ﬁres are frequent during dry thunderstorms in summer, although wildﬁres have long been supplemented by ﬁres set by aborigines. European settlers have markedly altered previous ﬁre regimes by annual winter burns. The result has been characterized as “disastrous” for productivity, particularly when carried out in association with heavy grazing by sheep (Suijdendorp, 1981). Part of the deleterious effect of winter ﬁres is associated with a shift in production from grasses to less palatable annual forbs. Suijdendorp (1981) advocated a return to summer ﬁres at ﬁve-year intervals with judicious use of grazing, in order to maximize the production of the more desirable grasses and inhibit scrub encroachment. The role of ﬁre in North American grasslands has been a focus of study throughout the twentieth century (Weaver, 1954), and the subject has been comprehensively reviewed (Daubenmire, 1968; Vogl, 1974; Wright and Bailey, 1982; Kucera, 1992). Much of the evidence for production responses to ﬁre comes from the range-management literature, and data are usually conﬁned to above-ground tissue. In the tallgrass prairie, enhanced yields of forage following burning have been demonstrated for many grass species although there are notable exceptions, particularly among the cool-season grasses (Wright and Bailey, 1982). In the mixed prairie, the response to ﬁre is weather dependent: some grasses, for example Bouteloua curtipendula (sideoats grama) and Stipa leucotricha (Texas wintergrass) may be severely reduced by ﬁre in years of average rainfall (Wright and Bailey, 1982). However, it is not clear whether the effect of moisture is mediated through ﬁre temperature or the physiological state of the plants at the time of the burn. For shortgrass prairie, Wright and Bailey (1982)

Ian K. BRADBURY

concluded that a ﬁre in dry years is usually detrimental to most species, with a delay of around three years before preburn yields are attained. In wetter years in contrast, recovery is much more rapid, and there may be no loss of yield during the ﬁrst post-ﬁre growing season. The balance between the graminoid and forb components during post-ﬁre recovery largely depends on the season of ﬁre. Forbs are much less vulnerable to autumn burns than to burns in spring when growth is rapid. Although caution is needed when comparing data from studies differing in methods, ﬁre intensity, species, phenological state, and weather, it appears that in North American grasslands ﬁre generally promotes production in the more humid tallgrass prairie, but the growth-enhancing effect declines with average rainfall, and may be reversed in the drier parts of the grassland zone. Over extensive tracts of western North America, herbaceous vegetation forms an understorey in open woodland, and management is frequently directed towards encouraging forage production. Here, repeated ﬁres tend to shift the balance in favour of grasses and forbs (Vogl, 1974). Litter removal by ﬁre can markedly increase production, as has been shown in northern Arizona grasses (Pearson et al., 1972), and in the open Pinus ponderosa-dominated forests of eastern Oregon and Washington (Weaver, 1974). Mechanisms for post-ﬁre changes in productivity In this section the processes which link ﬁre to primary productivity are examined. While the magnitude of response varies between community types, processes of general signiﬁcance include modiﬁcation of microclimate, changes in nutrient availability, soil erosion, and the removal of soil organic matter. Factors which are commonly invoked to explain enhanced productivity after ﬁre include increased nutrient and moisture availability, a more favourable temperature regime for biotic activity, the removal of phytotoxic substances in litter, and an increase in the ratio of photosynthetic to non-photosynthetic tissue (Daubenmire, 1968; Vogl, 1974; Kucera, 1992; Whelan, 1995). Because several factors operate simultaneously, it is often difﬁcult to assess the relative contribution of each one. There is widespread agreement that ﬁre has a major impact on nutrient dynamics. Woodmansee and Wallach (1981) categorized nutrient responses to ﬁre as either primary, secondary, or tertiary. The key primary

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS

responses are nutrient volatilization and ash deposition. The secondary responses, which are abiotic, include erosion by wind and water, leaching of nutrients, and the adsorption of cations on soil colloids. Tertiary responses, which are biotic, include nitrogen ﬁxation and transformations, and nutrient uptake by plants and microorganisms. From this scheme it is clear that ﬁre can raise soil nutrient availability in the short term, but also lead to nutrient loss from ecosystems. Whatever the longer-term effects of ﬁre on nutrient budgets, it is almost axiomatic that the combustion of vegetation and litter raises the concentration of available nutrients near the soil surface (Whelan, 1995). This is partly due to the ashing of vegetation and litter, and partly to the change in site conditions in the post-ﬁre period. Higher temperatures, due to a lower albedo, stimulate nitrogen mineralization and nitrogen ﬁxation by both free-living and symbiotic nitrogenﬁxing microbes (Rundel, 1981). Plant taxa which form associations with nitrogen-ﬁxing symbionts, Ceanothus species for example, are characteristic of some disturbed areas. Fire also tends to increase the pH of acidic soils, which may further stimulate nutrient availability. In Californian chaparral, Rundel (1983) attributed nitrogen increases to direct chemical alteration of organic matter by high temperatures and, possibly, the post-ﬁre stimulation of free-living nitrogen-ﬁxing bacteria. Christensen (1973) suggested that post-ﬁre increases of soil nitrogen in chaparral were due to enhanced nitriﬁcation rather than conversion of organic nitrogen to ammonium compounds. The role of nitrogen in post-ﬁre recovery in shrublands in the western United States was indicated by fertilization experiments (Vlamis and Gowans, 1961; DeBano and Conrad, 1978). In these studies, nitrogen additions elicited no increase in production on burnt plots, but there was a positive response in unburnt controls. Despite the tendency for ﬁre to raise nutrient availability, it cannot be assumed that this factor is always responsible when enhanced productivity is demonstrated (Raison, 1979). In grasslands, for example, Daubenmire (1968) suggested that the fertilizing effects of ﬁre are generally small and much less important than the direct effects of litter removal. Evaluating the relative contribution of different factors to enhanced productivity requires careful, and replicated, experiments. A good example of this approach is the work of Hulbert (1988) in tallgrass prairie in Kansas (U.S.A.), where early-season ﬁre

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generally leads to increased productivity. In this study, treatments comprised burning, artiﬁcial soil heating, clipping, additions of ash to unburned plots, shading, and nitrogen fertilization. The results suggested that post-burn increases in productivity were due to the combined effect of increased radiation and nitrogen availability, with light playing a key role. The effect of ﬁre in temporarily raising nutrient availability may partly explain why non-mycorrhizal plant species are frequently associated with disturbed ground (Newman, 1988; Brundrett, 1991). This is because mycorrhizal associations appear to be least advantageous to host plants in nutrient-rich situations. While ﬁre normally results in some increase in soil nutrient status in the short term, it also causes the loss of nutrients by volatilization, and predisposes nutrients to loss by erosion. Volatilization losses depend on temperature and the physicochemical properties of the element in question. The highest ﬁre temperatures (800–1000ºC) appear to be associated with forestry operations when considerable quantities of combustible material (slash and logs) remain at the surface, particularly when these are moved into windrows (Rundel, 1981). Shrub-dominated communities (e.g., chaparral in California and older stands of Calluna vulgaris in British heathland) can burn at temperatures well in excess of 600ºC. Signiﬁcant losses of nitrogen and sulphur by volatilization can be expected at temperatures above 200–300ºC (Rundel, 1981; Wright and Bailey, 1982). Losses of phosphorus, although variable, are generally much less than for nitrogen. Losses of cationic elements from Calluna-dominated heathland in Britain tended to increase with temperature but remained low overall except in burns of very high intensity (Allen, 1964; Evans and Allen, 1971). Nutrient dynamics may also be inﬂuenced by the distribution of the nutrient store between the aboveand below-ground plant components. Rundel (1983) suggested that the nitrogen balance of communities with characteristically high ratios of below- to aboveground biomass (e.g., tundra, desert, grasslands) is little affected by ﬁre. The effects of ﬁre on erosion, and other geomorphic processes, have been reviewed by Swanson (1981). Erosion is an effective mechanism for nutrient loss, and alters site characteristics in other ways which potentially reduce productivity. Vegetation and litter removal can precipitate catastrophic erosion in some situations, but more typically the erosional response is insidious and its consequences for primary productivity

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are extremely difﬁcult to quantify. The key variables controlling the amount of erosion after a burn are soil characteristics, steepness and length of slope, season, the amount and intensity of rainfall, vegetation type and ground cover, severity of ﬁre, and length of time since the previous burn. While it should usually be possible to gauge erosion risk at a particular site from the prevailing conditions, some relationships between site variables and erosion potential may be counterintuitive. In the Californian chaparral, for example, Biswell (1974) cited evidence that dry soils on steep slopes are less cohesive than wet soils, and hence subject to greater rates of removal downslope following severe ﬁres. In Californian chaparral, particularly where topography is steep, soil erosion usually accelerates after a ﬁre (Biswell, 1974; Rundel, 1983). On steep, unconsolidated slopes, the replacement of deep-rooted chaparral shrubs by grass species exacerbates the erosion hazard, although this is not necessarily the case on gentle slopes. The rate of recolonization by vegetation has a major inﬂuence on post-ﬁre erosion, which typically peaks in the ﬁrst winter after a ﬁre. In contrast to the chaparral, post-ﬁre erosion risks over most of the grasslands of the North American Great Plains are relatively low on account of low fuel loads, low ﬁre intensity, and gentle topography (Daubenmire, 1968). Similarly, in the savanna zone of West Africa, annual burns normally appear to have little effect on site conditions (Sanford and Isichei, 1986). The risk of erosion is increased when ﬁre leads to the development of water-repellent layers in the soil, as happens in Californian chaparral (Biswell, 1974). These form when hydrophobic substances from plant litter are carried into the surface soil by percolating rainwater. Intense ﬁres volatilize these substances, which then diffuse downwards before condensing on soil particles. For site productivity the signiﬁcance of water-repellent layers lies in their role in hydrologic budgets. If inﬁltration of rainwater is impeded, particularly during storm events on unvegetated steep slopes, soil losses may be very high. Coarse-textured soils, which are common in chaparral, are particularly susceptible to the development of water repellency because of a relatively low surface area per volume of soil (Biswell, 1974). Although the mechanisms by which ﬁre can reduce site productivity are well known, in practice it is extremely difﬁcult to evaluate the effects of ﬁre on

Ian K. BRADBURY

productivity, particularly over comparatively long timescales. Key issues are the extent to which, ﬁrstly, nutrient losses can be replaced by inputs between ﬁres and, secondly, other post-ﬁre physicochemical and biotic effects are reversible. Intuitively, nutrient losses have greatest impact on sites of low fertility – for example, heathland in western Europe. Even here it is suggested that, with ﬁre temperatures below about 400ºC, nutrient losses, with the possible exception of phosphorus, should be replaced by aerial input within a few years (Hobbs and Gimingham, 1987). In some situations, however, it is reasonable to assume that repeated ﬁres lead to lower soil nutrient status, and thus lowered productive potential, particularly when longer time-scales are considered. Such a scenario has been postulated by Le Hou´erou (1981) for the drier parts of the Mediterranean Basin, where several thousands of years of anthropogenic ﬁres have resulted in reduced nutrient and moisture availability as a consequence of erosion. It was claimed also that increased ﬁre frequency favoured pyrophytic species, which are generally less productive than nonpyrophytic species (Le Hou´erou, 1981). Fire may also inﬂuence site productivity by the combustion of soil organic matter, particularly in surface horizons (Gimingham, 1981; Rundel, 1981; Wright and Bailey, 1982). Humus plays a key role in cation exchange, as a store of nitrogen, in determining hydrologic characteristics, and in insulating the underlying mineral horizons. Also, the organic horizon may be the principal, sometimes the only, site for root development and nutrient recycling, as well as the main store of seeds and organs of vegetative regeneration. The effect of loss of organic matter on primary productivity is, therefore, likely to be most pronounced where a relatively thin organic rooting zone lies directly upon, or close to, the unweathered parent material, as is the case over large tracts of Shield country in Canada (Wright and Bailey, 1982). Even where mineral soil underlies organic horizons, hightemperature burns can considerably impair site recolonization from propagules and perennating organs. It is suggested that, under some conditions, the removal of surface organic matter effectively raises the soil water table, which again has implications for primary productivity (Gimingham, 1981). Although the combustion of humus may often lower site productivity, there are situations where vegetation recolonization is fostered by organic removal. As discussed earlier, there is evidence that the successful

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS

regeneration of pine species in the north-central United States depends on the exposure of mineral soil. Using such evidence, MacLean et al. (1983) have conjectured that the latitudinal tree-line in North America is determined by the ability of ﬁre to remove surface organic matter and expose the underlying mineral horizons. Where organic horizons are not removed, regeneration of spruce and pine is prevented. Over much of the northern coniferous forest and its outliers in North America, cool and often wet conditions inhibit decomposition and encourage the development of a carpet of bryophytes, principally Sphagnum and feather mosses (Heinselman, 1981). In consequence, nutrients are progressively immobilized, and soil temperatures in summer are lowered by the insulating effect of humus and moss. Fire plays a key role in maintaining productivity in such environments, by stimulating nutrient release from dead organic matter, and by altering site conditions and microclimate. The stimulation of the nitrogen cycle is particularly important in raising productivity (MacLean et al., 1983). The prolonged absence of ﬁre, or its failure to remove surface organic matter, can lead to the gradual transformation of forest to muskeg (Van Cleve and Viereck, 1981). Long ﬁre-free intervals on inherently wet sites have led to the formation of extensive peatlands in North America, such as those of the Lake Agassiz plain in Manitoba, Minnesota, and Ontario, and in the Hudson Bay lowlands (Heinselman, 1981). Although burning is usually associated with the loss of surface organic matter, there is evidence that ﬁre can initiate peat formation in some situations, with consequences for primary productivity. Wein (1983) suggested that this may occur when the loss of vegetation causes a reduction in the amount of water returned to the atmosphere, with a consequent rise in the water table. With a continuously high water table, the vegetation shifts from forest to peat-forming mosses. Primary productivity would be lowered in this situation, ﬁrst by the replacement of vascular plants by mosses with their inherently low productivity, and second by nutrient limitations (Heilman, 1966; MacLean et al., 1983). Fire plays an important role in those parts of the permafrost zone which support closed forest and foresttundra. (Further north, the true tundra only rarely supports ﬁre from natural ignition.) By removing vegetation, the effects of ﬁre are to raise insolation at ground level, and to reduce albedo. Fire therefore increases

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the thickness of the active layer during summer thaws, which in turn stimulates nutrient cycling and raises primary productivity (Heinselman, 1981; Wein and MacLean, 1983). The magnitude of response to ﬁre is determined by the amount of vegetation and organic matter removed, and its timing. Fires during early summer, when organic layers are cold and wet, may have no effect on the active layer, but when ﬁres consume large amounts of vegetation and organic matter, the thickness of the active layer may increase signiﬁcantly (Brown, 1983). Black and Bliss (1978) reported that a severe ﬁre in forest-tundra in the Northwest Territories of Canada led to a thickening of the active layer, to gully erosion on bare hillsides, and to thermokarst activity on exposed permafrost. Release of meltwater created a new drainage network and caused widespread subsidence. Despite considerable alteration of physical conditions, however, most of the area was covered with herbaceous vegetation after two years, woody species were colonizing the site a few years later, and a layer of organic matter was accumulating rapidly on the previously ash-covered surface. Though no measurements were made, productivity appeared to be relatively high a few years after the burn, an effect attributed to nutrient release and higher surface temperatures (Black and Bliss, 1978).

HETEROTROPHIC ORGANISMS

Heterotrophic organisms inﬂuence primary production in a variety of ways. Particularly signiﬁcant is the consumption of tissue or solutes (herbivory), which probably affects all plants at some time during their life (Crawley, 1983). Infections caused by pathogenic organisms weaken host plants, and in some situations may play an important role in community dynamics (Mueller-Dombois, 1986). Physical disruption of plants is exempliﬁed by wood-boring insects. Soil physicochemical processes are affected by burrowing rodents. Beavers (Castor canadensis), by altering local hydrological budgets, can effect major changes in the landscape over relatively short time-scales (Pollock et al., 1995). Both the intensity of heterotrophic activity and the response of primary producers can be considered as continua which, theoretically, are amenable to quantiﬁcation. Strictly, therefore, one cannot identify the point of separation between a “normal” level of

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heterotroph activity (or severity of impact) and one which constitutes a “disturbance” [see Schowalter and Lowman (Chapter 9, this volume) for further discussion with respect to forest herbivory by insects.] Moreover, unlike ﬁre, a biological disturbance event typically extends over more than one year. Plant consumption Grazing instantly reduces the amount of plant tissue used in gathering light, water or nutrients, and the plant’s reserves of energy and nutrients. Not surprisingly, above certain thresholds of grazing intensity, primary productivity declines. However, it is also abundantly clear that primary productivity is not simply a negatively linear function of either grazing intensity (number of grazers, duration of grazing) or severity (amount of tissue lost). Rather, there are a variety of responses, many of which indicate that plants should not be regarded simply as passive participants in the grazing process (McNaughton, 1979; Crawley, 1983; Belsky, 1986). Complicating any attempt to generalize the relationship between grazing and primary productivity is the fact that responses to grazing have been studied at the level of the individual plant, the population, and the community (Crawley, 1983). While individual species may be adversely affected by grazing, community-level productivity may be stimulated by lowered competitive pressure on unaffected plants, and the release of resources. Another difﬁculty is that mechanical defoliation, which is often used as a surrogate for grazing, may elicit responses different from those to grazing (Jameson, 1963). In evaluating the research literature, it is therefore vital to establish the conditions under which observations have been made. Compensatory growth In many grazing studies, the removal of plant tissue has elicited a positive growth response in the grazed plants. The term “compensation” is now commonly used when the rate of loss of plant weight due to grazing is less than the rate of consumption by the feeding organism (Crawley, 1983). The phenomenon of compensatory growth is clearly a key issue in any discussion about the relationship between herbivory and primary productivity. Belsky (1986) suggested that the epithets “over”,

Ian K. BRADBURY

“exact”, and “under” should be used when compensation growth of grazed plants is, respectively, greater than, equal to, or less than, that of ungrazed controls. Undercompensation was further categorized as “partial” (1–99% of ungrazed controls), “nil” (zero compensation) and “damage” (< zero). Belsky (1986) also advocated the use of these terms in communities where grazed and ungrazed plants coexist. In such a situation grazed plants could release resources for use by ungrazed plants, and thereby raise productivity above that of the ungrazed community. McNaughton (1979) suggested a number of possible mechanisms, both “internal” and “external”, by which grazing could stimulate plant growth. Factors internal to the plant were enhanced photosynthetic rates of residual tissues, reallocation of assimilates, reduction of leaf area index to nearer the optimum level, increased survival of photosynthetic tissue, and the redistribution of hormones. External factors proposed were conservation of soil moisture following a reduction in leaf area, increased light penetration, more rapid nutrient cycling from dung and urine and, for grazing by ruminants, the growth-enhancing effects of saliva. Although the efﬁcacy of some of these mechanisms has been challenged (Belsky, 1986), this scheme provides a useful framework for considering mechanisms by which plants may respond to grazing. While there is a consensus that plants should not be regarded simply as passive participants in the grazing process, there is much less agreement concerning the contribution of grazing to plant ﬁtness, and the extent to which grazing may be considered beneﬁcial to plants (McNaughton, 1983; Belsky, 1986). In a review of grazing in western North America, Lacey and Van Poolen (1981) reported that herbage production on moderately-grazed range exceeded that of ungrazed range by an average of 68%. In the Serengeti grasslands of Tanzania, McNaughton (1979) found that moderate grazing doubled above-ground primary production, while severe grazing did not depress production below controls. The exclusion of grazers resulted in a loss of low-growing grasses which were dominant under grazed conditions, an effect repeated for individual species in controlled tissuereduction experiments. Grazing history affected the response to defoliation of Panicum coloratum (Poaceae) from the Serengeti (Dyer et al., 1991). Plants from lightly and heavily grazed locations exhibited similar photosynthetic rates

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS

before a grazing treatment was applied using grasshoppers. After grazing, plants from the heavily-grazed location revealed full compensatory growth, while the biotype from the rarely-grazed location yielded 20% less than controls. Grazing studies of the lesser snow goose (Anser caerulescens) in graminoid-dominated swards around Hudson Bay, Canada, revealed differential responses according to plant species and site conditions. In general, productivity was either unaffected or stimulated by grazing. For the grass Puccinellia phryganodes, the removal of faecal deposits prevented compensatory growth. As compensatory growth in this species is associated with enhanced foliar nitrogen, it appears that the geese stimulate the nitrogen cycle. For other plant species, however, such an effect was not demonstrated, despite evidence of compensatory growth (Hik and Jefferies, 1990; Zellmer et al., 1993). In communities subjected to different grazing intensities in semiarid Rajasthan (India), the highest productivity was associated with intermediate grazing pressure, although there were edaphic and ﬂoristic differences between sites (Kumar and Joshi, 1972). Vinton and Hartnett (1992) studied the effects of grazing by bison (Bison bison) on primary production of tallgrass prairie in Kansas (U.S.A.). While compensatory growth resulted in similar amounts of above-ground biomass in different treatments in the year of grazing, in the following year growth rates in grazed swards were lower than in ungrazed swards. This study emphasizes the possible transitory nature of compensatory shoot growth, and its dependence on stored reserves. Mammalian herbivores typically consume a large proportion of above-ground tissue of the herb Ipomopsis aggregata in northern Arizona (U.S.A.) (Paige and Whitham, 1987). Here, the loss of apical dominance resulted in a doubling of plant biomass. Furthermore, seed production and seedling survival was about 2.4 times greater for grazed than ungrazed plants, suggesting a way in which grazing might contribute to ﬁtness. Edenius et al. (1993) suggested that compensatory growth should be less common in tree species than in shrubs and herbs, and less common in gymnosperms than angiosperms. However, they demonstrated compensatory growth, although not overcompensation, in Pinus sylvestris (Scots pine) following clipping of leading shoots to simulate winter grazing by moose

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(Alces alces). Hjalten et al. (1993) examined the interaction between competition and type of defoliation in young Betula (birch) trees. Overcompensation occurred in low-density trees from which the top 1 cm of the leading shoot had been removed. All other treatments, including defoliation, led to reduced performance. These authors suggested that more attention should be paid to identifying the circumstances under which a particular response is to be expected, rather than thinking simply in terms of positive and negative responses to grazing. Below-ground plant tissues are grazed by a variety of organisms, including insects and other arthropods, nematodes, molluscs, and small mammals, but grazing losses are notoriously difﬁcult to quantify. Very little is known about compensatory growth below ground, although Brown and Grange (1990) suggested that low levels of herbivory may encourage the production of relatively large root systems. Defoliating insects Among the best-studied plant–herbivore relationships are those between trees and phytophagous insects. The nature of these relationships is explored in detail by Schowalter and Lowman (Chapter 9, this volume). Here, remarks are conﬁned to the problem of quantifying the response of primary productivity to insect activity, and some ﬁeld observations. The impact of defoliating insects on primary productivity is not well quantiﬁed, and it is therefore difﬁcult to make meaningful comparisons between studies (Myers, 1988). A major obstacle is the choice of measure to best reﬂect tree production (Kulman, 1971). The inﬂuence of herbivory on diameter at breast height, for example, depends upon the distribution of defoliation within the tree crown. Measurements made with dendrometers are inﬂuenced by the degree of tissue hydrature. Also limiting the scope for generalization is the plethora of variables which may inﬂuence the response to defoliation. In addition to the amount of tissue lost, Kulman (1971) cited foliage age, foliage location, time of defoliation, and stage of development as contributory factors. Coniferous trees may respond to defoliation by heavy production of cones, thus diverting resources from wood production. Importantly, defoliation tends to increase the susceptibility to attack from other biological agents and, by lowering overall water content, to ﬁre. Kulman (1971), relying principally on circumstantial evidence, claimed that,

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when tree death is associated with defoliation, it is normally bark beetles, borers, and fungi which are the proximate causes. Myers (1988) listed eighteen species of North American and European Lepidoptera causing mass defoliation during more or less cyclic population eruptions which typically remain at their peak level for 1–3 yr. After heavy defoliation, a reduction in tree growth occurs, which is expressed in smaller, or possibly missing, growth rings. However, widespread tree mortality is unusual (Myers, 1988). For example, when deciduous, broad-leaved trees are heavily defoliated in early summer refoliation normally occurs later in the summer. Periodic eruptions of Choristoneura fumiferana (spruce budworm) populations in eastern Canada cause considerable tree mortality, particularly in overmature, even-aged stands dominated by Abies balsamea (balsam ﬁr). Kulman (1971) suggested that the growth increment of spruce and ﬁr in mixed stands in the decade following major defoliation by spruce budworm is only half that in the ten years prior to defoliation. While the decline in tree productivity is obvious in the short term, it is less easy to gauge the impact on productivity in the longer term because the rate of gain in carbon in old stands may be extremely low in the years prior to defoliation. Heterotrophic activity and site effects In some situations, the activities of animal species have a major impact on the physicochemical characteristics of a site. Such species, characterized by Lawton and Jones (1995) as “allogenic engineers”, are likely to affect primary productivity, by altering either physical state variables or resource availability. One of the most striking examples of animal activity affecting the physical environment is the beaver in North America. Although the range of this species, and the size of its populations, have been much reduced during the past two centuries (Pollock et al., 1995), beavers remain a potent inﬂuence on ecosystem dynamics in parts of the boreal-forest zone. By cutting standing timber to create dams, beavers increase sediment and organic matter in stream channels, and create a mosaic of wetland communities (Pollock et al., 1995). Where open water replaces essentially terrestrial habitats, a decline in primary productivity is to be expected, although the patterns of productivity of rooted vegetation in wetland sites is less predictable.

Ian K. BRADBURY

Naiman et al. (1994) studied the effects of beaver impoundments on the Kabetogama Peninsula (Minnesota, U.S.A.). Here, a large increase in the beaver population between the 1920s and the 1980s has led to a major conversion of forest to wetland and open water in less than ﬁfty years. Primary production, as determined by peak living biomass and litter, did not differ greatly between shallow ponds and meadows which experienced different degrees of waterlogging. However, the standing stock of a range of nutrients, including available nitrogen, increased markedly in all the communities formed as a result of impoundment. Naiman et al. (1994) postulated the following scenario. Impoundment results in a transfer of nutrients from living vegetation, either to the soil or sediment, or out of the system in drainage water. The nutrients which are stored, under predominantly anaerobic conditions, become readily available when aerobic conditions return following dam abandonment, and are responsible for the relatively high productivity of meadow communities. The effect of beaver activity on nutrient dynamics is likely to be particularly signiﬁcant in the boreal zone, where productivity is frequently limited by nutrient availability (Naiman et al., 1994). Burrowing herbivorous rodents occur on all continents except Australia, and are associated particularly with temperate grassland and open scrubland. It is difﬁcult to separate the direct effects on primary productivity of feeding by such animals from the indirect effects associated with burrowing activity. It is clear, however, that when large amounts of soil are excavated and redeposited by burrowing rodents a major redistribution of nutrients also occurs. Excluding Cynomys ludovicianus (prairie dogs) experimentally from plots of North American grassland had no effect on net annual primary production, but the yield of nitrogen in above-ground tissue was lower in the absence of these animals (Whicker and Detling, 1988). This result suggested that prairie dogs encourage nitrogen mineralization, an effect attributed either to higher temperatures on cleared areas, or to the excretions of the animals. The mounds created by burrowing activity of Geomys bursarius (pocket gopher) tend to have greater nutrient concentrations than surrounding areas, although exceptions have been reported. However, primary productivity tends to be highest in areas fringing the mounds (Huntly and Inouye, 1988). In steppelands of the former Soviet Union and Mongolia, it has been claimed that the physical effects of rodents on soils more than compensate

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS

for their consumption of plant tissue (Lavrenko and Karamysheva, 1993). Improved soil conditions were attributed to a combination of mechanical disturbance and the input of faeces and urine. Moreover, excluding small mammals had the effect of shortening the growing season by two weeks. In contrast, during intervals of population eruption, bunch-grass communities are converted to “bare desert” (Lavrenko and Karamysheva, 1993).

CONCLUSIONS

The range of ecological processes affected by different disturbance agents, together with non-monotonic relationships between disturbance intensity and vegetation response, make it unlikely that simple relationships between productivity and disturbance can be established. The identiﬁcation of patterns in these relationships is hampered by the problems inherent in measuring primary productivity accurately, and in comparing data from studies in which many variables differ. Most of the available evidence comes from studies of the effects of ﬁre or grazing on vegetation. In some communities, recovery from ﬁre is typically rapid, and productivity may be stimulated by enhanced photosynthetic capacity, increased resource availability, and altered microclimates. Fire has the potential to lower site productivity by causing nutrient loss during combustion, and predisposing nutrients to erosion after the burn. The combustion of organic matter may reduce or increase site productivity depending on circumstances. Where natural ﬁres are rare, particularly on sites of high erodibility, ﬁre may quickly lower site productivity. The relationships between grazing and primary productivity are complex, largely because plants tend to respond actively to tissue consumption. For individual plants and populations, the productivity response is inﬂuenced by growth stage, tissue type, and amount of tissue removed. At the community level, the relationship is inﬂuenced by competition and resource availability. While compensatory growth in response to grazing has been demonstrated in many studies, plant growth is ultimately constrained by energy and nutrients. Some animals, for example beavers and burrowing rodents, can have a major impact on physicochemical site conditions. When such activities alter microclimate, local hydrology, or nutrient distribution and

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availability, they have implications for primary productivity. It may be difﬁcult, however, to establish whether observed changes in productivity are due to physical alterations of the site or to tissue consumption. Overall, the literature covering disturbance effects on productivity is uneven in both quantity and quality. Particularly lacking are reliable quantitative data, reﬂecting the considerable problems involved in determining the magnitude of disturbance effects and accurately measuring primary productivity. The mechanisms which link disturbance and primary productivity are complex, and remain poorly understood. The fact that several processes are operating simultaneously during and after a disturbance event suggests that carefully designed and controlled experimentation will be required if greater understanding of the underlying mechanisms is to be achieved.

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Ian K. BRADBURY Kamnalrut, A. and Evenson, J.P., 1992. Monsoon grasslands in Thailand. In: S.P. Long, M.B. Jones and M.J. Roberts (Editors), Primary Productivity of Grass Ecosystems of the Tropics and Subtropics. Chapman and Hall, London, pp. 100–126. Kucera, C.L., 1992. Tall-grass prairie. In: R.T. Coupland (Editor), Natural Grasslands: Introduction and Western Hemisphere. Ecosystems of the World 8A. Elsevier, Amsterdam, pp. 227–288. Kulman, H.M., 1971. Effects of defoliation on growth and mortality of trees. Annu. Rev. Entomol., 16: 289–324. Kumar, A. and Joshi, M.C., 1972. The effects of grazing on the structure and productivity of the vegetation near Pilani, Rajasthan, India. J. Ecol., 60: 665–674. Lacey, J.R. and Van Poolen, H.W., 1981. Comparison of herbage production on moderately grazed and ungrazed Western ranges. J. Range Manage., 34: 210–212. Lambers, H. and Poorter, H., 1992. Inherent variation in growth rate between higher plants: a search for physiological causes and ecological consequences. Adv. Ecol. Res., 23: 187–261. Lavrenko, E.M. and Karamysheva, Z.V., 1993. Steppes of the former Soviet Union and Mongolia. In: R.T. Coupland (Editor), Natural Grasslands: Eastern Hemisphere and Resume. Ecosystems of the World 8B. Elsevier, Amsterdam, pp. 3–59. Lawton, J.H. and Jones, C.G., 1995. Linking species and ecosystems: organisms as ecosystem engineers. In: J.H. Lawton and C.G. Jones (Editors), Linking Species and Ecosystems. Chapman and Hall, New York, pp. 141–150. Le Hou´erou, H.N., 1981. Impact of man and his animals on Mediterranean vegetation. In: F. di Castri, D.W. Goodall and R.L. Specht (Editors), Mediterranean-Type Shrublands. Ecosystems of the World 11. Elsevier, Amsterdam, pp. 479–521. Long, S.P., Jones, M.B. and Roberts, M.J. (Editors), 1992. Primary Productivity of Grass Ecosystems of the Tropics and Sub-tropics. Chapman and Hall, London, 267 pp. MacLean, D.A., Woodley, S.J., Weber, M.G. and Wein, R.W., 1983. Fire and nutrient cycling. In: R.W. Wein and D.A. MacLean (Editors), The Role of Fire in Circumpolar Ecosystems. John Wiley, Chichester, England, pp. 111–132. McNaughton, S.J., 1979. Grazing as an optimization process: grass– ungulate relationships in the Serengeti. Am. Nat., 113: 691–703. McNaughton, S.J., 1983. Compensatory growth as a response to herbivory. Oikos, 40: 329–336. Miller, G.R., 1979. Quantity and quality of annual production of shoots and ﬂowers by Calluna vulgaris in north-east Scotland. J. Ecol., 67: 109–129. Mooney, H.A. and Gulman, S.L., 1983. The determinants of plant productivity: natural versus man-modiﬁed communities. In: H.A. Mooney and M. Godron (Editors), Disturbance and Ecosystems: Components of Response. Springer-Verlag, Berlin, pp. 146–158. Mueller-Dombois, D., 1986. Perspectives for an etiology of standlevel dieback. Annu. Rev. Ecol. System., 17: 221–243. Myers, J.H., 1988. Can a general hypothesis explain population cycles of forest Lepidoptera? Adv. Ecol. Res., 18: 179–242. Naiman, R.J., Pinay, G., Johnston, C.A. and Pastor, J., 1994. Beaver inﬂuences on the long term biogeochemical characteristics of boreal forest drainage networks. Ecology, 75: 905–921. Newman, E.I., 1988. Mycorrhizal links between plants: their functioning and ecological signiﬁcance. Adv. Ecol. Res., 18: 243–270. Paige, K.N. and Whitham, T.G., 1987. Overcompensation in response

DISTURBANCE AND PRIMARY PRODUCTION IN TERRESTRIAL ECOSYSTEMS to mammalian herbivory: the advantages of being eaten. Am. Nat., 129: 407–416. Pearson, H.A., Davis, J.R. and Schubert, G.H., 1972. Effects of wildlife on timber and forage production in Arizona. J. Range Manage., 25: 250–253. Peet, R.K., 1981. Changes in biomass and production during secondary forest succession. In: D.C. West, H.H. Shugart and D.B. Botkin (Editors), Forest Succession: Concepts and Applications. Springer-Verlag, New York, pp. 324–338. Peet, R.K., 1992. Community structure and ecosystem function. In: D.C. Glen-Lewin, R.K. Peet and T.T. Veblen (Editors), Plant Succession: Theory and Prediction. Chapman and Hall, London, pp. 103–151. Pollock, M.M., Naiman, R.J., Erickson, H.E., Johnston, C.A., Pastor, J. and Pinay, G., 1995. Beaver as engineers: inﬂuence on biotic and abiotic characteristics of drainage basins. In: C.G. Jones and J.H. Lawton (Editors), Linking Species and Ecosystems. Chapman and Hall, New York, pp. 117–126. Raison, R.J., 1979. Modiﬁcation of the soil environment by vegetation ﬁres, with particular reference to nitrogen transformations: a review. Plant Soil, 51: 73–108. Rundel, P.W., 1981. Fire as an ecological factor. In: O.L. Lange, P.S. Nobel, C.B. Osmond and H. Ziegler (Editors), Encyclopedia of Plant Physiology, Vol. 12A, Physiological Plant Ecology 1: Responses to the Physical Environment. Springer-Verlag, Berlin, pp. 500–538. Rundel, P.W., 1983. Impact of ﬁre on nutrient cycles in Mediterranean-type ecosystems with reference to chaparral. In: F.J. Kruger, D.T. Mitchell and J.U.M. Jarvis (Editors), Mediterranean Type Ecosystems: The Role of Nutrients. SpringerVerlag, Berlin, pp. 192–207. Sanford, W.W. and Isichei, A.O., 1986. Savanna. In: G.W. Lawson (Editor), Plant Ecology in West Africa: Systems and Processes. John Wiley, Chichester, pp. 95–149. Sousa, W.P., 1984. The role of disturbance in natural communities. Annu. Rev. Ecol. System., 15: 353–391. Specht, R.L., 1966. The growth and distribution of mallee– broombrush (Eucalyptus incrassata/Melaleuca uncinata association) and heath vegetation near Dark Island Soak, Ninety-Mile Plain, South Australia. Aust. J. Bot., 14: 361–371. Specht, R.L., 1969. A comparison of the sclerophyllous vegetation characteristics of Mediterranean type climates in France, California and South Australia. II. Dry matter, energy and nutrient accumulation. Aust. J. Bot., 17: 293–308. Sprugel, D.G., 1985. Natural disturbance and ecosystem energetics. In: S.T.A. Pickett and P.S. White (Editors), Natural Disturbance and Patch Dynamics. Academic Press, Orlando, pp. 335–352. Suijdendorp, H., 1981. Responses of the hummock grasslands of

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Chapter 25

PATTERNS AND PROCESSES IN PRIMARY SUCCESSION Lawrence R. WALKER

INTRODUCTION

Humans have been fascinated by changes in the biotic environment around them for centuries, and ecologists have documented these changes for over 100 years. Many explanations have been proposed to explain species change, but no general theory has emerged that encompasses the complexities of species replacements across a wide range of habitats. Some argue such a search is futile (Miles, 1987; van der Maarel, 1988), yet much insight is gained by comparing species change in various habitats and seeking general explanations for such a fundamental process (Pickett et al., 1987a,b; Walker and Chapin, 1987; Pickett and McDonnell, 1989; Glenn-Lewin et al., 1992; Bazzaz, 1996). There is also an urgent practical need to understand how to manipulate and perhaps accelerate species change for the restoration of damaged communities (Luken, 1990; Marrs and Bradshaw, 1993), whether or not sitespeciﬁc lessons are generalizable. Disturbance can be deﬁned as the destruction of biomass by physical or biological processes, and contrasts with stress, which can be deﬁned as any existing environmental condition that limits photosynthesis or productivity (Grime, 1979; White, 1979; Sousa, 1984). A disturbance can be examined by what causes it (e.g., a landslide), the mechanism by which the damage occurs (e.g., erosion of soil), or the effect it has on the biota (loss of vegetation cover; opportunity for new colonization events) (Glenn-Lewin and van der Maarel, 1992). Disturbance generally initiates species change (but see Grubb, 1988) and can continue to inﬂuence it through time. Autogenic processes or disturbances promote species change, and originate from biotic interactions and biotic alterations of the physical environment (e.g., herbivory). Allogenic processes originate from the physical environment

(e.g., wind). Although autogenesis is normally from within a community and allogenesis from outside, the distinction is not always clear (as in the displacement of a vegetation community by a plant invader, or the disruption of soils by animals: cf. Burrows, 1990). Succession in vegetation can be considered as a sequential change in species composition (or other ecosystem characteristics) over time spans ranging from about 2 to 2000 yr. Insect or microbial succession may occur rapidly (<10 yr: Goddard and Lago, 1985; Hanski, 1987) because the organisms involved have relatively short life cycles. Changes in vegetation that take ~10–100 years are referred to as ﬂuctuations, cyclic succession, gap or patch dynamics (Pickett and White, 1985; Burrows, 1990; Glenn-Lewin and van der Maarel, 1992; Callaway and Davis, 1993). Longer-term changes (>100 yr), often due to widespread climatic shifts, are termed secular succession or geohistorical changes (Burrows, 1990; Glenn-Lewin and van der Maarel, 1992). Some authors argue that the term “succession” is now too loaded with connotations of directionality (to a climax) or convergence of multiple pathways, and should be discarded (Burrows, 1990). If not deﬁned either too broadly (Gleason, 1926) or too narrowly (Clements, 1916), succession is still a functional and widely recognized term. Primary succession can be deﬁned as species replacement on substrates with no prior soil development. Changes in other community and ecosystem attributes (e.g., productivity, soil nutrients) will be considered in this chapter only as they are associated with ﬂoristic change (cf. Miles, 1987; Glenn-Lewin and van der Maarel, 1992). This deﬁnition generally implies succession on a newly exposed, often dry, and nearly sterile substrate with little or no organic matter present,

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low nutrient status, and no residual biological legacy (e.g., seeds or other propagules). Secondary succession is species replacement where a biological legacy is present. However, a continuum exists between clearly primary seres (successional sequences) (e.g., newly-exposed rock from mining or lava ﬂows) through seres with some nutrients, organic matter and propagules (e.g., a ﬂoodplain with seeds and sediments from upstream communities), to clearly secondary seres (forest regeneration on abandoned pastureland) (Vitousek and Walker, 1987). Primary succession is the biological response to an extreme allogenic disturbance. In contrast, secondary succession may occur following allogenic disturbance (e.g., ﬁre) or autogenic disturbance (e.g., herbivory). Because primary succession embodies the complete reconstruction of biotic communities, many patterns and processes differ between primary and secondary succession. For example, primary seres may be less predictable, with more potential pathways, in part because of the important yet unpredictable role of dispersal of propagules to the site. Similarly, nitrogen-ﬁxing species may be more important in soil development and facilitation of the growth of other species during primary succession (Vitousek and Walker, 1987). Both primary and secondary succession can occur on land or in the water, but only primary xerarch (dry land) succession will be considered in this chapter. For discussions of examples of primary hydrarch (wet land) succession one may refer to Sousa (1979) or Burrows (1990). Along the continuum from primary to secondary seres, there is a wide diversity of types of primary seres. This chapter will examine that diversity and explore similarities among primary seres. Types of disturbance regimes, interactions of disturbances, and the spatial and temporal variability of disturbances will be discussed, both within and among types of primary succession. Plant and animal responses to disturbance, and their role in changing the physical environment (e.g., soil development, light attenuation) will also be compared. Finally, the processes that drive primary succession will be analyzed in the context of general theories about succession. Natural primary successions do not cover a large proportion of the earth’s surface, yet rapidly increasing habitat destruction by humans makes it imperative that one tries to understand how plant and animal communities recolonize severely disturbed sites.

Lawrence R. WALKER PATTERNS OF DISTURBANCE

Clements (1928) described succession as the net effect of six basic processes: nudation, migration, ecesis (or establishment), competition, reaction (or site modiﬁcation by organisms) and stabilization. All of these processes interact during succession, although stabilization can be considered a net effect of the other ﬁve (Pickett et al., 1987b). Nudation refers to the creation of bare terrain, and will be considered in this section. The other four processes will be considered in later sections. In this section I discuss the general characteristics of the major types of disturbances that initiate primary succession, and compare the spatial and temporal variability within each disturbance type. Spatial variability includes the scale or magnitude of the disturbance, the severity of the disturbance, including the nature of the substrate, and the patchiness or proportion of the system that is disturbed. Figure 25.1 compares the spatial scale (in m2 ) to the severity (in % of original biological content remaining) of the eight major types of disturbance. The relatively large biological legacy of landslides and ﬂoodplains, compared to glacial moraines, mine wastes, and volcanoes, may be noted. Temporal variability includes the past history of the disturbance, its current frequency (return time and predictability), changes in its severity over time, and interactions of disturbance regimes of different types. Figure 25.2 indicates causal relationships among the disturbances, and estimated ranges of return intervals for each disturbance type. Disturbances on the top row are ones that cause other disturbances. Disturbances on the bottom row may or may not be consequences of other disturbance types. For example, a landslide may ultimately be caused by destabilization from volcanoes, dunes, glacial activity, mining, or road construction. However, landslides can be triggered simply by rainfall on steep slopes. Return-interval estimates vary widely. Anthropogenic disturbances such as mining and transportation effects (e.g., road construction) have generally lower upper limits than natural disturbances. Most disturbance types have potentially short return intervals. Despite the possibility of renewed volcanic eruptions or surging glaciers, recurrence of these two disturbance types is least likely. The spatial and temporal variability of eight disturbance types are discussed in approximate order from least to most similar to secondary succession (least to most biological legacy).

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Fig. 25.1. The spatial scale (logarithmic) and severity (% of original biological legacy remaining) are compared for eight types of disturbance that initiate primary succession. VO, volcanoes; DU, dunes; GL, glacial forelands; MI, mine wastes; TR, transportation; FL, ﬂoodplains; LA, landslides; RO, rock outcrops. MI and TR are purely anthropogenic disturbances.

Fig. 25.2. Interactions among eight types of disturbance that initiate primary succession. Disturbances in the upper row can cause other disturbances in the lower row, as shown by the lines linking them. Horizontal arrows also point to a disturbance that can be caused by another disturbance. First-row disturbances are generally more severe than those on the second row. Estimated ranges of return intervals in years are noted for each disturbance. Abbreviations are as in Fig. 25.1.

Volcanoes Volcanic eruptions (and their associated tectonic counterparts, earthquakes) are among the most destructive of natural disturbances, and they occur on most continents and many islands. Their effects can be felt far beyond the immediate vicinity of the volcano. For example, ash “veils” from Krakatoa (Indonesia, 1883), Katmai (Alaska, 1912), Agung (Bali, 1963), and Chich´on (Mexico, 1982) reduced solar radiation and lowered temperatures throughout the world (Simkin and Fiske, 1983). The various types of volcanic ejecta

differ widely in form, texture, and chemistry, and therefore in their immediate and long-term effects on colonizing biota. Volcanic substrates are generally devoid of nitrogen, but with respect to bases (calcium, magnesium, potassium, sodium) may be either rich (basalt) or poor (rhyolite). Toxic ﬂuorine and sulfuric acid may inhibit colonization of volcanic substrates (Burrows, 1990). Lava (termed a’a or pahoehoe depending on the amount of dissolved gases) completely destroys any biota, and reshapes the landscape, often creating new land, as in the recent expansion of the island of Hawaii by the eruption of the volcano Kilauea (Monastersky, 1995). The smooth pahoehoe lava may be more rapidly colonized than the porous, crumbly a’a because water collects in cracks in the pahoehoe (Atkinson, 1970). Microclimatic inﬂuences strongly inﬂuence succession, even on the same lava ﬂow (Aplet and Vitousek, 1994). Explosive eruptions produce slag, pumice, cinder, and ash, which may or may not ﬂow downslope. As noted, ﬁner particles may reach the upper atmosphere and affect the entire earth. Locally, the severity of the disturbance depends on the depth of deposition, most vegetation not surviving ash depths greater than 70 cm (Brown et al. 1917; Eggler, 1948; Zobel and Antos, 1992; see also del Moral and Grishin, Chapter 5, this volume). Ash depths of about 30 cm promoted plant growth in Alaska (Griggs, 1933). Deep cinder or ash deposits may be very slowly colonized due to rapid water percolation or instability. In general,

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survival is a function of depth of deposition and distance from the eruption (del Moral and Grishin, Chapter 5, this volume). In addition, patchiness of the disturbance increases with distance from the volcano, undisturbed islands or kipukas becoming more frequent at a distance. These undisturbed patches serve as sources of propagules for recolonization. Volcanic eruptions have been an inﬂuential part of geological and human history (del Moral and Grishin, Chapter 5, this volume) and there are currently active volcanoes in many parts of the globe, including New Zealand, Indonesia, Japan, Alaska, Hawaii, the Caribbean, and Iceland. Volcanologists are becoming increasingly proﬁcient at predicting eruptions on the basis of inﬂation of belowground magma reservoirs. Volcanoes typically erupt only periodically (Mount St. Helens, Washington, U.S.A.), but Kilauea (Hawaii, U.S.A.) has been erupting continuously for 12 years (Monastersky, 1995). Over geological time, volcanic activity depends on a continuous source of lava and the movement of crustal plates (Walker, 1990). Deprived of a lava source, volcanoes eventually erode. Local and landscape-level site changes occur during volcanic activity and following its cessation, through earthquakes, erosion, subsidence, landslides, and biotic activities. Enormous landslides have punctuated the formation of the Hawaiian Islands (Walker, 1990). One interesting and current interaction of disturbance types is in Iceland, where a volcano erupted through a glacier, causing extensive glacial melting and ﬂooding (Monastersky, 1996). Mine wastes Mine wastes incorporate a large variety of substrates and can include the abandoned pit, quarry, or mine, as well as wastes, whether processed (e.g., crushed and cyanide-leached gold-bearing ore) or unprocessed (e.g., non-ore bearing rock). The size of mine-waste accumulations has grown with improved extraction technology, from local quarries to open pits 3 km wide and thousands of hectares that are altered by strip mining. Extractive industries are common throughout the world (Cooke, Chapter 14, this volume) and mine wastes cover a signiﬁcant portion of the coterminous United States (about 0.2% or 1.5 million hectares as of 1971: Iverson and Wali, 1992). Marrs and Bradshaw (1993) have listed 12 different types of waste from extraction of such products as ore, oil, limestone, sand, or clay. Mine wastes constitute some of the

Lawrence R. WALKER

worst possible substrates for plant colonization. In addition to a complete absence of organic matter, mine wastes are typically very acidic or basic, depending on their source. Nitrogen is generally very low; and phosphorus, if present, is immobilized by the extremes in pH. Aluminum, copper, iron, manganese, and other elements may be present in amounts toxic to plant growth (Leopold and Wali, 1992). Compaction, salinity, and severe drought conditions can also reduce the rate of recolonization (Bradshaw and Chadwick, 1980; Iverson and Wali, 1992; Marrs and Bradshaw, 1993). However, some mine wastes may present more favorable growth conditions than surrounding unmined substrates (Bussler et al., 1984). Mine wastes may be interspersed with undamaged areas (e.g., hilltops not covered by wastes) that could serve as recolonization nuclei (Cooke, Chapter 14, this volume); but mine wastes are generally continuous, not patchy. Glacial forelands Glacial forelands, areas denuded by the advance of glaciers and exposed after glacial retreat, may represent as broad a variety of habitats as volcanic substrates. Glaciers are found in mountainous terrain at all latitudes, and across a wide range of elevations, including at high elevations in tropical mountains (Matthews, 1992). Although glacial forelands can be many kilometers wide, forelands tend to be long and relatively narrow, surrounded by non-glaciated surfaces. Generally considered a severe disturbance, glaciers may actually leave soil or organic matter (including erect tree trunks) in situ when they retreat (Porter, 1989; Matthews, 1992). Other substrates include eroded bedrock surfaces, and outwash (alluvial deposits from glacially-fed streams) and glacial till (unconsolidated, unsorted material) in deposits called moraines. Each of these has a characteristic texture and consequent water-holding capacity, although organic matter and nutrients are generally minimal (Burrows, 1990; Matthews, 1992; Chapin et al., 1994). During the last 500 000 years a large portion of the earth’s land masses have been sculpted by glaciers. Alternating glacial and interglacial periods have certainly inﬂuenced the evolution of the biota (Pielou, 1979). Since the last glacial retreat (beginning 18 000 BP), global temperature shifts have caused local retreats (peaking 7000 BP) and advances (most recently 400–100 BP) of glaciers (Porter, 1989). Recent global warming may increase melting worldwide, particularly

PATTERNS AND PROCESSES IN PRIMARY SUCCESSION

at higher latitudes (Heusser, 1956; Matthews, 1992), although the geometry of valleys is also a factor in determining the extent of deglaciation (Porter, 1989). Local site conditions on a glacial foreland are subject to many inﬂuences after the initial melting of the ice. Alluvial erosion may continue to sculpt the foreland, at least until extensive vegetative cover is achieved. Earthquakes and landslides may occur from isostatic rebound after the weight of the ice is removed. Forelands at high elevations are subject to frost heaving, soliﬂuction, and other destabilizing forces (Matthews, 1992). Geologically unstable forelands may even experience volcanic activity (Monastersky, 1996). As with other disturbance types, the removal of the initial disturbance is just the beginning. Vegetation changes then occur against a backdrop of other allogenic disturbances. Unconsolidated dunes Dunes, like volcanoes and glacial forelands, are widely distributed across the earth’s surface (Doing, 1985). Inland dunes are common features of deserts, discussed in Volume 12 of this series, and coastal dunes (see Volume 2 of this series) are found on the shorelines of oceans, lakes, and some large river systems. Active dunes typically move 3–15 m yr−1 and create a patchy mosaic of varying sand depths (Chadwick and Dalke, 1965; Belsky and Amundson, 1986; McLachlan et al., 1987). Dune-forming sands may contain variable amounts of organic material and nutrients, but water availability and/or substrate stability are generally considered the most important factors in colonization (Ayyad, 1973; Holton and Johnson, 1979; van der Laan, 1979; Danin, 1991). Inland desert dunes may represent islands of relatively mesic habitat where substrate stability rather than water is the principal limiting factor (Pavlik, 1985; Rundel and Gibson, 1996). On coastal dunes, water availability and substrate stability may increase as particle size and spray decrease going inland from marine (Barbour et al., 1985; Ishikawa et al., 1995) or lake (Lichter, 1998) shores. These zonal abiotic patterns may affect plant colonization (Kutiel and Danin, 1987). Historically, large areas of the world were covered with dunes, so dune-adapted species migrated more freely in the past than they do today (e.g., during the Holocene in the southwestern United States: Pavlik, 1989). Stabilization by plants, shifting climates, and wind conditions (Moreno-Casasola, 1986; Lancaster,

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1993) can alter dune formation and distribution. Human activities, including cutting of oak woodlands in Poland (Piotrowska, 1988), sand quarry activity in Siberia (Forbes, 1995), and dike building in the Netherlands (Olff et al., 1993), can also promote dune formation or expansion. Transportation Trails, roads, railroads, airports, power lines, buried pipelines, parking lots, and other disturbances related to transportation of people, goods, or services are rapidly altering urban, rural, and formerly pristine landscapes. Trails and unpaved roads are much better candidates for revegetation than asphalt and concrete. Roads created by bulldozing can have rates of recovery different from those of non-bulldozed tracks (L. Walker, pers. obs.). Compaction can be one of the more difﬁcult conditions for potential colonizers (Webb, 1983; Webb and Wilshire, 1983). Recolonization will also depend on the climatic regime; where climates are arid (L. Walker, pers. obs.) or cold (Roxburgh et al., 1988; Forbes, 1995; Auerbach et al., 1997), road surfaces are recolonized more slowly than in warmer, more humid climates (Sader, 1995), although a poor supply of soil nutrients may limit colonization even in humid climates (Young, 1994). Dust, lead from burning gasoline, or asbestos from wear of tires, are additional hazards along transportation corridors (Barrow, 1991; Forbes, 1995). Patches of semi-natural vegetation can nevertheless survive in such oases as freeway median strips, and certain roadside weeds may actually ﬂourish where competition is reduced. Disturbance by transportation activities has clearly increased as human populations and associated technology grow (Spellerberg, 1998). Site amelioration (by ripping old road beds, for example) is now practiced in some areas where roads are abandoned and revegetation is desired. Roads can destabilize slopes, causing landslides. More than half of a set of 173 landslides formed between 1964 and 1989 in eastern Puerto Rico were road-related (Guariguata and Larsen, 1990). Another study of 1859 landslides in eastern Puerto Rico found that road-related landslides were larger than ones not near roads (Larsen, 1995). Rock outcrops Rock outcrops may result from natural disturbances such as landslides, or from anthropogenic activities

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such as mining or construction. Long-stabilized rock outcrops can also be considered simply as a part of the abiotic environment. Natural rock outcrops occur most commonly in alpine, polar, and boreal regions where severe climatic conditions, and perhaps the paucity of propagules, keep them relatively free of vegetation (Rydin and Borgegard, 1988). Rock outcrops in lowland temperate or tropical climates may remain bare because of steep, unstable slopes, or constant erosion (shorelines, cliff faces), and/or from toxic conditions resembling mine wastes. The steep, upper slopes of karst hills in subtropical Puerto Rico, for example, remain bare primarily from erosion induced by rapid wetting and drying (Monroe, 1976), but the calcareous substrate also limits potential colonists. Rock outcrops in temperate and tropical areas are generally discrete patches surrounded by more heavily-vegetated areas. Unique ﬂoras (Fernald, 1907; Billings, 1973; Wiser, 1998) capable of surviving extreme desiccation and shallow soils (Burbanck and Platt, 1964; Sharitz and McCormick, 1972) may be adapted to the sparselyvegetated rocks. Rock outcrops may represent remnant patches of recently deglaciated terrain that will be colonized within a few decades, or semi-permanent patches in a landscape that are dependent for their presence on continual erosion (Oosting and Anderson, 1939; Shure and Ragsdale, 1977; Houle, 1990). Landslides Landslides are a common type of soil erosion on any steep terrain. They may be triggered by rains, earthquakes, or volcanoes, and are most common where other disturbances have created unstable slopes (e.g., glacial melting, mine wastes, tree cutting, or road construction). More than 60% forest cover can help reduce erosion (Pimentel and Harvey, Chapter 4, this volume). Landslides from rains (e.g., in Puerto Rico: Walker et al., 1996) are typically localized, ranging in size from less than 100 m2 (landslips) to 25 000 m2 , and can affect 1–15% of the landscape in the course of a century (Garwood et al., 1979; Walker et al., 1996). Landslides associated with volcanoes (also termed lahars, debris ﬂows, or mudﬂows) can be much larger, affecting many square kilometers (Dickson and Crocker, 1953; Heath, 1967; Dale, 1986; Walker, 1990). Landslide surfaces are typically very heterogeneous disturbances, with an upper scar zone where bedrock is exposed and no biological legacy

Lawrence R. WALKER

remains, a middle chute zone with minimal organic remnants, and a lower deposition zone where much of the organic debris from the upper two zones is deposited and where islands of vegetation or viable rhizomatous species may remain intact (Lundgren, 1978; Adams and Sidle, 1987; Guariguata, 1990; del Moral, 1993; del Moral and Grishin, Chapter 5, this volume). Therefore, the deposition zone is a much more favorable habitat for plant colonization than the upper two zones (Flaccus, 1959; Mark et al., 1964; Walker et al., 1996). Landslide scars may remain for centuries (Garwood et al., 1979; Guariguata, 1990; Curtin, 1994) and may be susceptible to renewed erosion (White, 1979), particularly if the landslide started below the top of the slope (Larsen and Torres-S´anchez, 1995). Average recurrence intervals for rain-caused landslides in Puerto Rico may range from 3000 to 10 000 years (Brown et al., 1995; Scatena and Lugo, 1995). Floodplains Floodplains are found at all latitudes where large, slow rivers deposit sediments (e.g., on the inside of meanders or in deltas). They are narrow or non-existent in steep terrain where river ﬂow is more rapid or sediment loads are reduced. Floodplain sediments may contain substantial organic matter, nutrients, animals, and plants (segments capable of vegetative growth, seeds, spores) (Luken and Fonda, 1983). Soil texture, soil stability, and soil water-holding capacity may be more important determinants of colonization than nutrient availability (Fonda, 1974; Frye and Quinn, 1979; Johnson et al., 1985; Kalliola et al., 1991; Shaffer et al., 1992). Floodplains are very patchy, depending on the limiting width of the river basin, water and sediment volumes, ﬂooding history, and the stabilizing inﬂuences of vegetation (Mann et al., 1995; Friedman et al., 1996). Tectonic uplift, rainfall, and the variables noted above all make ﬂoodplains very ephemeral at both geological and ecological time scales. Floodplains are often the product of silty, glacially-fed rivers, and can also be the source of dune-building sands. The construction of dams has altered both the amount and timing of stream ﬂow, thereby altering vegetation dynamics on ﬂoodplains (e.g., Busch and Smith, 1995; Higgins et al., 1996). Humans also modify ﬂoodplain succession through water pollution, logging,

PATTERNS AND PROCESSES IN PRIMARY SUCCESSION

and thermal pollution (Muzika et al., 1987; Davey and Rothery, 1993). Other primary seres Other terrestrial examples of primary succession occur in alpine regions (Griggs, 1956) or polar regions (Gold and Bliss, 1995) where severe climatic conditions often restrict vegetation following disturbance by cryoturbation or soil erosion, and the seres may involve nonvascular plants exclusively (Lewis Smith, 1993; Vestal, 1993). Wind erosion can remove soils in arid lands to promote desertiﬁcation (Barrow, 1991), and is involved in the creation and destruction of soil mounds in dry lake beds, creating conditions suggesting primary succession (Vasek and Lund, 1980). Development of biotic communities in mudﬂats, whether naturally emerging (Shaffer et al., 1992) or artiﬁcial (Gray, 1993), as well as in marshes, reefs, peatlands, and shorelines, can also be considered primary succession (Gray, 1993), but will not be incorporated in this discussion of predominantly xerarch seres. Humans create a number of severe environments not already mentioned. Clear-cutting a forest or overgrazing a pasture can lead to erosion of the soil (Pimentel and Harvey, Chapter 4, this volume). Construction sites (McKendrick, 1987; Bishop and Chapin, 1989), and toxic domestic, agricultural, or industrial wastes can create severe environments (e.g., sludge, sawdust piles, pesticide spills, nuclear wastes: Barrow, 1991). Military activities, from tank movements to waste dumps, can create severe disturbances (Demarais et al., Chapter 15, this volume). Most surfaces of the urban environment (Sukopp and Starﬁnger, Chapter 16, this volume) are also potential sites for primary succession.

RESPONSES TO DISTURBANCE

In the following section, I compare and contrast plant, animal, and ecosystem responses to the major disturbances already discussed. Plant responses are organized around life-history characteristics (dispersal, establishment, growth, and longevity), followed by a discussion of animals in primary succession. Then I examine ecosystem responses in terms of soil biota, plant and soil nutrients, water balance, and light conditions.

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Plants Dispersal and initial establishment Dispersal or migration (Clements, 1916) of viable propagules to the site is an essential step in most primary seres. In some cases, soil organisms or viable propagules (seeds, spores, vegetative parts capable of sprouting) survive the disturbance (e.g., landslides, ﬂoodplains, or shallow ash deposition). Initial sterility is unlikely except on new lava, as soil bacteria are readily found in most environments, including newly exposed glacial forelands in Antarctica (Wynn-Williams, 1993) or subglacial soils in the Alps (Matthews, 1992). Widely dispersed spores (and windblown fragments) may account for the importance of cryptogams (algae, fungi, lichens, bryophytes, ferns) in many primary seres (Brock, 1973; Adams and Adams, 1982; Whittaker et al., 1989; Lewis Smith, 1993; Walker, 1994). Cryptogamic crusts and fungi are considered important as soil stabilizers on dunes (Danin, 1991), on glacial forelands (Worley, 1973), and on older desert soils (Rychert et al., 1978), perhaps providing a favorable microhabitat for germination of seeds of vascular plant species (Griggs, 1933; Carpenter et al., 1987; Danin, 1991; Walton, 1993; Chapin et al., 1994). However, cryptogams are by no means necessary precursors to vascular species [e.g., on rock outcrops (Winterringer and Vestal, 1956) or landslides (Veblen and Ashton, 1978)]. Vascular species present in the seed rain generally have small, light, winddispersed seeds, although large-seeded species can be important pioneers (e.g., oaks on quarry walls: Grubb, 1987). On a glacial foreland in New Zealand (Archer et al., 1973) and mine wastes in the U.S.A. (Leopold and Wali, 1992), animal dispersal was second in importance to wind dispersal; in Norway, wind and stream dispersal during the ice-free season were more important than dispersal by animals or by winter winds across ice ﬁelds (Ryvarden, 1971, 1975). Dispersal dynamics may strongly inﬂuence succession on isolated volcanic islands such as Krakatau (Whittaker, 1993) and Surtsey (Fridriksson, 1987), or large open areas such as Mount St. Helens (del Moral, 1993). Pioneer species can have a strong inﬂuence on subsequent succession (Cooper, 1923; Dale, 1986; Wood and del Moral, 1987) through the process termed nucleation (Yarranton and Morrison, 1974), in which pioneer species form favorable nuclei for colonization by the same or later species (Blundon et al., 1993). Long-distance dispersal is not as important on glacial

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forelands [Matthews (1992); but see Chapin (1993)], landslides (Halpern and Harmon, 1983; Frenzen et al., 1988; Walker and Neris, 1993), mine wastes [Larson (1984); McClanahan (1986); but see Ash et al. (1994)], or ﬂoodplains (Walker et al., 1986) where propagule sources are nearby. In such areas, clonal expansion of pioneers may be important (Leisman, 1957; Kimura, 1991; Prach and Pyˇsek, 1994; Walker, 1994). Many more propagules arrive at a site than successfully establish (Wood and Morris, 1990; Matthews, 1992). Grubb (1986, 1987) suggested that substrate particle size is critical in determining successful establishment in primary succession, and that the range of growth forms depends on the substrate quality. He noted that grasses invade silty and sandy areas, herbs invade gravels, trees invade rock crevices, and bryophytes invade exposed rock surfaces. Thus, certain species attributes are either directly (life span) or inversely (growth rates) related to particle size (trees and lichens are longer-lived, grow more slowly, and establish on larger particle sizes). Grubb (1986, 1987) further suggested that sites with low water and low nutrient availability are colonized by long-lived species. Grubb’s hypotheses (regarding substrate particle size and resource availability) certainly provide a useful framework that needs to be further examined in a variety of primary seres. Grubb (1987) noted several exceptions, but some of his interpretations rest on whether a species is a “principal pioneer” (one that produces most of the biomass). Tree invasion of ﬁnetextured soils on ﬂoodplains (Johnson et al., 1976; Nanson and Beach 1977; Walker et al., 1986; Kalliola et al., 1991) or landslides (Walker et al., 1996) is another possible exception. Trees may be successful in such unstable soils because of their ability to germinate rapidly and grow roots and shoots sufﬁciently long to resist erosion or deposition (cf. Nechaev, 1967). In addition, many other types of colonizers such as microbial crusts (Wynn-Williams, 1993), cryptogamic invasion of the interstitial spaces within rocks (Vestal, 1993), perennial ferns (including tree ferns; Beard, 1945; Walker, 1994; Walker and Aplet, 1994; Russell et al., 1998) and vines need to be considered. Human inﬂuences (e.g., ﬁre: Bond and van Wilgen, 1996) can also alter colonization sequences. Clearly, many local factors (e.g., soil nutrients, substrate stability, water availability, propagule pool, seed mass) and regional factors (e.g., disturbance histories, climatic regime) affect colonization success.

Lawrence R. WALKER

Growth and longevity Contrasting life-history characteristics of organisms can help to explain species change during primary succession. Assuming little or no interaction between species, the effects of arrival times, growth rates, and longevity can be sufﬁcient to explain the pattern of succession on ﬂoodplains (Walker et al., 1986) or glacial moraines (Chapin et al., 1994), but may be most important in secondary succession where abundant resources allow for higher growth rates (Grime, 1979; Walker and Chapin, 1987). Early colonizers of severe disturbances are often fast-growing and short-lived (r-selected) and are frequently replaced by better competitors that are slower-growing and longer-lived (K-selected) as resources increase. The particular balance of stress, disturbance, and competition will help determine the suite of successful colonizers at a particular site. A combination of too much stress with disturbance results in no species surviving (Grime, 1979; Johannesson, 1989). Early colonizers of stressful environments must cope with severe microclimates and nutrient deﬁciencies; colonizers of disturbed environments must accommodate ephemeral habitats and unstable substrates. Mechanisms that enable survival in stressful environments include slow growth that demands very few resources, resource accumulation and concentration, efﬁcient retranslocation, reduced transpiration, the evergreen habit, and extensive allocation of resources to belowground growth (Bazzaz, 1979; Grime, 1979; Chapin, 1980; Tilman, 1985; Chapin and Braatne, 1986). Long-lived species are typically found in severe environments (e.g., bristlecone pines on alpine ridges, or perennial desert shrubs that survive for thousands of years: Ferguson, 1968; Vasek, 1980), and limited species replacements may occur. Some primary seres (e.g., glacial moraines in polar environments) are colonized slowly by bryophytes (Lewis Smith, 1993). On a glacial moraine in Norway, Crouch (1993) found both dispersal dynamics and responses to environmental stresses affected the relative abundance of cryptogams and phanerogams. In contrast to highstress environments, mechanisms that enable survival in disturbed environments can include rapid growth, high allocation to shoots, and rhizomatous growth (Walker and Chapin, 1986; Chiba and Hirose, 1993). In such environments, the plants are usually short-lived. More mesic, nutrient-rich substrates (e.g., ﬂoodplains in temperate environments) present a third alternative; they are usually rapidly colonized by fast-growing trees

PATTERNS AND PROCESSES IN PRIMARY SUCCESSION

(Fonda, 1974; Frye and Quinn, 1979; Walker et al., 1986; Kalliola et al., 1991; Prach, 1994), in which aboveground allocation is emphasized. In such habitats, succession may be determined by relative growth rates and competition for light (Huston and Smith, 1987), and species replacements may happen too fast for potential longevities of each species to be important. Animals Colonization Animals are critical to the dispersal of plants in primary succession (Archer et al., 1973; Matthews, 1992; Carlquist, 1994). Dispersal is either by ectozoochory (adherence to the outside of the animal by burrs, hairs, or in attached mud) or endozoochory (from the consumption of fruits). Long-distance animal dispersal to volcanic islands most likely occurs via ectozoochory on birds (Carlquist, 1994), but giant tortoises may also be dispersers (Fenner, 1985). Dispersal to volcanic islands closer to mainland seed sources may be predominantly by endozoochory (Whittaker and Bush, 1993) or ectozoochory (Ball and Glucksman, 1975). On Mount St. Helens, a mainland volcano, gopher mounds increased both species survival under ash and subsequent species richness (Andersen and MacMahon, 1985). Human-dispersed seeds are common along roadsides and trails, but wind (dunes), gravity (landslides), and water (ﬂoodplains) are probably more important agents of dispersal on other primary seres. Animal dispersal per se is also important to primary succession. Aeolian dispersal of crickets, spiders, and other invertebrates establishes early, self-sustaining populations of heterotrophs on new lava (Howarth, 1987; Edwards, 1988; Crawford et al., 1995). These animal communities may increase nutrient levels in the ecosystem and therefore facilitate the establishment of later plant arrivals (Edwards and Sugg, 1993). Colonization by larger animals is usually dependent on the establishment of plant food sources [e.g., early successional shrubs favored by hares on ﬂoodplains (Bryant and Chapin, 1986), and by moose on glacial moraines (Dinneford, 1990) and ﬂoodplains (Wolff and Zasada, 1979) in Alaska]. Succession In addition to dispersing plant propagules, heterotrophic organisms inﬂuence plant succession in a number of critical ways (Majer, 1989). Mycorrhizal fungi and nitrogen-ﬁxing bacteria are critical to the

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successional development of plant communities (Janos, 1980; Sprent, 1993; Boerner et al., 1996). Earthworms, ants, mites, termites, nematodes, bacteria, and fungi are all essential to the formation of soil structure, the decomposition of plant litter, and subsequent nutrient cycling (Swift et al., 1979). Decomposers often follow a succession themselves. Groups of bacteria and fungi that are most mobile and decompose simple sugars precede those that are less mobile and decompose more resistant material (Begon et al., 1990). Decomposers of wood (Torres, 1994) or carrion (Goddard and Lago, 1985; Early and Goff, 1986) also frequently follow a succession, where motility, size of mouthparts, and diet determine the sequence of species (see Willig and McGinley, Chapter 28, this volume). However, Schoenly and Reid (1987) cautioned that the causes of heterotrophic succession are still poorly understood. Edwards (1988) recorded a general pattern of animal succession on volcanoes from wind-blown spiders and crickets that are scavengers and detritus feeders, followed by omnivores, then herbivores, and lastly predators and parasites. Animals also affect plant succession by pollination and herbivory (van der Maarel et al., 1985; Majer, 1989; Bryant and Chapin, 1986; Brown et al., 1988). Pollinators determine plant reproduction and thereby either the maintenance of a species in a successional stage or its entry into the seed bank. Herbivores alter plant productivity, ﬁtness, competitive balances between plant species, and susceptibility to disease, drought, or other environmental factors (Urbanek, 1989). Large herbivores can retard succession (Whelan, 1989), producing a grazing disclimax (Clements, 1928), or accelerate succession by preferentially eating (Bryant and Chapin, 1986) or otherwise damaging (Amman, 1977) early successional species and creating gaps favorable to later successional species (Whelan, 1989). Herbivores can also alter the composition of later successional communities (Clarkson and Clarkson, 1995). Plants, in turn, inﬂuence animal succession by ameliorating microsites or by providing food, as noted for beetle succession on mine wastes in the Czech Republic (Hejkal, 1985) and Wyoming, U.S.A. (Parmenter and MacMahon, 1987). Ecosystems Soils and nutrients Soil development is inﬂuenced by climate, parent

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material, topography, organisms, and time (Jenny, 1980) and is usually a rather slow process (decades to millenia) during primary succession. Without direct measures of such slow changes, chronosequences must be used, with the often inadequate assumption that older sites have gone through the same stages of soil development as younger sites (Pickett, 1989; Fastie, 1995). Further complications include pockets of surviving soils (particularly on landslides or ﬂoodplains but also on glacial forelands) and continual inputs of soil or sediments during succession. Plants affect soil development by stabilizing substrates, breaking up the substrate with roots, transferring nutrients from subsoils to the surface, hosting nitrogen-ﬁxing bacteria and mycorrhizal fungi, exuding proteins and other compounds from their roots, and providing litter for decomposition. Invasion by trees with nitrogen-ﬁxing symbionts can dramatically increase soil nitrogen, for instance on volcanic substrates (Vitousek and Walker, 1989), dunes (Witkowski, 1991) and ﬂoodplains (Van Cleve et al., 1971; Luken and Fonda, 1983; Walker, 1989). Plants may also reduce nutrient availability through nutrient immobilization by litter, perhaps altering successional sequences (Maheswaran and Gunatilleke, 1988; Schimel et al., 1996). Animals contribute through soil aeration, facilitating litter decomposition, and vertical transfer of litter and nutrients. Matthews (1992) noted that the ﬁrst four of Jenny’s factors vary spatially, complicating the interpretation of soil development on glacial forelands, and that spatial heterogeneity can occur from the microsite to the global scale. Nevertheless, some factors show repeatable trends through time. Soil organic matter, soil nitrogen, and cation exchange capacity generally increase during succession on glacial forelands; pH generally decreases (Messer, 1988; Matthews, 1992). Walker (1993) also found, in several types of primary seres, that soil nitrogen increased rapidly for 50–200 years, then reached an asymptote at 2000–5000 kg N ha−1 . The asymptote was reached within 100 years on mine wastes and ﬂoodplains, but took longer (1000 years) on dunes. Other variables (aluminum, iron, phosphorus, soil texture) do not show generalizable patterns on glacial forelands (Matthews, 1992), although several studies have reported a gradual decline in phosphorus in dune succession over many centuries in New Zealand (Walker and Syers, 1976) and Australia (Walker et al., 1980).

Lawrence R. WALKER

Water Changes in soil water and plant water status can affect the rate and direction of primary succession (Grubb, 1986; Wood and Morris, 1990). Clements (1916) classiﬁed primary seres by the moisture conditions of the site. Soil water availability depends on soil texture, microclimate, and soil organic matter. Both soil texture and microclimate (but not soil organic matter) may be more heterogeneous spatially than temporally. Comparisons among disturbance types are difﬁcult because of such heterogeneity. Surface ﬂoodplain soils can be wet, dry, or variable (Kalliola et al., 1991; Van Cleve et al., 1993). Mine wastes (Schramm, 1966), volcanic surfaces (Adams et al., 1986; Braatne and Chapin, 1986), and rock outcrops (Shure and Ragsdale, 1977; Uno and Collins, 1987) tend to be dry substrates, but impermeable surfaces that collect rainwater can create wet microsites. Dunes in humid coastal areas may receive surface moisture, whereas inland dunes may be perched over water tables (Barbour et al., 1985). Danin (1991) noted that a minimum of 20 mm of rain per year was needed to support plant life on dunes in the Mediterranean region. Plant water status during succession depends on physiological and morphological traits of the plant species (e.g., stomatal conductance, rooting depth) as well as environmental factors (e.g., soil water, accumulation of organic material, microclimate). Light The responses of plants to light depend on nutrient availability when water is not limiting (Fetcher et al., 1983). Tilman (1988) suggested that the high light/nutrient ratio in early succession favors certain species, which are gradually replaced by species adapted to decreasing light and increasing nutrient availability. The germination of early successional species is generally light-induced (Fenner, 1985), but light-inhibited germination by some dune colonists may ensure burial (wetter soils, less erosion) before germination (Barbour et al., 1985). Photoinhibition, or the reduction of potential carbon gain under high light conditions, may limit growth of colonizers on landslides (Fetcher et al., 1996) or in other primary seres. As canopies develop in primary succession, shading is a mechanism that can speed or slow the rate of succession. If the species creating the shade cannot regenerate under its own shade, but another, potentially taller species can, shading may promote succession. When the shade is so dense that no regeneration

PATTERNS AND PROCESSES IN PRIMARY SUCCESSION

occurs, succession may be delayed or halted until the species creating the shade dies (Niering et al., 1986; Vitousek and Walker, 1989). The latter case is termed “competitive inhibition” and is discussed below.

PROCESSES IN PRIMARY SUCCESSION

In this section I examine the basic processes that drive primary succession in the context of current theories and models of succession. Successional theory has developed to incorporate both primary and secondary succession, but the relative importance of different processes varies along the continuum of environmental severity between primary and secondary succession. For example, facilitation is generally considered more important in primary succession, competition in secondary succession. I explore the relative importance of plant life histories, competition, and facilitation, and recommend future avenues for research. Successional theory History Succession is a process that has been observed and incorporated into human activity (hunting, agriculture, land reclamation from wetlands) for centuries (Clements, 1928). Warming (1895) and Cowles (1901, 1911) were among the ﬁrst authors to formulate general concepts about vegetation dynamics, derived in both cases from their observations of primary succession on dunes. Clements (1916, 1928, 1936) developed a detailed theory about succession that is still very inﬂuential today. One of his ideas, now largely discredited, was that successional change in plant communities resembled the development of an organism. He also asserted that, within a given climate, succession converged to a stable end point called a climatic climax. Numerous qualiﬁcations (e.g., proclimax, disclimax, subclimax) were needed to adapt this broad concept to the local variations in soil, climate and biota, suggesting that it was not realistic (Tansley, 1935). Clements’ attempts at a general theory of succession have been widely criticized in recent years, but his elucidation of basic processes in succession still provides the framework for current successional theory. Often forgotten is his wealth of keen ﬁeld observations, including his early distinction between primary and secondary succession and his integration of disturbance and succession (Clements, 1916).

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An alternative view of succession was simultaneously developed by Gleason (1917, 1926), which refuted Clements’ organismic analogy and described succession as a more indeterminate process of vegetation change dependent on the properties of individual species for its often unpredictable trajectories. Gleason (1926) also drew on examples of primary succession (e.g., rock outcrops, dunes) to make his points, albeit much less extensively than Clements (1928). Several subsequent studies integrated the emphasis of Clements on process or time and the emphasis of Gleason on pattern or space. Whittaker (1953) proposed a climax-pattern hypothesis that recognized a diversity of climaxes that corresponded to environmental gradients. Interestingly, he used the analogy of a primary sere (a braided stream or ﬂoodplain) to emphasize the continuously changing pattern of the vegetation. Watt (1947) provided a more inclusive synthesis of pattern and process when he proposed that the landscape is a shifting mosaic of interacting patches, each at a unique stage of succession. Watt’s approach was a preview of modern landscape ecology (Naveh and Lieberman, 1984) and patch dynamics theory [Pickett and White (1985); see also Pickett et al., Chapter 32, this volume). Most recent literature on succession has continued to reﬂect the early dichotomy between Clements and Gleason, with neo-Clementsians emphasizing holism, ecosystem-level processes that change in predictable ways (Odum, 1969), and cybernetic analogies where the system limits outside input as it develops during succession (Margalef, 1968). Clements’ emphasis on autogenic change or site modiﬁcation by organisms (reaction), largely discounted for several decades, has been recently re-emphasized (Bertness and Callaway, 1994; Callaway, 1995). Neo-Gleasonians have continued to emphasize reductionism, allogenic disturbance, spatial variability, and the stochastic nature of colonization and species replacements. They still argue whether succession, as a directional process, even exists (Miles, 1979). Raup (1957) saw succession as a process of recovery from past disturbance, not progression toward a climax. Egler (1954) suggested an initial ﬂoristics model, where all species were present initially, but became sequentially more conspicuous, reﬂecting the variation in plant growth rates and size at maturity. Drury and Nisbet (1973) concurred with Egler, and emphasized change at the population rather than the community or ecosystem level. They noted the paucity of data

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on primary succession, but suggested that the role of soil development (or at least of the build-up of organic matter) in primary succession was overstated, especially given examples of rapid colonization of glacial forelands (Cooper, 1923) and dunes (Richards, 1952). Connell and Slatyer (1977) presented a set of three alternative models of succession that have been widely cited. They considered only changes that occur without allogenic disturbance. Their facilitation model resembles Clementsian ideas that early colonizers alter the environment in ways that are favorable for later colonizers, but not for their own reproduction. Connell and Slatyer suggested that evidence of this type of succession came from studies of glacial forelands in Alaska, U.S.A. (Glacier Bay: Crocker and Major, 1955), and sand dunes in Michigan, U.S.A. (Cowles, 1899; Olson, 1958), as well as from heterotrophic succession on decomposing logs and carcasses (Payne, 1965; see also Goddard and Lago, 1985; Early and Goff, 1986). Connell and Slatyer’s other two models suggest that early colonizers either inhibit (competition model) or have no effect on (tolerance model) later colonists. They used examples of marine or secondary successions to support these models. Although the models were intended simply to explain the net effect of early colonizers on later ones (positive, negative, or neutral: Connell et al., 1987) they are easily interpreted as attempts to explain whole seres. Thus, while many subsequent successional studies have attempted to explain their data with reference to one or more of the models, most seres comprise a mixture of positive, negative, and neutral interactions between species, or aspects of each model (Walker and Chapin, 1987; Pickett et al., 1987b; Burrows, 1990; Matthews, 1992; Callaway and Walker, 1997). Perhaps models of species interactions that specify successional type, stage, and details of disturbance regimes and resource availability can be more productively tested. Grime (1977, 1979) directly addressed the inﬂuence of disturbance on succession by expanding the familiar r–K continuum (of generalist colonizers or r-selected species adapted to disturbance, and specialists or K-selected species adapted to competition) to a threeway approach involving species adapted to disturbance, competition, and stress. He suggested succession generally begins with disturbance-adapted species, goes through competition-adapted species, and ends with stress tolerators. He did not address primary succession in his model, but noted that in low-productivity

Lawrence R. WALKER

sites the dominance of competitive species would be reduced. Menges and Waller (1983) noted that certain disturbances (e.g., a ﬂood or a ﬁre) can be regarded by r-selected species as a disturbance, but as a stress by long-lived species that survive the perturbation in situ. They suggested that Grime’s threeway approach reduces to two axes for ﬂoodplain herbs – ﬂooding frequency and physiological or morphological adaptation to survival of ﬂoods. Thus, competitive species do well with low ﬂooding frequency, whereas both stress tolerators and r-selected species do well with high ﬂooding frequency, but adapt to ﬂooding differently. Noble and Slatyer (1980) proposed a series of vital attributes that, when understood for each species, could be used to predict succession. The vital attributes included the time to reach critical life-history stages (e.g., reproduction), the requirements for establishment, and the method of recovery from disturbance. They used only data from secondary succession (after ﬁre) to test their model, and it may not be easily applicable to primary seres (Matthews, 1992). Tilman’s (1985, 1988) resource-ratio model of succession proposed that the balance between nutrient and light availability drives species change in succession. He also used glacial-foreland succession at Glacier Bay, Alaska, to support the application of his model to primary succession, but with no direct evidence of how resource availability alters the competitive balance between species at Glacier Bay (cf. Matthews, 1992). In Tilman’s model, decreases in light and increases in nutrients drive succession. This generalization may best apply to primary seres that begin in open areas with low nutrients, but many seres, primary or secondary, do not follow this simpliﬁed pattern. Long-term effects of the pattern of initial colonization or environmental factors other than nutrients (e.g., salinity or soil moisture) can also affect succession (e.g., Olff et al., 1993). Further, a competitive equilibrium between species (at a given light/nutrient ratio) can only occur if species change is faster than the change in resource availability (Tilman, 1985; Huston and Smith, 1987). Rapid, allogenically driven changes in primary seres may not meet this condition. In 1987, three papers addressed the limitations of the Connell and Slatyer (1977) models. Only two (Huston and Smith, 1987; Walker and Chapin, 1987) dealt explicitly with primary succession. Huston and Smith (1987) noted that facilitation, competition, and tolerance are relative terms that are not mutually

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exclusive, and proposed a theory of succession based on individuals rather than populations. They suggested that individual plants respond uniquely to competition (e.g., for light) in ways that change throughout their life span. Because plants alter their physical environment, and because no one species or individual of a species can be a superior competitor in all habitats, successional change occurs. They modelled a primary succession where a nitrogen ﬁxer is competitively superior initially, but as light decreases and nutrients increase, other species become dominant. Pickett et al. (1987a,b) furthered the argument that mechanisms such as facilitation and competition are relative, not absolute, and must be deﬁned within speciﬁc parameters to be useful. They suggested that, although the Connell and Slatyer (1977) models have been used to explain whole seres, the models best apply to speciﬁc interactions of species. For example, a shrub invading old ﬁelds facilitates the growth of tree seedlings by inhibiting grasses (Pickett et al., 1987b). Removal of the shrub at a later date would probably enhance the continued growth of the trees, and the trees may eventually inhibit the shrub. Clearly, competition, facilitation, and even tolerance are all involved, depending on the organism and successional stage involved. Pickett et al. (1987b) suggested a hierarchy of successional causes (cf. O’Neill et al., 1986) to help investigators identify both the organizational levels and the processes that must be speciﬁed before mechanisms that drive succession can be described. Walker and Chapin (1987) also noted the confusion arising from applying the Connell and Slatyer (1977) models to entire seres, and used their work on an Alaskan ﬂoodplain (Walker et al., 1986; Walker and Chapin, 1986) to demonstrate how tolerance, inhibition, and facilitation were each important to species change. They argued that succession is a result of the interactions of many processes which co-occur, and which vary in relative importance in different environments throughout succession (Walker and Chapin, 1987). They proposed that stochastic events (e.g., allogenic disturbance), facilitation, and the presence of mycorrhizae are more important determinants of species change in early primary succession than in secondary succession. Insect herbivory may be more important in late primary than late secondary succession, if the plants are stressed by drought or nutrient deﬁciency. Walker and Chapin (1987) suggested that experimental approaches are necessary

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to distinguish the relative importance of each process to species change. Matthews (1992) argued that none of the previously described models of succession adequately took into account the spatial and particularly the temporal heterogeneity of allogenic forces in glacial-foreland succession. He noted that landslides, streams, leaching, frost heaving, aeolian deposition, and micro- and mesoclimatic changes on glacial forelands continue to be important inﬂuences on succession long after the ice melts. Therefore, Matthews proposed a geoecological model in which the relative rate of decrease of allogenesis and increase of autogenesis through succession is dependent on environmental severity. He noted that patterns of species richness and biomass, differences in species composition among pioneer and mature communities, and the rate of succession can all be inﬂuenced by the relative importance of autogenic and allogenic processes, in both convergent and divergent succession. One problem with Matthews’ model is that autogenic and allogenic processes are ends of a continuum (White, 1979) and difﬁcult to separate. Burrows (1990) summarized many types of vegetation change, and suggested that the idea of a dichotomy between primary and secondary succession be dropped. For example, does primary succession become secondary succession after disequilibrium occurs from either autogenic or allogenic causes? He noted that many types of vegetation change (e.g., ﬂuctuations, cycles, direct replacement, response to gradual climatic change), or even the lack of change, are not generally included in models about succession. Instead, most models only address traditional linear replacement of species. He proposed a theory incorporating all types of vegetation change in ﬁve basic modes or patterns of change. Processes Cowles (1901) once termed succession a variable approaching a variable rather than a constant. Succession is still as great an enigma, despite a century of attempts to codify succession by end point (Clements, 1928), ecosystem properties (Odum, 1969), life-history attributes (Noble and Slatyer, 1980), population-level dynamics (Peet and Christensen, 1980), individual responses (Huston and Smith, 1987), hierarchical causes (Pickett et al., 1987b), process interactions (Walker and Chapin, 1987; Callaway and Walker, 1997), or the balance of external and internal forces (Matthews, 1992). However, a general consensus is being reached

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about the major processes that are important in directing successional change. These include stochastic events, allogenic disturbance, mycorrhizal interactions, soil development, life-history characteristics (patterns of differential establishment, growth and longevity), allelopathy, herbivory (including predation), disease, competition, and facilitation (Pickett et al., 1987b; Walker and Chapin, 1987). I now review how lifehistory characteristics and the processes of competition and facilitation apply to primary succession, and use one well-studied primary sere (Glacier Bay, Alaska, U.S.A.) as a case study to illustrate these processes. I choose these three processes as central to autogenic succession, although all the above-listed processes interact (Walker and Chapin, 1987). Life-history characteristics: An emphasis on lifehistory characteristics is the logical outcome of an increasingly reductionistic view of succession. Following Gleason’s lead that each plant species acts independently (Gleason, 1926), explanations of succession have been proposed based both on the population (Peet and Christensen, 1980) and on the individual (Huston and Smith, 1987). The models of Connell and Slatyer (1977) promoted an experimental approach to examining life-history characteristics. Establishment in primary succession has been considered dependent on site modiﬁcation (reaction or facilitation) and soil development (but see Drury and Nisbet, 1973); but studies of primary succession on glacial moraines (Adams and Dale, 1987), mine wastes (Finegan, 1984), and ﬂoodplains (Walker et al., 1986) indicate that establishment dynamics often resemble Egler’s (1954) initial ﬂoristics model – many species in the succession arrive very early in the sere. However, Matthews (1992) pointed out that initial ﬂoristics and facilitation were not mutually exclusive, and that “late successional” trees which colonize recently deglaciated moraines are often not very healthy (cf. Cooper, 1923). Matthews suggested that most colonization on glacial forelands resembles relay ﬂoristics (the orderly replacement of plant communities in response to environmental amelioration); yet such patterns could also merely be the consequence of differences in dispersal dynamics rather than indicate a requirement for site modiﬁcation. The stochastic nature of dispersal is perhaps predictable only in the context of a particular local ﬂora. Establishment of later successional species may require site improvement in the majority of primary seres. Differences in growth rates among plant species

Lawrence R. WALKER

during primary succession may determine the pattern of species replacements on glacial moraines (Burrows, 1990; Chapin et al., 1994) or ﬂoodplains (van Auken and Bush, 1985; Walker and Chapin, 1986). The tolerance model of Connell and Slatyer (1977) can be interpreted as the absence of species interactions (passive tolerance sensu Pickett et al., 1987b), succession being driven simply by differential arrival and growth, or the sorting of species by their increasing tolerance to reduced light and other resources through successional time (active tolerance: Pickett et al., 1987b). There are few data on the inﬂuence of relative growth rates or changes in physiological adaptations of species in succession (Pickett and McDonnell, 1989); most of the data come not from primary but from secondary succession (e.g., Bazzaz, 1979, 1996). Differential species longevity may determine species change in primary succession, but for long-lived trees such measurements are rare, and models must be used. Peet and Christensen (1980) suggested that, after a short initial establishment phase, mortality is the dominant process in thinning of stands of trees during secondary succession. In primary succession, there is typically a longer period of dominance by thicketforming shrubs or vines, and self-thinning of rapidly growing stands may not be as important. Competition: Competition has long been recognized as an important force structuring both plant and animal communities (Clements, 1916; Connell, 1983; Schoener, 1983; Tilman, 1988). However, ecologists are still not clear how important the role of competition in structuring communities is, relative to other forces such as predation, mutualism, or dispersal probabilities (Connor and Simberloff, 1979), how competition should be measured or deﬁned (Keddy, 1989), or how it varies with density (Berkowitz et al., 1995) and productivity (Grime, 1979; Thompson, 1987; Tilman 1987). The role of competition in succession was emphasized by Clements (1928), and competitive inhibition was the model for which Connell and Slatyer (1977) found the most supporting evidence. Although most evidence of competitive inhibition (earlier species delaying the establishment of later species) comes from secondary succession (e.g., Niering et al., 1986), it has also been demonstrated in primary seres. Competition is probably most important in mid-succession, when plant densities are often highest (Walker and Chapin, 1986; Matthews, 1992). However, competition may be important even in sparsely vegetated early primary succession, as a

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result perhaps of active root competition among plants that are efﬁciently scavenging the limited available resources (Matthews, 1992; Marrs and Bradshaw, 1993). The most common example of competition in primary succession is in the formation by early successional species of dense thickets that resist invasion by later colonists. Examples of this process occur on volcanoes (Tagawa et al., 1985; Whittaker et al., 1989; Walker and Vitousek, 1991), mine wastes (Prach, 1994), glacial forelands (Heusser, 1956; Chapin et al., 1994), dunes (Morris et al., 1974; Moore, 1982; Binggeli et al., 1992), landslides (Walker, 1994), and ﬂoodplains (Halwagy, 1963; Nanson and Beach, 1977; Luken and Fonda, 1983; Walker and Chapin, 1986). This type of competition is generally asymmetric (Keddy, 1989). The early successional, thicket-forming species may simultaneously inhibit the later-colonizing understory species through light reduction or root competition for water or nutrients, and facilitate their growth through the addition of nutrients and organic matter, and perhaps by protection from herbivory (Bryant and Chapin, 1986; Walker and Chapin, 1987; Chapin et al., 1994). Competition for light is central to species replacements in primary succession (Cooper, 1931; Lawrence, 1979; Huston and Smith, 1987; Tilman, 1988; Olff et al., 1993) and is dependent more on relative height and leaf area than on plant density. Regeneration dynamics under a canopy (either low light tolerance or rapid growth of suppressed seedlings in a canopy gap) are clearly critical in determining the species composition of vegetation during primary succession (Burrows, 1990). Competition and facilitation can be operating simultaneously in many environments (Binggeli et al., 1992; Callaway, 1995; Callaway and Walker, 1997) but the net effect may be all that inﬂuences succession. Indirect effects (species A inhibiting species B by facilitating species C, a competitor of species B) also complicate the successional outcomes of species interactions (Walker and Chapin, 1987). Facilitation: Facilitation has long been assumed to be essential (obligatory facilitation) for species replacements to occur in primary succession. Clements (1928) used “reaction” to describe the modiﬁcation of the physical environment by plants. Matthews (1992) distinguished “reaction” from “facilitation”, which he deﬁned as the positive consequence of the effects of

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an early colonizer on establishment and/or growth of later colonizers. Experimental tests of the importance of facilitation (Davis et al., 1985; Walker and Chapin, 1987; Wood and del Moral, 1987; Blundon et al., 1993; Chapin et al., 1994) have often found evidence for non-obligatory or facultative facilitation (a later successional species grows better after the environment has been altered by earlier colonists, but all species can establish in early successional environments). Obligatory facilitation may be most critical in very severe primary seres (e.g., Birks, 1980), although even on mine tailings or rock outcrops late-successional trees can establish in rock cracks (Grubb, 1987; Miles and Walton, 1993), and Matthews (1992) concluded that there is no evidence for obligatory facilitation on glacial forelands. In less severe seres such as ﬂoodplains or landslides, obligatory facilitation is even less likely to occur. The mechanisms of facilitation can occur in a variety of ways in primary succession. Plant dispersal to a site can be facilitated by trees or posts providing perches for birds which subsequently defecate seeds McDonnell and Stiles, 1983; Vitousek and Walker, 1989), by vegetation that captures seeds (Day and Wright, 1989; Choi and Wali, 1995), or by animals directly or indirectly involved in dispersal (e.g., Matthews, 1992). Plants that ameliorate the physical environment may facilitate colonization by less tolerant species. Dunes (Ayyad, 1973; Moreno-Casasola, 1986) or ﬂoodplains (Johnson et al., 1985; Friedman et al., 1996) can be stabilized by plants. Wind speeds are reduced on glacial forelands by plant cover (Viereck, 1966). Shade is provided in arid lands (e.g., the nurseplant effect: Niering et al., 1963; Turner et al., 1966; McAuliffe, 1984; Valiente-Banuet and Ezcurra, 1991). The build-up of soil or snow around plants affects microclimate in ways that may reduce damage to later colonizers from ﬂooding (Nechaev, 1967; Menges and Waller, 1983), extreme temperatures (see examples in Matthews, 1992) or high concentrations of salts (Vasek and Lund, 1980). Soil accumulation is generally associated with an increase in organic matter, which is also considered essential for growth in severe environments (Birks, 1980; Adachi et al., 1996). Plants with nitrogen-ﬁxing symbionts (Walker, 1993) or nonsymbiotic ﬁxers (Vitousek et al., 1987) can improve soil nutrient status for later colonizers on most types of primary succession, including volcanoes (Vitousek et al., 1987; Wood and del Moral, 1987; Clarkson, 1990), mine wastes (Palaniappan et al., 1979; but

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see Larson and Vimmerstedt, 1983), glacial forelands (Crocker and Major, 1955; Birks, 1980; Blundon et al., 1993), dunes (Olson, 1958), and ﬂoodplains (Van Cleve et al., 1971). Burrowing animals can facilitate plant establishment by reducing competition and enhancing nutrient turnover (Andersen and MacMahon, 1985). Animals that break down substrates in the earlier stages of heterotrophic succession facilitate both the decomposition process and the establishment of later successional heterotrophs (Schoenly and Reid, 1987). Bacteria and fungi may facilitate the development of vascular plant communities in most primary seres but are not considered necessary precursors (Walton, 1993). The process of facilitation is now recognized as universal (Callaway, 1995). Its role is no longer seen as obligatory for primary succession, but it may accelerate species change (Marrs and Bradshaw, 1993; Chapin et al., 1994). Studies can now be focused on the relative balance of facilitation and competition throughout primary succession (Connell and Slatyer, 1977; Walker and Chapin, 1987; Callaway and Walker, 1997). A case study: Glacier Bay, Alaska The glacier forelands of Glacier Bay constitute perhaps the most cited example of primary succession in the world, and are among the best-studied. Detailed observations of plant species change began in 1916 when Cooper (1923) established permanent vegetation plots on surfaces less than 40 years old throughout the west arm of the ﬁord system. The regular monitoring of these plots (Cooper, 1923, 1931, 1939) and subsequent studies (Lawrence, 1958, 1979; Reiners et al., 1971; Lawrence et al., 1967; Noble et al., 1984; Chapin et al., 1994; Fastie, 1995) represent some of the longest sequences of vegetation change in the world that have been continuously recorded. Detailed geological (Field, 1947) and photographic records, as well as oral histories from indigenous populations (Lawrence, 1958) and dendroecological measurements (Fastie, 1995), have veriﬁed the dates of ice removal from various locations in the bay (Field, 1947; McKenzie and Goldthwait, 1971), allowing researchers to age the plant communities accurately. I will summarize the history of successional studies at Glacier Bay to illustrate four ongoing debates within studies of primary succession: a) the predictability of dispersal dynamics and their importance to successional trajectories; b) the importance of facilitation; c) the use of

Lawrence R. WALKER

chronosequences; and d) the possibility of generalizing from one sere to another. Cooper (1923) and others (Crocker and Major, 1955; Lawrence et al., 1967) have noted the spatial heterogeneity of colonization at Glacier Bay, generally attributing it to microsite suitability or the stochasticity of dispersal (Crocker and Major, 1955; Chapin, 1993). Following early colonization by a cryptogamic crust (Worley, 1973; Lawrence, 1979), the small nitrogenﬁxing (Lawrence et al., 1967) rosaceous shrub Dryas drummondii often forms a dense mat. Mid-successional alder (Alnus sinuata) (also a nitrogen-ﬁxer) colonizes the East Arm of Glacier Bay, but not the West Arm, usually taking several decades to expand to dense thickets from isolated, colonizing clusters (Lawrence et al., 1967). Cottonwood (Populus trichocarpa) trees sometimes form abundant stands, colonizing before or during the expansion of alder. Spruce (Picea sitchensis (spruce)) trees dominate later stages of succession, but may colonize newly de-glaciated moraines, particularly if the distance to mature spruce forests is not great (Cooper, 1931; Chapin et al., 1994). Finally, there is debate about whether gradual paludiﬁcation (expansion of bogs from rising water tables) leads to a ﬁnal stage dominated by western hemlock (Tsuga heterophylla) trees interspersed with bogs (Lawrence et al., 1967; Reiners et al., 1971; Ugolini and Mann, 1979). Do dispersal dynamics set the course for the entire succession? Why is the vegetation so different in the two arms of the bay? How many successional trajectories are there? Is there eventual convergence to a climax stage? These questions, important in all successional studies, are not yet answered, even at well-studied Glacier Bay. The answers hinge in part on a better understanding of the role of dispersal and establishment of the dominant plant species. Glacier Bay has been used as a classic example of facilitation. This is largely due to several inﬂuential studies, including those of Crocker and Major (1955) and Lawrence et al. (1967). Crocker and Major (1955) used the series of surfaces of known age at Glacier Bay to provide one of the earliest well-documented links between soil and vegetation development in primary succession. They showed that alder increased soil nitrogen and soil organic matter and caused soil pH to decline, but merely speculated about how soil nitrogen affected plants. In fact, they are most cited for their speculation that the intriguing drop in their total soil nitrogen values after 100 years results from the loss of the alder stage and the uptake of soil

PATTERNS AND PROCESSES IN PRIMARY SUCCESSION

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Fig. 25.3. Effects during four successional stages on natural and transplanted spruce (Picea sitchensis) seedlings in Glacier Bay, Alaska. SOM, soil organic matter (from Chapin et al., 1994).

nitrogen by the developing spruce forest (Crocker and Major, 1955). Lawrence and colleagues (Lawrence, 1958, 1979; Schoenike, 1958; Lawrence et al., 1967) experimentally examined the role of alder and Dryas drummondii. They concluded that the addition of alder leaves (and to a lesser extent the addition of Dryas leaves) increased cottonwood growth and that nitrogen ﬁxed by both alder and Dryas was inﬂuential in determining plant succession. More recent studies (Ugolini, 1968; Bormann and Sidle, 1990) have conﬁrmed the important effect plants such as alder have on soil development at Glacier Bay. Chapin et al. (1994) showed experimentally that alder facilitated growth of spruce seedlings through additions of organic matter and nitrogen (Fig. 25.3). However, they also provided evidence that alder thickets inhibit growth of spruce seedlings through competition for light and at the root level, and supported earlier conclusions (Cooper, 1923) that spruce did not need alder to be present in order to establish and grow (obligatory facilitation). Vegetation at each stage inhibited establishment of species of the following stages, life histories of the species determined the overall pattern of succession, and species interactions determined the rate of succession (Chapin et al., 1994). A recent study by Fastie (1995) questioned the validity of the assumption in the Crocker and Major (1955) chronosequence that the older and younger spruce forests had a similar biological history. Succession at the mouth of the bay, with mature spruce forests in close proximity to the newly exposed terrain, probably did not have the prolonged alder stage that

the more remote and more recently deglaciated sites further inland had (Chapin et al., 1994; Fastie, 1995). Therefore, the apparent decrease of soil nitrogen during the early spruce stage reported by Crocker and Major (1955) and used by many as the basis for the argument that facilitation is important at Glacier Bay, resulted from multiple pathways of succession occurring at different nitrogen levels, not from in situ reduction of soil nitrogen. Fastie (1995) attributed the multiple pathways at Glacier Bay to dispersal and establishment dynamics (e.g., certain species can not establish either early or late in succession), to species interactions (alder both facilitates and inhibits spruce growth), and to landscape position (alder dominates where distance from seed source is greatest on the younger moraines). Is succession at Glacier Bay typical of other seres on glacial forelands or even of primary succession in general? Walker (1995) made a detailed comparison between primary succession at Glacier Bay (Chapin et al., 1994) and on the ﬂoodplain of the Tanana River in central Alaska (Walker et al., 1986; Walker and Chapin, 1986, 1987; Walker, 1989). Very similar methods were used in the two studies, and species composition during succession was similar. Walker found that higher initial levels of nitrogen in the ﬂoodplain soils reduced the importance of facilitation by alder. He further showed that soil nitrogen accumulation is slower at Glacier Bay than on other glacial forelands with vascular species present that have nitrogen-ﬁxing symbionts (Walker, 1993, 1995). He suggested that the slow nitrogen accumulation coupled with the mesic environment of coastal Alaska may provide optimal conditions for alder

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to grow and facilitation to be important. Matthews (1992) also concluded, after an extensive survey of glacial foreland succession, that Glacier Bay was wetter and warmer than many seres on glacial forelands, perhaps favoring the dominance by alder and more rapid forest development. The evidence of facilitation at Glacier Bay must therefore not be taken to imply that facilitation is equally important elsewhere.

SUMMARY AND FUTURE DIRECTIONS

Many reﬁnements have been made to the basic processes that Clements (1916) identiﬁed as important in driving succession, but no new processes have been added. Dispersal is still unpredictable, despite creative efforts to study (Zasada and Lovig, 1983; Augspurger and Kitajima, 1992) and model it (Janzen, 1970; Hubbell, 1979; Becker et al., 1985). Most seed dispersal is over short, not long distances, so that the position of the disturbance in the biotic landscape will be an important determinant of which species establish ﬁrst. Site modiﬁcation (reaction, facilitation), once considered obligatory for primary succession, although without experimental evidence, is now recognized as facultative and more important in determining the rate than the direction of succession (Chapin et al., 1994). A full range of types of facilitative effects in primary and secondary succession has recently received more attention (Callaway, 1995; Callaway and Walker, 1997). Competition has had its supporters (e.g., Tilman, 1988) and detractors (e.g., Connor and Simberloff, 1979) since Clements’s day, but clearly also has a role in determining the rate, if not the direction, of primary succession (Walker and Chapin, 1987). Despite the emphasis on the same processes for almost a century, in more recent years contributions have been made by modeling how those processes might interact through succession (Connell and Slatyer, 1977; Huston and Smith, 1987; Pickett et al., 1987b; Walker and Chapin, 1987; Tilman, 1988) on the basis of the life history, density, and physiology of the interacting species (Callaway and Walker, 1997). Future studies should focus on several issues. First, good autecological data are needed on key species. Without such data, the importance of life history characteristics and physiological adaptations cannot be assessed. Measurements of seed rain, conditions needed for seed germination and seedling establishment, and measures of longevity are critical. Second, long-term

Lawrence R. WALKER

monitoring of permanent plots is essential. One cannot rely on sequences in space to represent sequences in time. If researchers revisited old ﬁeld sites, valuable data could be collected without waiting for decades of change to occur. Third, experiments that tease apart the relative importance of competition and facilitation will advance understanding of species interactions. Do they only alter rates of succession, or do they also alter trajectories? Fourth, the effects of plants on soils and soils on plants need to be studied. Future research might continue studying the nutrient implications of nurse–plant interactions and the role of nitrogen ﬁxers in primary succession. Fifth, the spatial relationships of plants (e.g., clusters, and nurse–plant effects involving light and water competition) can offer insight into species interactions, especially when measured across successional time. Sixth, plant–animal interactions are little understood in a successional context, including the importance of herbivory and parasitism. Finally, I would emphasize the importance of comparisons within and among types of primary succession, perhaps using explicit climatic gradients (e.g., soil moisture or air temperature). Without such comparisons, conclusions remain site-speciﬁc. Standardization of measurements (particularly for soil sampling) would aid cross-site comparisons. The study of primary succession is relevant to society principally to the extent it enables one to offer suggestions on the best methods to reclaim damaged areas. Useful predictions for managers depend on robust generalizations that are not site-speciﬁc.

ACKNOWLEDGMENTS

Comments by R. Callaway, C. Fastie, and S. Smith improved the manuscript and D. Dean assisted with Figs. 25.1 and 25.2. Development of the ideas presented here occurred while I was supported by NSF grants 88-1789, BSR-8811902, and DEB-9411973. The latter two were to the Terrestrial Ecology Division, University of Puerto Rico, and the International Institute of Tropical Forestry as part of the Long Term Ecological Research Program in the Luquillo Experimental Forest.

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Chapter 26

PLANT INTERACTIONS DURING SECONDARY SUCCESSION Scott D. WILSON

INTRODUCTION

Interactions among plants are thought to drive secondary succession, but the relative importance and, in some cases, even the direction (positive or negative) of these interactions is not well known. Plant interactions have received surprisingly little attention in succession studies. Rejm´anek (1990) reviewed 518 studies of old ﬁeld succession from 1910 to 1988, but only 18 dealt with competition. Similarly, recent books on primary (Miles and Walton, 1993) and secondary succession (Osbornov et al., 1990; Glenn-Lewin et al., 1992) have no chapters devoted to competition. In his list of processes inﬂuencing succession, Bradshaw (1993) listed plant interactions last, partly because they occur only after earlier processes such as dispersal, but possibly also because of the practical challenges associated with studying interactions experimentally. Taking a different approach, Jackson (1981) found 332 papers published before 1959 dealing with competition, and 30% of these concerned succession. These early papers established that plants can compete during succession, and were followed by predictions about how plant interactions should contribute to succession (e.g., Drury and Nisbet, 1973; Grime, 1973; Horn, 1974; Connell and Slatyer, 1977; Connell, 1978; Huston, 1979; Tilman, 1985; Huston and Smith, 1987; Walker and Chapin, 1987; Peet, 1992). Some predictions are difﬁcult to test. For example, disturbance should reduce competition (Grime, 1973; Huston, 1979), but just a decade ago such a reduction was thought to be difﬁcult to measure (Orians, 1986). Measures of competition at various times since disturbance and at different levels of disturbance have appeared since (reviewed by Goldberg and Barton, 1992; Gurevitch et al., 1992). My objective here is to survey recent studies of plant interactions during

secondary succession, and to complement previous reviews (Van Hulst, 1978; Finegan, 1984; Pickett et al., 1987; Glenn-Lewin and van der Maarel, 1992). A better understanding of plant interactions may also increase understanding of plant invasions, because many successful plant invasions involve introduced species from early successional stages which colonize disturbances and prevent the establishment of latersuccessional native species. For example, introduced annual grasses have displaced native perennial grasses from parts of western North America (Mack, 1986) and introduced perennial grasses exclude native shrubs in Hawaii (D’Antonio and Vitousek, 1992). Plant interactions can be summarized as net negative or positive effects. I discuss these as competition and facilitation. I address secondary succession (vegetation change following biomass removal) as opposed to primary succession (vegetation change on new substrate) for two reasons. First, secondary succession is more common and therefore more often studied than primary succession. Second, the ever-increasing size, frequency, and intensity of disturbance by humans will increase both the area of land undergoing secondary succession and the need to understand mechanisms driving succession. Plant interactions during primary and secondary succession are compared brieﬂy near the end of this chapter. Plant interactions vary through successional time, so the terms “early” and “late” will appear, but these are always in the context of the plants under discussion, so that, during a typical temperate old-ﬁeld succession, “early in succession” for an annual species occurs before “early in succession” for a woody species. Four general ideas about plant interactions during secondary succession recur in reviews and models, and will be evaluated here. These are that facilitation

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is most important early in succession; that competition increases in importance over time; that latesuccessional species are better competitors than those present early in succession; and that competition causes late-successional species to have narrower niches than early species. More than one of these ideas may be addressed by a single experiment. For example, an increase in the growth of a plant following removal of neighbors indicates competition whereas a decrease indicates facilitation, so that both competition and facilitation may be tested in the same experiment. Secondly, growing one target species with and without neighbors at different successional stages gives information about the intensity of competition (or facilitation) in the different successional stages, as well as information about the competitive effects of the species dominating the different stages. Lastly, using more than one target species gives the same two kinds of information noted above, but also allows comparisons of the competitive responses of the targets. As a result, some experiments are evaluated more than once in this review. Each evaluation pertains to the particular question being addressed. The strongest tests of ideas about interactions during succession come from experiments that allow explicit testing of statistical interactions between, for example, competition and successional stage, or between suppression and the identity of target species (Goldberg and Barton, 1992). Much can also be learned, however, by cautiously comparing the results of separate studies in different successional stages. I therefore include garden, greenhouse, growth-chamber, and pot experiments, mostly as examples of interactions during the ﬁrst stages of succession. I have included many kinds of studies from a variety of habitats and experiments. Generalizations are therefore built on comparisons of widely differing systems and investigations, which readers may interpret as a strength or a weakness. The generalizations are skewed by the fact that many studies originate from one system, old ﬁelds on abandoned cropland, in one region, the northeastern United States. Many experiments involve more than one species, habitat, or life-history stage, and a frequent outcome of such complex studies is that a hypothesis is supported by some combinations of variables but is rejected by others. When summarizing such outcomes I refer to all possible combinations as “cases”. For example, a study which ﬁnds facilitation in only three species among the

Scott D. WILSON

eight examined, in only one environment out of four examined, is summarized as ﬁnding facilitation in 3 out of 32 cases. This review is organized around three approaches to studying interactions. Pattern analysis includes descriptions of variation in resources, individuals, populations, and communities among different successional stages. The main limitation in understanding process from studies of pattern is that observed patterns might be produced by processes other than plant interactions (Horn, 1981; Keddy, 1989; Lepˇs, 1991) or successional stage (Pickett, 1989). This problem can be overcome to some extent through the use of experiments (Hairston, 1989), keeping in mind that experiments are limited by their unknown impacts on other factors, their inability to distinguish indirect effects (Connell, 1983, 1990), and their generally limited spatial and temporal extent (Tilman, 1989; Goldberg, 1995). The second approach, population-level experiments, measures groups of plants, whereas the third approach, individual-level experiments, measures single plants. Population- and individual-level interaction experiments are distinguished because individual plants react only to neighbors, whereas individuals in a population respond both to neighbors and to other experimental plants. Results from population-level experiments can be difﬁcult to interpret. If, for example, the population size of an annual species naturally decreases during succession, the population might show little competitive release following removal of interspeciﬁc neighbors early in succession, simply because there are few interspeciﬁc neighbors (Silvertown and Dale, 1991). The same species should show more competitive release following removal of interspeciﬁc neighbors later in succession simply because it would have more space to ﬁll. In this instance, the annual population is responding partly to variation in its own size, so that unequivocal measurements of neighbor effects are not possible. Another example involves richness. If richness is lower at one stage than another, and the most abundant species is removed from both stages, then the competitive release would be greatest in the stage with lowest richness, because the numerical subordinates would have more space to ﬁll. Thus, removal experiments at the population level can indicate that species with relatively large mass at one time during succession do indeed have large competitive effects at that time, but population removals cannot be used to infer the per-plant competitive effects of the neighbors, or the competitive abilities of the released species.

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

Interpretations of the responses of individual plants are simpler. The boundary between population- and individuallevel experiments is generally clear, but some experiments manipulate groups of plants consisting of individuals which probably do not interact with each other, and these I have included as individual-level experiments. For example, a population of seeds may be sown into a ﬁeld, and germination recorded, but the newly germinated seedlings probably do not interact, so that their responses reﬂect individual behavior. The same thing can happen when transplants are planted in a plot at such a low density that they do not inﬂuence each other. Kindell et al. (1996) have presented a method to test this assumption. One experiment can measure both individual responses such as germination, and population responses such as growth in the resulting population of seedlings.

DOES FACILITATION DECREASE WITH AGE?

Neighbors might enhance plant performance in early succession by providing shade, protection from grazers, and habitat for seed-dispersing animals. Neighbors may be especially beneﬁcial in primary succession (Connell and Slatyer, 1977; Walker and Chapin, 1987; Bradshaw, 1993) where there is initially little plant cover. Further, neighbors early in primary succession may increase soil fertility by adding organic matter and nitrogen (Del Moral and Wood, 1993; Chapin et al., 1994). Facilitation may be less likely during secondary succession where soil nutrients are present from the beginning and competing vegetation quickly covers a disturbed site. Pattern analysis Positive associations between species are often interpreted as facilitation. Early-successional annuals and biennials show positive associations more often than do later-successional perennials (Myster and Pickett, 1992), suggesting that facilitation is more prevalent early in succession. Early-successional woody neighbors, mostly shrubs, increased the survival of transplanted trees (Pseudotsuga menziesii) in New Mexico (Coffman, 1975), and tree seedlings were more common among shrubs than in herbaceous vegetation in old-ﬁelds in Michigan (Werner and Harbeck,

613

1982). Positive associations between early- and latesuccessional trees occurred in one out of three stands examined in Colorado subalpine forest (Rebertus et al., 1991). Some pattern analyses indicate that facilitation is not restricted to early succession. Interspeciﬁc neighbors enhanced growth in Pinus strobus–Populus tremuloides mixtures in a 60-year-old Michigan forest (Peterson and Squiers, 1995), and the frequency of positive associations among species in a New Jersey old-ﬁeld increased from 7% to 16% during the ﬁrst 13 years (Myster and Pickett, 1992). An alternative explanation for positive associations, however, is that resource-rich patches promote the establishment of more than one species (Callaway, 1995). Stronger evidence for facilitation would be provided by the establishment of late-successional species only after early-successional species had colonized a site (Finegan, 1984), but even this pattern could be caused by resource-rich patches if the late arrivals simply have lower dispersal abilities. Other analyses ﬁnd little evidence for facilitation early in succession. Positive associations were not found during early years of succession following ﬁre in New Zealand grasslands (Gitay and Wilson, 1995). Tree-seedling mortality measured for two years after ﬁre in Eucalyptus scrub increased with increasing seedling density (Wellington and Noble, 1985), a result which supports competition, not facilitation. Population-level experiments If facilitation decreases with age, then facilitation should be most apparent in early-successional communities, but experiments in populations of old-ﬁeld annuals have produced no examples of facilitation. Removal experiments showed that annual and perennial populations had no effect on each other in a 1-year-old Ohio ﬁeld (Hils and Vankat, 1982). Similarly, neighbor removal had no negative effect on early-successional species in tilled ﬁelds in Michigan (Miller and Werner, 1987) or Oregon (McEvoy et al., 1993). Miller (1994) used removals in a tilled ﬁeld in Michigan to test for indirect positive effects of neighbors. Indirect positive effects occurred, and partially mitigated negative effects, but the net effect was usually negative. Armesto and Pickett (1985) removed the dominant annual Ambrosia artemisiifolia from a 2-year-old New Jersey ﬁeld and the dominant perennial Solidago canadensis from a 7-year-old ﬁeld; neighbor biomass

614

did not decline as a result in either case, showing that facilitation did not occur. Comparisons of pure stands and mixtures of oldﬁeld species grown along crossed gradients of fertility and disturbance produced no examples of facilitation in one garden experiment (Campbell and Grime, 1992), but, in a similar experiment, one out of four species were facilitated by neighbors in disturbed treatments (Turkington et al., 1993). Experiments in some established perennial populations have found facilitation. Single-species removals in perennial salt marsh vegetation showed facilitation in two out of six cases (Bertness, 1991). In the salt marsh, plant removal created harsh conditions similar to those of early primary succession. Plots without vegetation experienced more sunlight, higher rates of evaporation, and higher salt concentrations in soil water. All of these conditions were ameliorated by neighbors, so that plant interactions in the salt marsh were beneﬁcial, as frequently occurs during primary succession (Walker and Chapin, 1987). Evidence for facilitation in perennial vegetation was also found by Belleﬂeur and Villeneuve (1984). Neighbor removal in clear-cut Quebec forest decreased the growth of commercial tree species in two out of three cases in the ﬁrst year, but this effect disappeared after four years. Similarly, removal of herbs and shrubs in a North Carolina clear-cut decreased tree mass in the ﬁrst year, but not in subsequent years (Wilson and Shure, 1993). Woody plants may also facilitate each other. The growth of seedlings of the early-successional shrub Ailanthus altissima in New Jersey old-ﬁelds was enhanced by leaf litter from oaks (Quercus spp.), apparently because the litter suppressed herbaceous perennials (Facelli, 1994). The dispersal of seeds of woody species into old-ﬁelds was signiﬁcantly increased by adding cut saplings in New Jersey (McDonnell and Stiles, 1983; McDonnell, 1986) and by the presence of an earlysuccessional shrub in abandoned Amazonian pastures (Vieira et al., 1994). The attraction of birds to perches may contribute to the presence of woody seeds in soil under early-successional woody plants in Illinois, and explain their absence from soil in recently abandoned ﬁelds dominated by herbaceous vegetation, as was found in Illinois by Burton (1989). Individual-level experiments Neighbor removal in a 1-year-old ﬁeld in Michigan

Scott D. WILSON

decreased emergence of Daucus carota by 17%, whereas removal in a 3-year-old ﬁeld increased emergence by 52% (Holt, 1972), suggesting that neighbor effects switched from positive to negative as succession proceeded. Similarly, transplants of three out of eight herbaceous species were facilitated by neighbors in tilled plots in Minnesota, whereas none were facilitated and most were suppressed in 40-year-old undisturbed plots (Wilson and Tilman, 1995). McConnaughay and Bazzaz (1990) grew single annual target plants in gaps in perennial grass swards. The addition of a single annual neighbor of a different species facilitated growth of the target individual in one case out of nine. Pairwise mixtures of old-ﬁeld species grown in a greenhouse showed facilitation in only one out of 49 cases (Goldberg and Landa, 1991). These results suggest that facilitation early in secondary succession affects only a small subset of species. Facilitation is also weak or uncommon in experiments involving individual woody plants. Facilitation of tree establishment from experimentally sown seeds decreased over time in two out of 13 cases in Illinois (Burton and Bazzaz, 1991). In three cases, the opposite pattern was found: facilitation increased with successional age; woody seedlings established best under young shrubs and trees, after the annual and perennial stages. In the same study, early-successional neighbors increased the survival of one woody species out of the four examined: survival of Prunus serotina was about 50% without neighbors but 95–100% with the perennials Festuca pratensis and Solidago altissima (Burton and Bazzaz, 1995). The impact of neighbors on photosynthesis or water potential was never positive (Burton and Bazzaz, 1995). Neighbors increased woody plant establishment in four out of nine cases in North Carolina old-ﬁelds, primarily by sheltering seeds and seedlings from predators (De Steven, 1991a). In this case, however, this effect did not vary signiﬁcantly between newly tilled plots and 10-year-old vegetation. Woody plant survivorship and growth were not facilitated by neighbors in ﬁelds of either age (De Steven, 1991b). Summary of facilitation during secondary succession There are examples of positive associations among species early in succession, but there are exceptions to this, and positive associations also occur late in succession. Further, positive associations might be

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

caused by processes unrelated to facilitation, so pattern analyses do not provide strong evidence for facilitation early in secondary succession. Experiments suggest that facilitation early in secondary succession affects relatively few species. Studies which supported early-successional facilitation tended to do so with only a minority of species examined (Table 26.1). Facilitation is likely to occur where neighbors ameliorate harsh abiotic factors, but, in secondary succession, these beneﬁts are often obscured by resource competition from the same neighbors (Bertness and Callaway, 1994). Facilitation could arise through indirect competitive effects (Miller, 1994; Facelli, 1994), but the net negative effects of neighbors probably outweigh indirect beneﬁts. Perhaps the most promising avenue for exploring facilitation during secondary succession is in especially stressful habitats, where interactions among plants and abiotic factors might be expected to be similar to those in primary succession.

DOES COMPETITION INCREASE WITH AGE?

Competition might increase during secondary succession because the ratio of resource demand to supply increases with increasing standing crop (Goldberg, 1987; Taylor et al., 1990). It seems reasonable that there is little competition early in primary succession before individuals ﬁll available space (Walker and Chapin, 1987; Bradshaw, 1993), and this might also apply to certain cases in secondary succession. Fire-prone trees, for example, may rely on decreased competition following disturbance for establishment opportunities (Mutch, 1970; Clark, 1991). Alternatively, competition in secondary successions might begin within days of disturbance, on account of the presence of seeds and vegetative propagules, and soils which, unlike those found in primary succession, are capable of supporting rapid plant growth (Marks and Mohler, 1985; Grubb, 1987). The high growth rate and high tissue nutrient concentrations of early successional species should provide intense competition early in succession (Bazzaz, 1979, 1986). Competition is very intense in agricultural ﬁelds, recently abandoned ﬁelds, and forest clear-cuts despite disturbance and abundant nutrients (Parrish and Bazzaz, 1982; Radosevich and Rousch, 1990; Pacala and Silander, 1990; Harvey et al., 1993; Cousens and Mortimer, 1995). The importance of gaps for recruitment into both young and

615

old grasslands (Platt and Weis, 1985; Goldberg and Gross, 1988) suggests that competition is intense in both. Pattern analysis In the same way that positive associations may reﬂect facilitation, negative associations among species might reﬂect competition, and increasing competition through time might cause an increase in negative associations. Instead of increasing, however, the proportion of negative associations decreased from 16% to 4% during the ﬁrst 31 years of succession in a New Jersey oldﬁeld (Myster and Pickett, 1992). Similarly, associations over 27 years following ﬁre in New Zealand grasslands did not become more negative (Gitay and Wilson, 1995). Thus, species associations do not suggest that competition increases with time. Competition might increase through succession if resource supply remains constant while plant mass and demand increase. Aber (1979) surveyed a range of forest ages in New Hampshire and found that leaf area index achieved its maximum value after just 3 years, suggesting that light competition rapidly increases in importance. Available soil nitrogen in old ﬁelds did not increase during decades of succession in Minnesota or Illinois, even though total nitrogen increased as soil organic matter accumulated (Robertson and Vitousek, 1981; Gleeson and Tilman, 1990; Burton and Bazzaz, 1995). The discrepancy between available and total nitrogen presumably reﬂects the high demand for nitrogen by vegetation, and suggests that competition for nitrogen is strong at all ages. Available nutrients can also decrease over succession, and for reasons unrelated to plant uptake. The accumulation of organic matter during succession in boreal forest in Alaska insulates and cools the soil, decreases decomposition rates, reduces nutrient mineralization and availability (Van Cleve and Viereck, 1981), and should increase nutrient competition. An increase in competition may be reﬂected by an increase in biomass allocation to resource-garnering structures. Community root allocation increased through time in Minnesota old-ﬁelds, presumably because community mass increased while the availability of limiting nutrients (e.g., nitrate) stayed constant (Gleeson and Tilman, 1990). Similarly, late successional trees in the temperate zone have shallower rooting patterns than early successional species (Gale and Grigal, 1987),

616

Scott D. WILSON

Table 26.1 Summary of studies that address the hypothesis that facilitation decreases (FD) during secondary succession or that competition increases (CI) Reference

Range 1

Levels 2

N3

System 4

Targets 5

Var 6

Results 7 FD

Pattern analysis Christensen and Peet (1981)

8–70

f

12

F

1w

s,r

Gitay and Wilson (1995)

1–27

11

1–4

P

c

sc

Gleeson and Tilman (1990)

1–60

28

1–2

OF

c

al

McDonnell and Stiles (1983)

3,13

2

1

OF

c

g

Myster and Pickett (1992)

1–31

f

10

OF

c

sc

Peterson and Bazzaz (1978)

1,2

2

1

OF

1a,1p

sc

Rebertus et al. (1991)

~100–400

3

2–6

F

c

sc

Schlesinger and Gill (1978)

0–20

3

1–7

SH

1w

s

Stewart and Thompson (1982)

0,Und

3

1

P,SH

6p

al

Yarranton and Yarranton (1975)

0–45

f

1

F

1w

s

Zobel et al. (1993)

5–>150

5

1

F

c

sc

Wellington and Noble (1985)

0–2

f

2

SH

1w

g

Werner and Harbeck (1982)

10,16

2

1

OF

c

sc

Armesto and Pickett (1985, 1986)

2,7

2

1

OF

c

sc

Belleﬂeur and Villeneuve (1984)

1–4

f

30

F

3w

r

Bertness (1991)

0,Und

2

10

W

3p

r

√ √

x

√

√ √ √

x √ 33

0

5

3

G

2a,4p

r

Gross and Werner (1982)

1–15

3

1

OF

4a

g

Hils and Vankat (1982)

0

1

4

OF

c

r

√ √ x √ √

x √

33

x

1,3

2

5

OF

1p

g

Marks and Mohler (1985)

0–1

f

2

OF

c

sc

McDonnell and Stiles (1983)

2

1

1

OF

c

g

McEvoy et al. (1993)

0–3

f

4

P

1a

r

x

Miller and Werner (1987)

1–2

1

1

OF

2a,3p

r

x

Monk and Gabrielson (1985)

3,15

2

1

OF

c

r

Turkington et al. (1993)

0

5

3

G

4p

r

Wilson and Shure (1993)

0–2

f

4

F

c

r

Burton and Bazzaz (1991)

0–19

9

1

OF

6w

g

Burton and Bazzaz (1995)

0–19

9

1

OF

4w

s

x √ √ √ √ √

√

√

66

√

x √

Holt (1972)

Individual-level experiments

√ √

Population-level experiments

Campbell and Grime (1992)

CI

°

√ 25

x

x √ √

Burton and Bazzaz (1995)

ph

x

Burton and Bazzaz (1995)

wp

x

Burton and Bazzaz (1995)

r

√

15 25

√

23

x √ √ x

De Steven (1991a)

1,10

2

2

OF

5w

g

De Steven (1991b)

1,10

2

2

OF

4w

s

x°

44

Goldberg and Landa (1991)

0

1

7

G

2a,5p

r

x

x° √ 25

continued on next page

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

617

Table 26.1, continued Reference

Range 1

Levels 2

2

N3

System 4

Targets 5

Var 6

McConnaughay and Bazzaz (1990)

0–1

10

P

3a

r

Reader et al. (1994)

10–25,Und 9

3–6

OF,P

1p

r

Wilson and Tilman (1993)

0–38

2

10

OF

1p

r

Wilson and Tilman (1995)

0–38

2

10

OF

4a,4p

r

Results 7 FD √ 11

CI √ √

√

38

√

° °

1

Community ages considered (yr); 0, newly disturbed ground; Und, undisturbed ground. Number of successional stages considered, for example, Gleeson and Tilman (1990) examined 28 different ages ranging from 1 to 60 years; f, follow-through study. 3 Number of replicates of successional stages. 4 F, forest; G, greenhouse or garden; P, permanent grassland; OF, old-ﬁeld; SH, shrubland; W, wetland. 5 Number and life histories of study species; a, annuals or biennials; c, community as a whole; p, herbaceous perennials; w, woody plants. 6 Variable measured; al, allocation; g, germination, dispersal or establishment; ph, photosynthetic rate; r, growth, biomass or size; s, survival; sc, species composition; wp, water potential. 7 √, hypothesis supported by trends; digits, % cases in which the hypothesis was supported; x, hypothesis refuted by trends; °, null hypothesis rejected statistically. 2

possibly because competition for nutrients being mineralized near the soil surface becomes more important through time. Both of these examples may reﬂect increases in root competition over time. Variation in biomass allocation can reﬂect variation in competition in ways unrelated to resource capture. Stewart and Thompson (1982) examined allocation in six species of herbaceous perennials in Crataegus scrub, a grassland, and a recently disturbed quarry in the UK. These three sites represented a gradient of increasing disturbance intensity since the grassland was grazed and the quarry was disturbed by mineral extraction. The proportion of individuals ﬂowering increased with disturbance for most species, from about 30 to 90%. Since ﬂowering plants were usually larger than non-ﬂowering plants, and since competition often regulates plant size, the results suggest that competition was least intense in the highly disturbed quarry. Patterns of establishment and replacement in small disturbances are consistent with intense competition beginning early in succession. Survivorship of seedlings in areas disturbed by badgers (Taxidea taxus) in Iowa prairie was 80% in cases where the mound was not previously colonized by plants, but 0% if the mound was previously colonized (Platt and Weis, 1985). High densities of the annual Ambrosia artemisiifolia apparently suppress the perennial Aster pilosus in ﬁrst year Illinois old-ﬁelds (Peterson and Bazzaz, 1978). Patterns in establishment and death with respect to

density and time since disturbance can be examined in woody species because dead stems are preserved and can be aged. The chaparral shrub Ceanothus megacarpus in California showed no death during the ﬁrst 3 years following ﬁre, maximum death rates 6– 10 years after ﬁre, and no death after that (Schlesinger and Gill, 1978). Death is presumably caused by intraspeciﬁc competition, since Ceanothus occupied 100% of community cover. Similarly, mortality in a stand of Pinus banksiana in Ontario increased steadily from 20 to 60 years following ﬁre and establishment (Yarranton and Yarranton, 1975). Pinus stands in North Carolina thin naturally after some years of growth, and the time to the start of thinning decreases with increasing tree density (Peet and Christensen, 1987). Thus, follow-through data for woody populations suggest that intraspeciﬁc competition starts when trees are large enough to affect each other, and that its intensity increases with increasing tree size. Woody population data, however, do not evaluate the restriction of growth by herbaceous plants early in succession. An alternative explanation for competitive effects appearing late in forest succession is that ongoing competition occurring since the beginning of succession has a visible effect only after enough time has passed for the most suppressed individuals to die. This possibility might be tested with experiments. Competitive exclusion might alter community heterogeneity during succession (Armesto et al., 1991;

618

Zobel et al., 1993). Species richness and small-scale heterogeneity decreased during the ﬁrst century of succession on abandoned ﬁelds in southern Sweden as individuals of the shrub Juniperus communis expanded in size (Rejm´anek and Ros´en, 1992). As for woody population data, however, this result does not distinguish between increasing competition and constant competition becoming apparent only as time passes. Population-level experiments Population-level removal experiments in old ﬁelds suggest that competition increases over time. The cover of annuals in a newly tilled New York ﬁeld was not affected by perennial removal (7% in control plots vs. 8% with perennials removed), but, in the following year, annuals had signiﬁcantly higher cover in plots from which perennials had been removed (9% vs. 31%; Marks and Mohler, 1985). Armesto and Pickett (1985, 1986) removed the dominant species from a 2-yearold New Jersey ﬁeld (Ambrosia artemisiifolia) and a 15-year-old ﬁeld (Solidago canadensis). Removals had no effect on diversity or cover of other species in the 2-year-old ﬁeld, but removals in the 15-year-old ﬁeld increased species richness and the cover of common annuals (3–10-fold for A. artemisiifolia, Erigeron annuus, Hieracium pratense). Similarly, the removal of perennials from a 3-year-old New Jersey ﬁeld had little effect on annuals (Jackson and Willemsen, 1976), whereas perennial removal from an 18-year-old South Carolina ﬁeld doubled the cover of annuals (Pinder, 1975). The biennial Verbascum thapsus is present in ﬁelds in Michigan one and two years old, but not in 3-year-old ﬁelds; removing interspeciﬁc neighbors from 3-year-old ﬁelds allowed it to germinate, grow and set seed (Gross, 1980). Removal of interspeciﬁc neighbors from plots in Oregon clear-cuts had no effect on the annual Senecio sylvaticus in the ﬁrst year, but signiﬁcantly increased biomass in the second year (Halpern et al., 1997), suggesting that competition increased over time. Increasing competition did not account for the eventual disappearance of S. sylvaticus, however, because its population in the third year was very small in all competition treatments. In a New England salt marsh (Bertness, 1991), pairwise species removals in undisturbed vegetation produced evidence of competition in four out of six cases, whereas the same treatments applied to

Scott D. WILSON

populations invading artiﬁcially cleared patches showed competition in just two out of six cases, suggesting that competition is reduced just after disturbance. Removal experiments in harvested forests in Quebec showed that early-successional herbs suppress woody plants during the ﬁrst three 3 years of succession (Belleﬂeur and P´etillon, 1983), but long-term monitoring in forests in Alabama, Florida, and Georgia has suggested that suppression lasts only for about the ﬁrst ﬁve years (Oppenheimer et al., 1989; Britt et al., 1990, 1991). Herbs in forests in Quebec and British Columbia decrease soil temperature (Belleﬂeur and LaRocque, 1983; Coates et al., 1991), as does moss in Alaska (Van Cleve and Viereck, 1981); this may suppress tree growth, either by reducing nutrient mineralization or by the direct negative effects of low temperature on tree growth. In contrast, herb removal in North Carolina clear-cuts had no positive effect on trees during the two years following harvest (Wilson and Shure, 1993). The effect of removing trees on herbs, however, was negligible in the ﬁrst year, but increased steadily throughout the second year. Taken together, the results suggest that the importance of competition from herbs and trees simply follows their successional patterns: trees are initially suppressed mostly by herbs until the herbs are shaded by trees, after which trees are mostly suppressed by trees (Peet and Christensen, 1987). As noted above (pp. 612–613), population-level removal experiments must be interpreted cautiously because the dependent variable, the size of the target population, differs among successional stages at the start of the experiment, and this difference may determine the outcome of the experiment. Differences in sizes of target populations among successional stages at the start of the experiment can be avoided by establishing experimental mixtures. Comparisons of pure stands and mixtures grown along crossed gradients of fertility and disturbance suggested that competition was reduced by disturbance in one experiment (Campbell and Grime, 1992), but not in a second (Turkington et al., 1993). The discrepancy may have arisen because the range of disturbance intensity was greater in the ﬁrst case, in which neighbors were clipped and the soil disturbed, than in the second, where neighbors were clipped without disturbing the soil. Individual-level experiments A comparison of ﬁve Michigan old-ﬁelds showed that

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

the suppression of transplants of the perennial grass Andropogon gerardi increased with community litter mass (Foster and Gross, 1997). Neighbors facilitated establishment of Daucus carota in a newly tilled ﬁeld in Michigan, but suppressed its establishment in a 3-yearold ﬁeld (Holt, 1972). Seeds of four biennial species established in a 1-year-old Michigan ﬁeld, but only two of the species could establish in a 15-year-old ﬁeld (Gross and Werner, 1982). The abilities of four early succession annuals to suppress individual annuals in a greenhouse experiment increased strongly with species mass and were not related to interspeciﬁc differences in architecture (Tremmel and Bazzaz, 1995), conﬁrming Goldberg’s (1996) prediction that competitive effects may be primarily a function of neighbor mass more than physiology or morphology. Thus, competition intensity should increase as community mass increases through succession. Competition intensity, calculated as the proportional decrease in individual performance caused by neighbors, increased with community mass as disturbance intensity decreased along an Ontario lakeshore (Wilson and Keddy, 1986a). Community mass is not always a reliable predictor of competition intensity. Annual vegetation in newly tilled plots in a 34-year-old Minnesota old-ﬁeld had higher peak biomass than perennial vegetation in undisturbed plots, but the growth of transplants of the perennial grass Schizachyrium scoparium was reduced by neighbors about 30% in tilled plots and 50% in undisturbed plots (Wilson and Tilman, 1993). Similar results were obtained when eight annual and perennial species were transplanted: competition intensity was signiﬁcantly lower in tilled plots (Wilson and Tilman, 1995). Competition intensity was lower in tilled plots, in spite of their higher peak biomass, probably because of the absence of neighbors at the start of the growing season. There are several examples of competition not increasing with time since disturbance. Neighbors suppressed two species of prairie grasses to about the same extent in burned and unburned plots in Manitoba (Wilson and Shay, 1990), presumably because most of the biomass and meristems in grasslands are belowground and unaffected by ﬁres. Burton and Bazzaz (1991) sowed tree seeds in Illinois old-ﬁelds dominated by annuals, perennials or early-successional trees, as well as in an unvegetated plot. For a minority of cases (sown species by year combinations), increasing competition intensity was suggested by high establishment in the

619

unvegetated plot and decreasing establishment with increasing successional stage. Similarly, the predawn water potential in the xylem of established plants decreased (became more negative) with increasing successional age, and the net photosynthesis per unit leaf area of the established plants was lowest in the presence of Prunus serotina, which represented the oldest successional stage examined (Burton and Bazzaz, 1995). Survival and stem mass, however, showed no obvious relationship to successional stage (Burton and Bazzaz, 1995). Lastly, the inﬂuence of neighbors on tree-seedling survivorship in newly disturbed and 10-year-old North Carolina ﬁelds did not vary signiﬁcantly with vegetation age, although, in the case of one species, Pinus taeda, neighbors tended (P < 0.1) to decrease survivorship more in the older plots (De Steven, 1991b). Negative neighbor effects are not always caused by competition. In New York in old-ﬁelds from one to 15 years since abandonment, neighbors increased predation on tree seeds, apparently by sheltering mice, but the importance of predation did not vary with ﬁeld age (Gill and Marks, 1991). Several studies in a variety of ﬁeld ages have examined the effect of neighbors on transplant growth, but only a subset of studies have included more than one successional stage. In order to synthesize these, I summarized the results of several studies which gave the growth rate of grass transplants with and without herbaceous neighbors. I included some studies reporting initial and ﬁnal transplant mass (Wilson, 1994; Gerry and Wilson, 1995), and thus enabling growth rates to be calculated. The studies were performed in ﬁelds undergoing secondary succession, and in apparently undisturbed grasslands. I considered studies in which both neighbor roots and shoots were removed, and for which ﬁeld age could be determined either from the paper or by communication with the authors (Wilson and Tilman, 1991, 1993, 1995; Wilson, 1993a, 1994; Reader et al., 1994; Gerry and Wilson, 1995; Bakker, 1996; Foster and Gross, 1997; Wilson, unpublished data; Peltzer and Wilson, unpublished data). I examined grasses with herbaceous neighbors in order to avoid drastic differences between transplant and neighbor morphology, which would have occurred if trees had been competing with grasses. I excluded two young sites from Reader et al. (1994) because the topsoil had been removed and the plots were undergoing primary succession (H. Olff, pers. commun., 1996). Fertilized plots (e.g., Wilson and Tilman,

620

Scott D. WILSON

disturbed grasslands (<100 yr; Fig. 26.1). This was true for the study by Reader et al. (1994) of a single transplant species, Poa pratensis (t-test: t = 4.3, P < 0.001), and for all the other studies considered together (t = 2.4, P < 0.05). These results emerged despite differences in transplant species, soils, climate, and neighboring vegetation, suggesting that ﬁeld age is important in controlling competition intensity. Experiments in old-ﬁelds at the Cedar Creek Natural History Area in Minnesota conﬁrmed this: competition intensity increased signiﬁcantly with time since disturbance (Fig. 26.1: regression: r 2 = 0.55, P < 0.01). Root and shoot competition

Fig. 26.1. Competition intensity as a function of ﬁeld age. Data are for perennial grasses transplanted into herbaceous neighborhoods. Competition was calculated as the difference between transplant growth rates without and with neighbors, divided by growth rates without neighbors. Some points have been offset left or right to prevent overlap. The line is a regression relationship between competition intensity and ﬁeld age from experiments at Cedar Creek Natural History Area in Minnesota (Wilson and Tilman, 1991, 1993, 1995; Wilson, 1994). Open squares, Wilson and Tilman (1991); large solid squares, Wilson and Tilman (1993); large open triangles, Wilson (1993a); open circles, Reader et al. (1994); small solid triangles, Wilson (1994); small open triangles, Gerry and Wilson (1995); medium solid squares, Wilson and Tilman (1995); crossed squares, Bakker (1996); solid circles, Foster and Gross (1997); plusses, Wilson (unpublished data); small solid squares, Peltzer and Wilson (unpublished data); crosses, Peltzer and Wilson (unpublished data).

1991, 1993, 1995) were not included. I calculated competition intensity [(growth with no neighbors – growth with neighbors)/growth with no neighbors] for each species, year and vegetation type. Field age was given a value of >100 years if the ﬁeld appeared not to have been cultivated within memory, but the soil in most of the very old ﬁelds had probably never been disturbed. These old sites included nutrient-poor glacial till and an alvar in southern Ontario (Reader et al., 1994), native grassland at Canberra (Reader et al., 1994), Saskatchewan prairie (Wilson, unpublished data; Peltzer and Wilson, unpublished data) and Australian alpine grassland (Wilson, 1993a). Competition intensity was higher in very old (>100 yr) or undisturbed grasslands than in more recently

Field experiments suggest that the relative importance of root competition is greatest on nutrient-poor soils (Putz and Canham, 1992; Wilson, 1993a,b). On poor soils, competition is probably among roots for most of the successional sere (Lauenroth and Cofﬁn, 1992). Tilling a nutrient-poor old-ﬁeld in Minnesota significantly reduced root competition but had no effect on shoot competition (Wilson and Tilman, 1993). Tilling fertilized plots, however, had no effect on root

Fig. 26.2. Possible trajectories of intensities of plant interactions during primary succession, secondary succession on poor soil, and secondary succession on rich soil. F, facilitation; SC, shoot competition; RC, root competition.

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

competition, but decreased shoot competition. Thus, succession on nutrient-poor soils may be accompanied by increasing root competition, as soil space is slowly ﬁlled by perennial roots. Succession on nutrientrich soils may be accompanied by increasing shoot competition (Fig. 26.2; see also Tilman, 1985). Shoot competition might reach maximum values more rapidly than root competition (Fig. 26.2) because the directional nature of light means that it is more easily preempted than are soil resources, which can be renewed in three dimensions. Also, plants might ﬁll the above-ground environment more rapidly than the soil. Evidence that root competition increases slowly through succession comes from grasslands, where competition is mostly among roots (Lauenroth and Cofﬁn, 1992; Wilson, 1993a,b; Gerry and Wilson, 1995). The relationship between competition intensity and ﬁeld age in grasslands (Fig. 26.1) suggests that root competition increases continually for more than 100 years. In contrast, leaf area index in a temperate hardwood forest reached its maximum value after just three years (Aber, 1979). Summary of competition during secondary succession Given the high growth rates and densities of earlysuccessional plants, competition may occur within days following disturbance. Competition usually increases in the course of succession (Table 26.1) because plant mass increases while available resources decline or remain constant.

ARE LATE-SUCCESSIONAL SPECIES BETTER COMPETITORS?

A widely held assumption is that differences in competitive abilities contribute to successional replacements (Drury and Nisbet, 1973; Connell and Slatyer, 1977; Van Hulst, 1978; White, 1979; Finegan, 1984; Grime, 1987; Tilman, 1988, 1994), but testing this assumption has been complicated by the fact that competitive ability has many facets (Tilman, 1985; Huston and Smith, 1987; Berendse, 1994; Goldberg, 1996). One approach is to recognize that competitive ability has two components: effect and response (Goldberg, 1990). Competitive effect refers to the extent to which a plant suppresses its neighbors. If competition

621

intensity increases with time, it is because competitive effects increase. This could result either because latesuccessional species simply have more biomass than early species, or because they have greater competitive effects per unit biomass (Goldberg, 1990). Latesuccessional species clearly have larger effects on light per unit biomass because of their greater stature and perennial nature (Botkin, 1981). Perennials may also have greater effects on soil resources per unit biomass by pre-empting annuals early in the spring (Peterson and Bazzaz, 1978). It is likely that both increasing community biomass and increased competitive effects per unit biomass contribute to the overall increase in competitive effects through time. Competitive response refers to the ability of a plant to grow in the presence of neighbors (Goldberg, 1990). Early-successional species might be poor response competitors because they allocate a large proportion of their biomass to dispersal instead of to resource-gathering organs that would help them compete against other plants (Hancock and Pritts, 1987; Tilman, 1988). Late-successional species might be good response competitors because they allocate more biomass to resource-collecting tissues (Tilman, 1988) and use nutrients efﬁciently (Reiners, 1981; Aerts, 1995). Competitive effects and responses can be separated in some experiments but not in others. For example, growing transplants of different species with and without neighbors in different neighborhoods allows calculation of the competitive effects of the different neighborhoods, and competitive responses of the different transplant species. In contrast, replacement series experiments, in which a set of species are grown in all possible pairwise combinations, produce measures of performance which are inﬂuenced by both effects and responses. In a replacement series, a species is deﬁned as a strong effect competitor if it greatly suppresses its neighbors; these suppressed neighbors are, by deﬁnition, poor response competitors. Thus, competitive effects and responses measured in replacement series are not independent (e.g., Wilson and Keddy, 1986b). Pattern analysis Seeds of the annual Ambrosia triﬁda from a 15-yearold ﬁeld in Illinois produced larger plants than those from an adjacent, continuously tilled ﬁeld, suggesting that competition had selected for higher competitive

622

ability in the older ﬁeld (Hartnett et al., 1987). Allocation of biomass to roots increased and allocation to reproductive organs decreased over the ﬁrst 60 years of succession in nutrient-poor Minnesota old-ﬁelds, suggesting that an ability to compete for soil resources increased over time (Gleeson and Tilman, 1990). Superior competitive abilities of later-successional perennial grasses were also suggested by a survey of ten New Jersey old-ﬁelds representing a 31-year chronosequence (Myster and Pickett, 1992). Late-successional grasses tended to show more negative correlations with other species than did early-successional species, suggesting that they were frequently successful at excluding other species. Neither competitive effects nor responses can be inferred solely from ﬁgures for species abundance in nature, since these could result from factors unrelated to competition. Population-level experiments Many experiments where early- and late-successional species are grown in mixtures show either no differences in their competitive abilities, or even higher competitive abilities in early-successional species. Additive experiments by Goldberg and Landa (1991) involving annual and perennial old-ﬁeld species grown together for ﬁve weeks suggested that competitive effect increased with plant size. Differences among species per unit biomass were small, and annuals and perennials did not fall out as distinct groups with regard to competitive ability. Similar results were obtained from mixtures grown along gradients of fertility and disturbance (Campbell and Grime, 1992). In mixtures of early-, mid-, and late-successional species (Parrish and Bazzaz, 1982), late-successional species were most likely to be suppressed, especially where nutrients were abundant. Early-successional species germinated earlier and grew faster. Similarly, during a four month experiment in an Oregon garden (Shainsky and Radosevich, 1992), the tree Pseudotsuga menziesii was suppressed by increasing density of the earliersuccessional tree Alnus rubra, whereas Pseudotsuga had no effect on Alnus. Other mixture experiments have shown that latesuccessional species or populations are better competitors. A biotype of Taraxacum ofﬁcinale from undisturbed habitats in Michigan was able to suppress a biotype from a disturbed habitat (Solbrig and Simpson, 1974). Later-successional grasses were superior competitors in a replacement series involving moorland

Scott D. WILSON

species in a pot experiment in the United Kingdom (Hester et al., 1991). Late-successional grasses displaced early-successional grasses in mixtures during a three-year experiment in Minnesota because they had higher allocation of biomass to roots (Tilman and Wedin, 1991a,b; Wedin and Tilman, 1993). Grasses transplanted from a 49-year-old ﬁeld into a garden were more successful in invading stands of neighbors than were grasses grown from commercial seed mixtures (Turkington, 1994). Similar results have been obtained from removal experiments. In a New England salt marsh, an earlysuccessional grass never suppressed later grasses, but the later grasses suppressed the early species (Bertness, 1991). Removal experiments in a Quebec forest clearcut suggested that the early-successional tree Betula alleghaniensis was suppressed by neighbors, whereas two later-successional trees, Acer saccharum and Fagus grandifolia, were not (Belleﬂeur and Villeneuve, 1984). Removal of litter from South Carolina old-ﬁelds produced a 5-fold to 10-fold increase in the density of annual plants but no change in perennials (Monk and Gabrielson, 1985), suggesting that the annuals were most suppressed. As noted above (pp. 612–613), earlysuccessional species may show the greatest release following the removal of late-successional neighbors simply because the early-successional species were in low abundance. Different environments can of course produce different results. In the case of Taraxacum ofﬁcinale biotypes, for example, a four-year garden experiment including two levels of disturbance showed that each biotype had superior competitive ability in its native disturbance regime (Solbrig and Simpson, 1977). Individual-level experiments Competitive responses Many individual-level experiments suggest that latesuccessional species are good response competitors. Two early-successional biennials in Michigan established in a 1-year-old ﬁeld but not in a 15-year-old ﬁeld, whereas two later-successional biennials were able to establish in both ﬁelds (Gross and Werner, 1982). Similarly, late-successional species were more successful than early-successional species in establishing in experimental stands of perennial grasses in England and in Michigan (Fenner, 1978; Gross, 1984). An experiment on transplants grown with and without neighbors in a Saskatchewan old-ﬁeld dominated

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

by the introduced perennial grass Bromus inermis suggested that native prairie plants were less sensitive to neighbor biomass than were introduced agricultural species (Gerry and Wilson, 1995). The slopes for the relationships between transplant growth rate and neighbor biomass for native species were between −0.08 and −0.17, whereas the slopes for the agricultural species Agropyron cristatum and Melilotus ofﬁcinalis were between −0.24 and −0.34, indicating that the natives were better response competitors. In contrast, in a similar experiment with eight species in Minnesota, the extent to which transplant growth was suppressed by neighbors did not vary between annual and perennial species (Wilson and Tilman, 1995). Late-successional species may also be better response competitors when light is the limiting resource. Photosynthetic response curves of 14 species from early, mid, and late succession showed that latersuccessional species were much less sensitive to light attenuation than were early species (Bazzaz and Carlson, 1982). This was particularly true for woody plants. Similarly, photosynthetic rate in the late-successional tree Acer saccharum was reduced by neighbors to 43% of that without neighbors, whereas rates for the earlysuccessional trees Gleditsia triacanthos and Prunus serotina were reduced to 19–25% (Burton and Bazzaz, 1995). Together, the experiments suggest that late-successional species are better response competitors in many cases, but that variation in species responses in different environments and neighborhoods make generalizations difﬁcult. Competitive effects The competitive effects of large species from undisturbed wetlands were greater than those of small plants from disturbed habitats (Gaudet and Keddy, 1988). Similar results were obtained from a replacement series experiment (Wilson and Keddy, 1986b). Competitive effects can be compared between earlyand late-successional species by comparing the extent to which they suppress transplant growth. The regression relationship between the growth rate of transplants of the perennial grass Schizachyrium scoparium and neighbor biomass in Minnesota was stronger in undisturbed plots dominated by perennial grasses (r 2 = 0.90) than in disturbed plots dominated by annuals (r 2 = 0.62), suggesting that perennial neighbors accounted for more of the variation in transplant growth than did annuals (Wilson and Tilman, 1993).

623

Further, the slope of the line describing the decrease in transplant growth rate [ln(g g−1 ) d−1 ] with increasing neighbor biomass (g m−2 ) was steeper with perennial neighbors (−0.0041) than with annual neighbors (−0.0026), suggesting that the proportional reduction of transplant growth by neighbors was greater for perennials than annuals. Slopes also declined steadily from low-nitrogen plots dominated by later perennial species to high-nitrogen plots dominated by earlier perennials (Wilson and Tilman, 1991). These results suggest that late-successional species have greater competitive effects than species earlier in the successional sequence. There is also evidence that early-successional species can have large competitive effects on others, especially early in succession. The perennial Aster pilosus was suppressed by a variety of annual neighbors more than it was by conspeciﬁcs when grown in pots (Peterson and Bazzaz, 1978), suggesting that the annuals had larger competitive effects. A comparison of annual and perennial grasses found that the annuals always had higher growth rates (Garnier, 1992), which probably enables them to exert large competitive effects in the earliest stages of succession. The suppression of transplants caused by annual neighbors increased with neighbor biomass in several trials in Minnesota oldﬁelds, whereas suppression never increased with the biomass of perennial neighbors (ﬁg. 5 in Wilson and Tilman, 1995). The difference may arise because the population of long-lived perennial neighbors is near the carrying capacity of the site, and they exert competitive effects regardless of their biomass. In contrast, the population size and thus competitive effects of annuals may be patchy and unpredictable, but sites with higher densities will have more biomass and the annuals there will exert stronger competitive effects. Plant species which are successful invaders tend to be early-successional species that apparently have strong competitive effects relative to species native to the invaded communities. Bakker (1996) measured the competitive effects of the introduced perennial grass Agropyron cristatum, which establishes easily in newly abandoned ﬁelds in the northern Great Plains of North America and continues to exclude native grasses for many decades (Looman and Heinrichs, 1973; Christian, 1996). Grass transplants were grown with and without neighbors in Saskatchewan oldﬁelds about 60 years old dominated either by the introduced grass A. cristatum or by native grasses. The competition term in analysis of variance was

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Scott D. WILSON

Table 26.2 Summary of studies that address the hypothesis that late-successional species have greater competitive abilities than early-successional ones Range 1

Levels 2

N3

System 4

Targets 5

Var 6

Gleeson and Tilman (1990)

1–60

28

1–2

OF

c

al

Myster and Pickett (1992)

1–31

f

10

OF

c

sc

Belleﬂeur and Villeneuve (1984)

1–4

f

30

F

3w

r

Bertness (1991)

0,Und

2

10

W

3p

r

Campbell and Grime (1992)

0

5

3

G

2a,4p

r

x

Goldberg and Landa (1991)

0

1

7

G

2a,5p

r

Gross and Werner (1982)

1–15

3

20

OF

4a

g

Hester et al. (1991)

0–2

f

5

G

2g,1w

r

x E, √ R √

Monk and Gabrielson (1985)

3,15

2

2

OF

c

r

Parrish and Bazzaz (1982)

0

1

6–12

G

6a,11p

s

x

Shainsky and Radosevich (1992)

0

1

3

G

2w

r

Solbrig and Simpson (1974)

0

1

1–2

G

2p

r

x √

Solbrig and Simpson (1977)

0–4

f

2–3

G

2p

r

Tilman and Wedin (1991a,b)

0–2

f

2–4

G

5p

al

Bazzaz and Carlson (1982)

0

1

4

OF,F

4a,2p,8w

ph

Burton and Bazzaz (1995)

0–19

9

1

OF

4w

ph

Fenner (1978)

0

1

11

G

4a,2p

g

Gaudet and Keddy (1988)

0

1

5

W

1a,43p

r

Gerry and Wilson (1995)

~30

1

10

P

4p,2w

r

Reference Pattern analysis

Population-level experiments

Individual-level experiments

Peterson and Bazzaz (1978)

0

1

10

G

1a,1p

r

Wilson and Keddy (1986b)

0

1

10

W

7p

r

Wilson and Tilman (1993)

0,38

2

10

OF

1p

r

Wilson and Tilman (1995)

0,38

2

10

OF

4a,4p

r

Results 7 √ √

√ √

66 √

R

√

x √

√ √ √ √ √

R R R R

x° √ √ √

°E °E, x °R

1

Community ages (years); 0, newly disturbed ground; Und, undisturbed ground. Number of successional stages considered; f, follow-through study. 3 Number of replicates of experimental treatments. 4 F, forest; G, greenhouse or garden; P, permanent grassland; OF, old-ﬁeld; W, wetland. 5 Number and life histories of study species; a, annuals or biennials; c, community as a whole; p, herbaceous perennials; w, woody plants. 6 Variable measured: al, allocation; g, germination, dispersal or establishment; ph, photosynthetic rate; r, growth, mass or size; s, survival; sc, species composition. 7 √, hypothesis supported by trends; digits, % cases in which the hypothesis was supported; x, hypothesis refuted by trends; °, null hypothesis rejected statistically; E, competitive effects; R, competitive responses. 2

greater in the ﬁeld dominated by A. cristatum than in that dominated by natives, suggesting that competition was more important in areas dominated by A. cristatum (Underwood and Petraitis, 1993). Competitive effects may be intense if a fast-growing species can quickly colonize a site and maintain dominance.

Summary of competitive ability and secondary succession Only in some cases does competitive ability increase with successional stage (Table 26.2) because different kinds of competitive ability are required for success at different successional stages and in

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

different environments (Grubb, 1985; Huston and Smith, 1987). Experiments performed in the relevant environments are, of course, the most informative. Among early-successional species rapidly ﬁlling newly available space, competitive effects are strong (Goldberg and Landa, 1991), and the rapid thinning evidenced by skewed size distributions suggests that early-successional species are generally poor response competitors (Weiner, 1990). Among later-successional species, competitive effects are strong on account both of their larger size (Goldberg and Landa, 1991) and their perennial nature, which allows them to pre-empt resources (Bazzaz, 1990). Late-successional species are good response competitors, which allows them to persist until early-successional species disappear.

DOES COMPETITION CAUSE NICHE DIFFERENTIATION?

The range of resources and conditions encountered on recently disturbed ground is likely to be greater than that on undisturbed ground, so early-successional species should be able to succeed in a wider range of environments than late-successional plants (Bazzaz, 1983, 1986, 1987). Thus, by comparison, latesuccessional species should have narrower and more differentiated niches than early-successional species. This could arise from competition over the course of succession, or simply from the need to cope with a greater range of conditions early in succession. If competition during succession is responsible for determining which genotypes persist, then surviving genotypes should be complementary in their abilities to coexist (Turkington and Mehrhoff, 1990). Competition might also select for complementarity among species in their competitive abilities (Aarssen, 1983, 1985). Pattern analysis The niches of several species of Solidago showed more overlap in Michigan old-ﬁelds than they did in Iowa prairie, probably because of the stochastic nature of establishment and pre-emption in the oldﬁelds (Werner and Platt, 1976). Multivariate analyses of North Carolina piedmont forests suggested that betadiversity, and thus niche differentiation, was greatest in the oldest forest (Christensen and Peet, 1981, 1984).

625

The proportion of cover contributed by different guilds in New Zealand tussock grasslands did not vary over 27 years since ﬁre, which does not support niche differentiation (Gitay and Wilson, 1995). Lepˇs (1991) pointed out that niches might appear to diverge if sampling is done across an environmental gradient early in succession but along the gradient later in succession. For example, an environmental gradient that was not apparent might be sampled at just a few points early in succession, and species would not seem to be differentiated along it. The same gradient might be more apparent later in succession and might be more carefully sampled, so that species would appear to be more differentiated. Examination of species interrelations in 31-yearold ﬁelds in New Jersey showed that only 2% of interactions involving introduced species were negative, whereas 6% of those involving native species were negative (Myster and Pickett, 1992). This could reﬂect a stronger niche differentiation among native species, which have presumably interacted for a longer time than introduced species. Niche divergence might also account for an increase in the proportion of random associations among species over the 31 years surveyed. Similar results emerged from a survey of 14 Ontario quarries that had been abandoned for periods of 1 to 21 years, which O’Connor and Aarssen (1987) interpreted as evidence that complementarity in competitive ability increases over time. Population-level experiments Populations of early-successional herbs tend to have wider niches along experimental gradients than populations of later species (Parrish and Bazzaz, 1982; Bazzaz, 1987). Population yield was usually reduced by interspeciﬁc neighbors in replacement series of annual species, but was increased by interspeciﬁc neighbors in 10 out of 20 mixtures of native perennial species, suggesting that late-successional species had more niche differentiation (Parrish and Bazzaz, 1982). Most of the existing knowledge about differences in competitive interactions associated with different ages of vegetation comes from comparisons of pastures of different ages by Turkington and his associates. Genets of perennials from a 40-year-old ﬁeld in British Columbia, studied in replacement series, grew best when grown with genets from their original neighborhoods, suggesting that specialization had occurred

626

(Aarssen and Turkington, 1985a). Aarssen (1988) collected seeds from four perennial species in 5-, 24and 43-year-old ﬁelds and grew them in mixtures; the consistency of rank order of competitive ability decreased with increasing ﬁeld age. This might reﬂect an increase in the ability of all genotypes to compete, but it could also result from niche divergence in the oldest ﬁelds decreasing competition intensity and so decreasing the predictability of the outcome of competition. Turkington and Joliffe (1996) grew populations from Welsh pastures from 6 to over 100 years old in replacement series at ﬁve densities; the negative impact of the dominant species, Lolium perenne and Trifolium repens, decreased with increasing pasture age, and the yields of mixes were highest for genotypes from the oldest ﬁelds, suggesting that divergence had occurred. Other evidence for competition-driven niche specialization through succession comes from an experiment which followed four generations of annuals grown along gradients of soil moisture, density, and neighbor diversity (Zangerl and Bazzaz, 1984). Population variation decreased through time along all gradients, suggesting that specialization was occurring. Changes in variation, however, were least predictable in the most diverse treatments, which highlights the importance of plant interactions in controlling population variation through succession. Individual-level experiments Comparison of germination responses of early- and late-successional trees found no decrease in niche width in late-successional species (Burton and Bazzaz, 1991). Genets of the perennial grasses Holcus lanatus and Lolium perenne and the legume Trifolium repens grown singly in pots showed decreasing population variability with increasing age of the pasture from which they had been collected (4–67 years), suggesting specialization over time (Aarssen and Turkington, 1985b). Genets of Trifolium repens from a Welsh pasture more than 100 years old grew better when transplanted back into their original neighborhoods than when planted into different neighborhoods; this effect disappeared when neighbors were removed, suggesting that the important differences among the neighborhoods were biotic (Turkington, 1989).

Scott D. WILSON

Summary of niche divergence during secondary succession Several studies illustrate that late-successional populations tend to be more differentiated than earlysuccessional populations. Attributing increased differentiation to competition, based on patterns alone, is difﬁcult (Connell, 1980; Keddy, 1989). Evidence that competition causes differentiation has been found in experiments which manipulate neighbors in mixtures of contrasting ages (e.g., Zangerl and Bazzaz, 1984; Turkington, 1989). INTERACTIONS DURING PRIMARY VS. SECONDARY SUCCESSION

Plant interactions during primary succession probably differ from those during secondary succession in several respects (Fig. 26.2). First, facilitation can be important in primary succession (Walker and Chapin, 1987), but appears to be relatively unimportant during secondary succession. Second, in primary succession there can be no competition until population densities are high enough to allow interactions, whereas competition begins very quickly during secondary succession. Third, early competition during primary succession is likely to be among roots, because soil resources are in short supply. In contrast, early competition in secondary succession may be among roots on nutrient-poor soils or among shoots on nutrient-rich soils. Fourth, competition during primary succession in relatively productive habitats may be characterized by a shift from root to shoot competition over time, as soil resources, standing crop and shade increase. In contrast, the relative importance of root and shoot competition might not change during secondary succession. CONCLUSIONS

Generalizations about the direction, magnitude and effects of plant interactions during secondary succession can be evaluated with a variety of pattern analyses and population-level and individual-level experiments. Facilitation is often suggested by positive associations among early successional species, but associations are not always positive, and positive associations could be caused by processes other than plant interactions. Experiments comprising several species and

PLANT INTERACTIONS DURING SECONDARY SUCCESSION

treatments often detect facilitation, but always in only a minority of cases (Table 26.1). Facilitation during secondary succession is most likely to occur in cases in which disturbance creates harsh conditions which are ameliorated by neighbors. The intensity of competition is likely to increase during secondary succession, where plant demand increases over time and resources either remain constant or decline. There are several instances of such resource dynamics. Population-level experiments are difﬁcult to interpret, since both the manipulated and measured variables vary through time. Individuallevel experiments show signiﬁcant competition early in succession and suggest that competition increases through time (Table 26.1, Fig. 26.1). Competition is probably primarily among roots throughout succession on nutient-poor soils, and among shoots on richer soils (Fig. 26.2). Because competition is more intense later in succession, late-successional species might be expected to persist or dominate because of superior competitive abilities. Inferring relative competitive abilities from pattern is uncertain because of the multitude of possible causes of any pattern. Experiments frequently ﬁnd few distinctions between the competitive abilities of earlyand late-successional species (Table 26.2). Because competition is a signiﬁcant force in both the youngest and oldest communities, and given the large changes in resources, conditions, and neighbors that occur during succession, it seems clear that different types of competitive ability will be required for success in different environments. This, combined with plant plasticity, makes it unlikely that consistent differences in competitive abilities can be assigned to early- and late-successional species. One way forward would be to measure competitive effects and responses in relevant environments with relevant neighbors over relevant time periods. Narrower niches in late succession could indicate competitive exclusion, or simply smaller ranges of conditions available for growth in undisturbed areas. Experiments that examine neighbors, ﬁeld age, and plant origin indicate that competition-driven niche restriction is possible. This evidence, however, is all from temperate old-ﬁelds and, as is the case for much understanding of succession, there is a need for testing these ideas in more systems. Many studies address hypotheses about plant interactions during secondary succession, but relatively few have designs allowing examination of the statistical

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interactions constituting explicit tests of these hypotheses. Replication of ﬁeld age could be increased in many cases (Table 26.1). Frequently, a ﬁeld of one age is compared with another, and differences between them are attributed to successional stage. It is always possible that other differences between ﬁelds are also important (Pickett, 1989). This could be overcome by studying several ﬁelds of each successional stage. Another solution is to use a regression-style approach and examine ﬁelds of many ages. This is now fairly common for pattern analysis (e.g., O’Connor and Aarssen, 1987; Gleeson and Tilman, 1990; Gitay and Wilson, 1995) but has not been used as much in experiments (cf. Gill and Marks, 1991; Burton and Bazzaz, 1995). Competition experiments performed at several successional stages would be very large, but would allow variation to be partitioned among several factors, including successional stage, neighbor removals, and transplant species, and should allow the relative importance of interactions to be assessed. ACKNOWLEDGMENTS

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Turkington, R., Klein, E. and Chanway, C.P., 1993. Interactive effects of nutrients and disturbance: an experimental test of plant strategy theory. Ecology, 74: 863–878. Underwood, A.J. and Petraitis, P.S., 1993. Structure of intertidal assemblages in different locations: how can local processes be compared. In: R.E. Ricklefs and D. Schluter (Editors), Species Diversity in Ecological Communities. University of Chicago Press, Chicago, pp. 39–51. Van Cleve, K. and Viereck, L.A., 1981. Forest succession in relation to nutrient cycling in the boreal forests of Alaska. In: D.C. West, H.H. Shugart and D.B. Botkin (Editors), Forest Succession: Concepts and Application. Springer-Verlag, New York, pp. 185–211. van Hulst, R., 1978. The dynamics of vegetation: patterns of environmental and vegetation change. Vegetatio, 38: 65–75. Vieira, I.C.G., Uhl, C. and Nepstad, D., 1994. The role of the shrub Cordia multispicata Cham. as a ‘succession facilitator’ in an abandoned pasture, Paragominas, Amazˆonia. Vegetatio, 115: 91–99. Walker, L.R. and Chapin III, F.S., 1987. Relative importance and interactions among processes controlling successional change. Oikos, 50: 131–135. Wedin, D. and Tilman, D., 1993. Competition among grasses along a nitrogen gradient: initial conditions and mechanisms of competition. Ecol. Monogr., 63: 199–229. Weiner, J., 1990. Asymmetric competition in plant populations. Trends Ecol. Evol., 5: 360–364. Wellington, A.B. and Noble, I.R., 1985. Post-ﬁre recruitment and mortality in a population of the mallee Eucalyptus incrassata in semi-arid, southeastern Australia. J. Ecol., 73: 645–656. Werner, P.A. and Harbeck, A.L., 1982. The pattern of tree seedling establishment relative to staghorn sumac cover in Michigan old ﬁelds. Am. Midl. Nat., 108: 124–132. Werner, P.A. and Platt, W.J., 1976. Ecological relationships of cooccurring goldenrods (Solidago: Compositae). Am. Nat., 110: 959–971. White, P.S., 1979. Pattern, process and natural disturbance in vegetation. Bot. Rev., 45: 229–299. Wilson, A.D. and Shure, D.J., 1993. Plant competition and nutrient limitation during early succession in the southern Appalachian Mountains. Am. Midl. Nat., 129: 1–9. Wilson, S.D., 1993a. Competition and resource availability in heath and grassland in the Snowy Mountains of Australia. J. Ecol., 81: 445–451. Wilson, S.D., 1993b. Belowground competition in forest and prairie. Oikos, 68: 146–150. Wilson, S.D., 1994. Initial size and the competitive responses of two grasses at two levels of soil nitrogen: a ﬁeld experiment. Can. J. Bot., 1349–1354. Wilson, S.D. and Keddy, P.A., 1986a. Measuring diffuse competition along an environmental gradient: results from a shoreline plant community. Am. Nat., 127: 862–869. Wilson, S.D. and Keddy, P.A., 1986b. Species competitive ability and position along a natural stress/disturbance gradient. Ecology, 67: 1236–1242. Wilson, S.D. and Shay, J.M., 1990. Competition, ﬁre and nutrients in a mixed-grass prairie. Ecology, 71: 1959–1967. Wilson, S.D. and Tilman, D., 1991. Components of plant competition along an experimental gradient of nitrogen availability. Ecology, 72: 1050–1065.

632 Wilson, S.D. and Tilman, D., 1993. Plant competition in relation to disturbance, fertility and resource availability. Ecology, 74: 599–611. Wilson, S.D. and Tilman, D., 1995. Competitive responses of eight old-ﬁeld plant species in four environments. Ecology, 76: 1169–1180. Yarranton, M. and Yarranton, G.A., 1975. Demography of a jackpine stand. Can. J. Bot., 53: 310–314.

Scott D. WILSON Zangerl, A.R. and Bazzaz, F.A., 1984. Effects of short-term selection along environmental gradients on variation in populations of Amaranthus retroﬂexus and Abutilon theophrasti. Ecology, 65: 207–217. Zobel, K., Zobel, M. and Peet, R.K., 1993. Change in pattern diversity during secondary succession in Estonian forests. J. Vegetation Sci., 4: 489–498.

Chapter 27

THE RESPONSE OF ANIMALS TO DISTURBANCE AND THEIR ROLES IN PATCH GENERATION Michael R. WILLIG and Mark A. McGINLEY

INTRODUCTION

Natural and anthropogenic disturbances play critical roles in molding the structure and function of many terrestrial ecosystems. Because the deﬁnition of disturbance encompasses a variety of meanings, it allows considerable latitude in classifying phenomena as sources of disturbance. In general, disturbance may be considered “any relatively discrete event in time that disrupts ecosystem, community, or population structure and changes resources, substrate availability, or the physical environment” (White and Pickett, 1985). Other notions of disturbance include events that remove a community from its “normal” state, processes that result in mortality or loss of biomass (e.g., Huston, 1994), processes that prevent the attainment of equilibria (e.g., Krebs, 1994), or events that open sites for regeneration (e.g., Harper, 1977). The variety of events that have been considered as disturbances in natural systems is reﬂected in a list compiled by White and Pickett (1985) which includes (1) climatic events such as hurricanes and windstorms, ice storms and freezes, or rainstorms and ﬂash ﬂoods; (2) large-scale geological events such as earthquakes and volcanic eruptions; (3) alterations in climatic conditions such as droughts; and (4) biotic processes such as tree-falls, insect outbreaks, disease, predation, and burrowing or building by animals. We use the deﬁnition of disturbance provided by White and Pickett (1985), but limit its application for both practical and conceptual reasons. First, we exclude the effects of trophic interactions (i.e., positive effects obtained by a consumer and negative effects suffered by the consumed as a consequence of feeding). Although trophic interactions have important effects on the structural and functional attributes of populations,

communities, and ecosystems (see Schowalter and Lowman, Chapter 9, this volume), appropriate conceptual contexts for examining these relationships already exist, and the beneﬁt to the theory of disturbance of including them as examples of disturbances is unclear. Massive outbreaks or epidemics may represent an exception to this disclaimer, but even then, it is difﬁcult to distinguish the population size or impact of the consumer necessary to distinguish cascading trophic effects from catastrophic effects. Second, we do not consider long-term alterations in climate as disturbances. We might consider a freeze to be a disturbance, but we would consider global warming as outside the domain of disturbance theory. The difference between these two kinds of events is not purely semantic. We believe that it is useful to limit the term disturbance to include “relatively discrete events in time” that may recur continually, but to exclude long-term changes or ﬂuctuations that are continuous in nature. Disturbances may be characterized by their frequency (mean number of events per time period), intensity (physical force of the event per unit area per unit time), and extent (area disturbed) within a domain of interest (White and Pickett, 1985). Typically, the domain of interest is an ecosystem within which population, community, or process characteristics are examined. The intensity of a particular event (e.g., landslide) may be great at a local scale, but less when considered at the level of the domain of interest. Elements of disturbance regimes may be visualized as potentially occupying regions of space deﬁned by each of three orthogonal axes (Waide and Lugo, 1992). Some disturbances are rare, high-intensity events that alter broad portions of the landscape

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H

Disturbance regimes H = Hurricane L = Landslide T = Treefall

T

sit en Int

E x t e n t

y

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of populations (e.g., density or distribution) or communities (e.g., richness or composition), rather than dynamics or interactions (e.g., rates of immigration, emigration, mortality, and natality for populations; strengths of competition or predation for communities). By necessity, our exposition reﬂects these limitations. For heuristic and historical purposes, we distinguish three effects of natural disturbances on animals: (1) direct effects; (2) indirect effects derived from an altered physical environment; and (3) indirect effects that result from an altered biological environment (Fig. 27.2). The direct effects are immediate and occur as a result of mortality associated with the disturbance, or the spatial redistribution of individuals occurring during the disturbance. Indirect effects that derive from the altered physical environment become manifest shortly after the initial disturbance, and likely occur because of altered mortality or the behavioral response of mobile species to a currently inhospitable physical environment. In practice, it may be difﬁcult or impossible to distinguish between direct and indirect effects, because of the logistic delays usually associated with sampling an area immediately after a disturbance.

Fig. 27.1. Diagrammatic representation of disturbance regimes as deﬁned by their frequency, intensity, and extent. These three attributes may be correlated in natural systems.

(e.g., hurricanes, ﬁres, ﬂoods), whereas others may be frequent, low-intensity events that affect smaller portions of the landscape (e.g., tree-falls, small animal disturbances). Between these extremes are infrequent, moderately intense disturbances that affect areas of intermediate extent (e.g., landslides). Perusal of the literature suggests that natural disturbances are not homogeneously distributed in this three-dimensional space (Fig. 27.1). Frequency may be inversely related to intensity and extent, whereas intensity and extent may exhibit a positive association. Nonetheless, these associations may be a biased reﬂection of the interests of investigators, as much as an accurate reﬂection of the distribution of disturbance attributes in nature. Most work that has examined the response of terrestrial animals to disturbance has focused on the two extreme types: broad, intense, and infrequent disturbances caused by atmospheric or geological processes, or narrow, less intense, and frequent disturbances with a biotic origin. Similarly, the taxonomic and geographic focus of research has not been homogeneous (e.g., mostly vertebrates). Moreover, the response of animals to disturbance primarily has considered static features

Fig. 27.2. Diagrammatic representation of the relative time periods during which direct, indirect physical, and indirect biological mechanisms associated with disturbance affect animal responses. The absolute times associated with the temporal axis will differ among disturbances (e.g., tree-falls versus ﬁres versus hurricanes).

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This is especially important when considering mobile organisms (i.e., animals) whose positions in geographic space are not static (as in plants). Finally, the changes in population and community attributes of the biota that derive from the ﬁrst two types of effects alter the frequency and strength of both intraspeciﬁc and interspeciﬁc interactions, thereby impinging on the natality, mortality, emigration, and immigration of the remaining species. This is the beginning of a cascading sequence of events, which, combined with the effects of dispersal into the area, results in trajectories of recovery during succession. The interplay of these three determines the response of animal populations and communities at any time after a disturbance.

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In this chapter, we consider two broad categories of disturbance based on the factors responsible for initiating the disturbance. Abiogenic disturbances are those primarily caused by changes in the physical environment (e.g., hurricanes, ﬁres, ﬂoods, earthquakes) as a consequence of climatological or geological processes, whereas biogenic disturbances primarily are initiated by the actions of species (rodent burrows, animal carcasses). The intensity of abiogenic events such as wind storms (e.g., wind velocity) and ﬁres (e.g., temperature) can be measured independently of the consequences of these events (i.e., severity sensu White and Pickett, 1985) on the biota. In most cases, one may expect the severity of a disturbance to be correlated with the intensity of the event that caused it. However, it is inherently difﬁcult to deﬁne the intensity of biogenic disturbances independent of their consequences. We do not intend to provide a comprehensive review of the literature concerning the response of animals to all types of disturbance. Instead, we consider the effects of disturbances on animals by focusing on illustrative examples that include destructive, climatic events such as hurricanes and ﬁres, as well as patchgenerating phenomena such as tree-falls. Moreover, we consider a variety of animal activities as agents of disturbance, and illustrate their effects on plants and animals. Finally, we present a case study of the effects of animal disturbances on the structure and function of a sand shinnery oak (Quercus havardii) ecosystem in western Texas.

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Hurricanes, typhoons and cyclones in tropical forests These infrequent, high-intensity storms affect large regions of the tropics and subtropics. Their impact on the structure and function of ecological systems has been particularly well documented in regions bordering the Caribbean (Walker et al., 1991, 1996; Wiley and Wunderle, 1993). Indeed, they are considered to be among the most important features that mold the biotic complexion of these ecosystems. Direct effects as a consequence of wind and rain (Kennedy, 1970; Walker et al., 1991) cause appreciable mortality in plants, but mortality of animals is rarely documented in the literature (e.g., Waide, 1991a; Will, 1991; Willig and Camilo, 1991). Nonetheless, anecdotal evidence and extrapolation suggest that rain, wind, and ﬂooding can cause mortality to animals during cyclonic wind storms (Wiley and Wunderle, 1993). Storm-induced tree defoliation opens the canopy and alters the microclimatic attributes of the forest with regard to temperature and humidity. Similarly, the felling of trees and massive input of debris (leaf litter, branch falls) caused by hurricanes alter the distribution of biomass, as well as the vertical and horizontal structure of the forest. Finally, destruction of food supplies (e.g., fruits and ﬂowers) and refugia (e.g., nest or roosting sites) can enhance post-hurricane mortality for an extended period after the initial impact (Jeggo and Taynton, 1980). Because of severe structural damage and attendant human misery (Safﬁr, 1991; Sparks, 1991), direct effects of tropical storms on populations or communities of animals are rarely documented, and it is quite difﬁcult to distinguish them from indirect effects subsequently caused by alterations in the physical environment or by modiﬁcations in biotic interactions. As a consequence, we use the terminology of Waide (1991b) to categorize the initial responses of animals to tropical storms (immediate, 1–3 months; shortterm, 4–6 months; mid-term, 7–18 months), with responses beyond 18 months considered to be longterm consequences. Immediate responses Effects of Hurricane Hugo (1989) on animal populations within 3 months of impact were quite variable in Puerto Rico, depending on taxon. For example, despite reductions in density associated with the direct effects of Hurricane Hugo, total spider densities were substantially higher 3 months after the hurricane

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(90 833 individuals ha−1 ) compared to before the hurricane (Pfeiffer, 1996). Most of this increase in density was attributed to the orb-weaver, Leucauge regnyi. Nonetheless, responses to Hurricane Hugo were species-speciﬁc, and some taxa (e.g., the pholcid, Modisimus signatus) decreased in abundance. Density responses were related to habitat use; taxa that attach webs to undersides and margins of live leaves decreased in abundance, whereas species that built webs in furled dead leaves suspended from fallen branches and twigs increased in abundance (Pfeiffer, 1996). Although detailed quantitative studies of many insect species are not available prior to Hurricane Hugo (Garrison and Willig, 1996), Torres (1992) reported conspicuous increases in the abundance of a number of insects in the Luquillo Forest within 90 days of impact by Hurricane Hugo. Diptera, especially fruit ﬂies (Drosophilidae), increased within one month of the hurricane’s impact. This increase was linked to inputs of decaying fruit on the forest ﬂoor as a consequence of wind and rain associated with Hurricane Hugo. Similarly, bark beetles (Scolytidae) and pin-hole borers (Platypodidae) increased in density, as a consequence of increased quantities of decaying tree trunks and branches. Among birds, frugivores (Columba squamosa, Euphonia musica, Geotrygon montana) and nectarivores (Anthracothorax viridis, Chlorostilbon maugaeus, Coereba ﬂaveola) clearly decreased as a consequence of Hurricane Hugo, whereas insectivores (e.g., Todus mexicanus) and omnivores (e.g., Margarops fuscutus, Melanerpes portoricensis, Neospingus speculiferus, Turdus plumbeus) increased in abundance (Waide, 1991a, 1996; Wunderle, 1995). Although lizards (Anolis spp.) did not exhibit differences in density three months after hurricane impact, their spatial (vertical) distribution was altered signiﬁcantly (Reagan, 1991, 1996). Moreover, susceptibility to the effects of hurricanes may be age-dependent; for example, the density of adult frogs (Eleutherodactylus coqui) was unaffected, while the density of juveniles decreased in the aftermath of Hurricane Hugo (Woolbright, 1991, 1996; Stewart and Woolbright, 1996). Like the Caribbean Basin, islands in the Indian and Paciﬁc Oceans are subject to cyclonic storms and typhoons at relatively frequent intervals (Bani, 1992; Craig and Syron, 1992; Robertson, 1992; Elmqvist et al., 1994). In the aftermath of cyclones, substantial reductions in densities of ﬂying foxes (Pterepodidae) have been reported at a number of sites, including Rodrigues Island (Carroll, 1984), Mauritius (Cheke

Michael R. WILLIG and Mark A. McGINLEY

and Dahl, 1981), Guam (Wiles, 1987) and Samoa (Craig and Syron, 1992; Craig et al., 1994). Speculation suggests that some individuals are killed directly or blown far out to sea (Craig and Syron, 1992). Most data indicate that death derives from reduced health or vigor caused by diminished food supplies and increased exposure to rain, wind, and heat, as well as from enhanced predation by domestic animals (e.g., cats, dogs, and pigs) and human hunters (Stinson et al., 1992). These same indirect effects may induce bat movements to areas less devastated by the disturbance; because pteropodids are strong ﬂiers, this includes other islands as well as protected habitats within islands (Pierson et al., 1996). Short-term responses The heterogeneity of response to hurricanes becomes greater when one focuses on the short term (4– 6 months), and expands the geographic basis for comparison. In Puerto Rico, frugivorous and nectarivorous birds still exhibited little recovery from Hurricane Hugo, whereas insectivorous and omnivorous birds had returned to pre-hurricane densities (Waide, 1991a; Wunderle, 1995). In sharp contrast, Yih et al. (1989) reported the virtual absence of birds in Nicaraguan lowland forest 4 months after the impact of Hurricane Joan. In response to Hurricane Gilbert in Jamaica, frugivorous and nectarivorous birds decreased in montane areas, but increased in the lowlands; no consistent responses in density were observed with regard to insectivorous and omnivorous species of birds (Wunderle et al., 1992). On the Yucatan Peninsula of Mexico, frugivorous, nectarivorous, insectivorous, and omnivorous birds had not yet, in 1990, recovered to levels prior to the impact of Hurricane Gilbert; in fact, many species of frugivores and nectarivores were rare or absent (Lynch, 1991). Although Puerto Rican lizards exceeded pre-hurricane levels within 4–6 months of disturbance (Reagan, 1991), frogs continued to exhibit the same age-speciﬁc response (see above) as detected in earlier surveys (Woolbright, 1991) – namely, that juveniles were affected but not the adults. A detailed study of two species of tree-roosting ﬂying foxes (Pteropus samoensis and P. tonganus) on islands in the South Paciﬁc before and after two consecutive cyclones (Ofa in 1990 and Val in 1991) provides considerable insight into the way in which large frugivorous bats respond to disturbance (Pierson et al., 1996). Prior to the impact of Cyclone Val in the Tafua Rain Forest Reserve (~5000 ha), average

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density estimates of P. tonganus indicated about 1400 individuals in the reserve. Four months later, density estimates averaged only four individuals. In addition to such marked reductions (>99%) in density, P. tonganus became more diurnal and was observed feeding alone rather than in groups, often in areas of human habitation or dominated by agriculture. Similarly, in Falealupo Rain Forest Reserve (also ~5000 ha), a colony of several hundred P. tonganus was reduced to no more than ﬁve individuals within four months of Cyclone Ofa. In contrast, P. samoensis persisted in precyclone roost sites, albeit in somewhat lower numbers, and continued to forage diurnally within forested areas. In response to the reduction in native fruits 4 months after Cyclone Ofa, P. samoensis consumed petioles of an unidentiﬁed vine that proliferated on defoliated trees, leaves of an epiphytic orchid (Eria robusta), and bracts of a storm-resistant liana (Freycinetia reineckeri) that survived the storm. Facultative folivory on the part of this species buffers it from the otherwise devastating effects of resource depletion, and is an important survival strategy in response to altered landscapes and tree phenology (Pierson et al., 1996). Mid-term responses Outbreaks of many species of Lepidoptera occurred in the aftermath of Hurricane Hugo (Torres, 1992). Peak lepidopteran densities, 7–9 months after the hurricane, were associated with the invasion of early-successional plants and increased proportional abundance of herbaceous plants and vines. After 12–18 months, vertebrates (birds, lizards, and frogs) differed from invertebrates in the degree to which they had recovered from the impact of Hurricane Hugo in Puerto Rico. Most of the populations of vertebrates (birds, lizards, juvenile frogs) partially or fully recovered, and some (adult frogs) appreciably exceeded pre-hurricane levels. In contrast, walking sticks or phasmids (e.g., Agamemnon iphimedeia, Lamponius portoricensis), snails (Caracolus caracolla, Nenia tridens, Polydontes acutangula), and a slug (Gaeotis nigrolineata) showed severe reductions in density and altered spatial distributions one year after Hurricane Hugo (Willig and Camilo, 1991). Nonetheless, the hurricane did not affect the sizedistribution of C. caracolla to an appreciable extent; hurricane effects were independent of size or age for this common snail. Similarly, the response of bats to Hurricane Hugo differed in a species-speciﬁc fashion (Gannon and

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Willig, 1994; Willig and Gannon, 1996). One year after the hurricane, Artibeus jamaicensis, a frugivore, showed a severe reduction in density; Stenoderma rufum, another frugivore, exhibited moderately reduced density; Monophyllus redmani, a nectarivore, exhibited a slight increase in density. More detailed demographic data for S. rufum suggested a 50% reduction in the proportion of reproductively active females, as well as a 60% reduction in the proportion of juveniles in the population. In addition, the home range and foraging range of S. rufum were smaller one year after the hurricane, compared to a two-year-period before it. After 11 months of recovery from Cyclone Ofa, mean population density of Pteropus tonganus was only 2% of pre-cyclone numbers in Tafua Rain Forest Reserve (Pierson et al., 1996). In the Falealupo Rain Forest Reserve, bat numbers increased to less than 10% of pre-cyclone numbers (Pierson et al., 1996). High juvenile mortality after the storm (virtually 100%), combined with slow recovery by a highly damaged, preferred food source (Syzygium inophylloides), likely contributed to the slow recovery of the bat population. Long-term responses Documentation of the long-term responses of animals to hurricanes is rare. The most extensive and intensive data exist for snails, slugs, frogs, and bats in the Luquillo Experimental Forest of Puerto Rico. These studies form the basis for the exposition that follows. Although initially their densities were much reduced after the hurricane, the most common species of snails (Caracolus caracolla, Nenia tridens, Polydontes acutangula) and slugs (Gaeotis nigrolineata), greatly exceeded their pre-hurricane densities (two- to sevenfold increases) as a consequence of ﬁve years of secondary succession in areas drastically affected by Hurricane Hugo (Bisley Watersheds) (Secrest et al., 1996; Willig et al., 1998). Large inputs of litter after the hurricane, followed by rapid development of the understory and its subsequent diminution as a consequence of canopy closure, provided increased quantities of food and substrate for these species. In areas less drastically affected by Hurricane Hugo (El Verde), pre-hurricane surveys of gastropods were not conducted, but an extensive monitoring of attributes at population and community level provides the necessary data to assess long-term responses to hurricanes (Secrest et al., 1996; Willig et al., 1998). Demographic responses (1991 to 1995) to disturbance differed among the more common species. For example, two snail

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species (Alcadia striata and Cepolis squamosa) showed asymptotic declines in density, two species (Nenia tridens and Platysuccinea portoricensis) exhibited increases in density, one species (P. acutangula) declined for 2.5 years and then increased for 2.5 years, and one slug species (Gaeotis nigrolineata) increased for 2.5 years and then decreased for 2.5 years. Complex patterns at the community level emerge as a consequence of these ﬂuctuations, even after controlling for differences in historical land-use. In general, species richness, evenness, and diversity declined in response to Hurricane Hugo in areas with appreciable or moderate anthropogenic disturbance in the past. In contrast, areas with little anthropogenic disturbance did not exhibit obvious or consistent trends with regard to indices of diversity. Unfortunately, previous landuse history and susceptibility to disturbance from Hurricane Hugo were confounded spatially, making interpretations of effects difﬁcult. Population density of Eleutherodactylus coqui, the most common frog in the tabonuco forest of Puerto Rico (tabonuco, Dacryodes excelsa, is the dominant hardwood tree) increased after Hurricane Hugo for 2.5 years, but then decreased during the subsequent 2.5 years almost to the levels before the hurricane (Stewart and Woolbright, 1996; Woolbright, 1996). The population increase was related to increases in the number of retreat sites provided by forest-ﬂoor debris from the hurricane, as well as to decreases in the abundance of predators. The subsequent decrease in frog population size was associated with the degradation of woody debris and the resurgence of predator populations. After ﬁve years of recovery from Hurricane Hugo, the two common frugivorous bats attained numbers much greater than those prior to the hurricane (Willig and Gannon, 1996; Gannon and Willig, 1998). In particular, Artibeus jamaicensis was almost seven times as abundant, and Stenoderma rufum was approximately twice as abundant. Moreover, the proportion of reproductively active female S. rufum returned to prehurricane levels of approximately 60%. Less than two years after the impact of Cyclone Ufa, another major cyclonic storm (Val) struck the islands of Samoa (Pierson et al., 1996). This disturbance again impinged on the Tafua Rain Forest Reserve, and within 1.5 months the population of P. tonganus had decreased to 5% of their pre-cyclone (but post-Ufa) numbers. After 8.5 months of secondary succession, densities were even lower – less than

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0.1% of pre-cyclone numbers. By mid-January, 1993 (slightly less than 3 years after Ufa and 14 months after Val), mean population had increased to about 24 individuals in the reserve. Because the population had not recovered from the severe reduction after Cyclone Ufa, it did not confront the same degree of resource depletion per capita after Cyclone Val. Fewer nutritionally compromised bats were observed, and levels of predation by humans and domesticated animals were less, at least in part because of the implementation of conservation education programs to reduce hunting pressure, especially on American Samoa (Daschbach, 1990). In summary, the direct effects of hurricanes on terrestrial animals are poorly documented and anecdotal in nature for the most part. Nonetheless, it is clear that in some instances (reviewed by Wiley and Wunderle, 1993), rain, wind, and ﬂooding can enhance mortality and lead to emigration, while altering migratory patterns and changing geographic distributions. Indirect effects are associated with changes in abundance of food supplies, alteration in the quality and quantity of nesting or roosting sites, changes in exposure to predation, modiﬁcation of the three-dimensional structure of vegetation, altered microclimate, and interactions with human activity. Animals respond with dietary shifts, altered use of habitat and microhabitat, and modiﬁed demographic patterns (i.e., natality, mortality, emigration, and immigration). Historical contingencies (history of previous disturbance and land-use) also affect the response of animals to contemporary disturbances. Grassland ﬁres Because ﬁres have been a consistent disturbance in grasslands over their history (Anderson, 1990), they should have important effects on population and community structure of both plants and animals. Warren et al. (1987) identiﬁed three distinct phases associated with ﬁre disturbance. The combustion phase is the period over which the ﬁre passes through an area. The shock phase is the period between the end of the combustion phase and the initiation of plant regrowth. The recovery phase is the period between the initiation of plant growth and the eventual recovery of animal populations. Fire has a variety of short-term and long-term effects on animal populations and communities. First, ﬁres affect animals directly by causing mortality or

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emigration during the combustion phase (reviewed by Lyon et al., 1978; Warren et al., 1987). Second, ﬁres affect animal populations indirectly by altering the characteristics of the physical environment. This effect should be the strongest during the shock phase. Finally, ﬁre affects animals indirectly by altering the biotic components of the community during the recovery phase. Direct effects Because grassland ﬁres can be produced experimentally, it is possible to examine the direct effects of ﬁre that occur during the combustion phase. Fire directly affects animals at this time. It may either kill the animals or cause them to emigrate from a burned area in an attempt to survive the ﬁre [nonetheless, some species are attracted to ﬁres and may move into a region after ﬁres have been started (Lyon et al., 1978)]. The effect of emigration on eventual population and community structure depends on the rate at which species recolonize following ﬁres. The rate of recolonization may be determined by the size of the ﬁre, or the extent to which the ﬁre alters the physical and abiotic environment. For example, mobile animals, such as birds, may be able to recolonize rapidly following a ﬁre if environmental conditions are appropriate. Many species of ground-nesting birds were observed to lay eggs in a recently burned ﬁeld, although most nests were built in regions of the ﬁeld that were not damaged by the burning (Kruse and Piehl, 1985). Fire directly causes mortality because (1) animals are consumed by the ﬁre, (2) temperatures generated by ﬁres (e.g., Gibson et al., 1990) are lethal (Howard et al., 1959), or (3) individuals suffocate because of reduced concentrations of atmospheric oxygen (Chew et al., 1958). The level of mortality depends on ﬁre intensity and timing, as well as on characteristics of the species. Indirect sources of mortality also occur during the combustion phase. For example, Komarek (1970) and Gillon (1972) observed that predation rates increase during a ﬁre because predators are attracted to the large numbers of escaping arthropods. Animals survive high temperatures associated with grass ﬁres (Gibson et al., 1990) either by ﬂeeing the approaching ﬁre or by remaining in a protected location. Thus, the survival of mobile animals should be greater than the survival of less mobile animals. For example, Warren et al. (1987) noted that orthopterans that are poor ﬂiers, such as crickets (Gryllidae),

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cockroaches (Blattidae), and mantids (Mantidae), experience high mortality during ﬁres, whereas shorthorned grasshoppers (Acrididae) were able to escape the ﬁre because of good ﬂying ability. In addition, parasites may be able to survive a ﬁre by being carried away by their hosts. For example, Stoddard (1946) observed that even the most intense ﬁres were unable to eliminate ticks, presumably because of the survival of individuals attached to vertebrate hosts that escaped the ﬁre. Animals that have immobile life stages may be especially susceptible to being killed by ﬁre. For example, eggs of species that nest on the ground or lay their eggs on vegetation, litter, or the soil surface may be exposed to increased risks of mortality. Fire destroyed all nests of upland bird species such as ringnecked pheasants (Phasianus colchicus) and northern bobwhite (Colinus virginianus) in Nebraska (Erwin and Stasiak, 1979). However, ﬁre is not necessarily fatal to ground-nesting birds. Kruse and Piehl (1985) observed that 69% of clutches of ground-nesting birds survived a ﬁre in a North Dakota grassland. Nonetheless, most of the nests that survived were in areas that did not experience ﬁre. Certain arthropods may be particularly susceptible to ﬁre during the egg stage. For example, ﬁre has been used as a mechanism to control mite species with aestivating eggs (Newman, 1936; Wallace, 1961) and to reduce the density of leaf hoppers that lay eggs on grass (Osborn, 1893; Osborn and Ball, 1897). Some gall-forming insects are highly susceptible to ﬁre as well. Fay and Samenus (1993) observed that all galls of the cynipid wasp Antistrophus silphii were destroyed by an experimental ﬁre on Konza Prairie in Kansas (U.S.A.). However, the large woody galls produced by some Australian Coccinoidea may be adaptations to ﬁre in eucalyptus forests (Koteja, 1986). Animals of low vagility must avoid ﬁre-dependent mortality by minimizing in situ effects. Burrowing is one of the most effective means to do so, because temperature decreases rapidly as depth beneath the soil surface increases (Ahlgren and Ahlgren, 1960; Cooper, 1961; McFayden, 1968). For example, surfacedwelling spiders experience high mortality due to ﬁres, whereas spiders that live in burrows have much lower mortality (Riechert and Reeder, 1972). Burrowing species such as ants also survive initial effects of ﬁre (Warren et al., 1987). In contrast to large mammals that ﬂee to avoid the effects of ﬁre (Lyon et al., 1978), most species of small mammals rely on burrowing to survive ﬁres (Kaufman et al., 1990). In fact, the

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survival of small mammals that nest below ground (e.g., the deer mouse, Peromyscus maniculatus) was much higher than that of species (e.g., the harvest mouse, Reithrodontomys megalotis) that nested above ground (Erwin and Stasiak, 1979; Kaufman et al., 1983, 1988). In general, ﬁre should not have a large direct effect on mortality of the soil macro- and micro-faunas because they are insulated from the direct effects of ﬁre (James, 1982; Seastedt, 1984). Animal species differ in their susceptibility to ﬁre at different life stages. As a consequence, the timing of ﬁres can have important effects on mortality rates. Some species avoid ﬁre by producing ﬁre-resistant stages or by living in locations where they are safe. For example, grasshopper species that overwinter in the egg stage are unaffected by spring ﬁres, whereas species that overwinter as nymphs experience extremely high mortality (Warren et al., 1987). Many attempts to use ﬁre as a means of pest control have shown that the timing of the ﬁre inﬂuences the level of mortality experienced by pest populations (Warren et al., 1987). Animals may reduce their susceptibility to ﬁre by selecting macrohabitats or microhabitats that have a lower probability of burning or experience less intense ﬁres. Fay and Samenus (1993) suggested that cynipid gall wasps choose to lay eggs in large clumps of their host plant rather than on isolated individuals, because these large clumps are less likely to be destroyed by ﬁre. Indirect effects The destruction of litter and vegetation by ﬁre results in an immediate change in the physical and biotic characteristics of the environment. Thus, animal species are affected indirectly by ﬁres during the shock and recovery phases. Fire reduces plant biomass and reduces the extent of the leaf-litter layer, thereby increasing the penetration of light to the soil surface. Increased light penetration results in an increase in soil temperature and a decrease in the soil moisture content (Daubenmire, 1968; Hulbert, 1969; Owensby and Smith, 1973). Whether these environmental changes have positive or negative effects is species-speciﬁc. Physical effects: Indirect responses to changes in the physical environment may be especially common during the shock phase. Some species living in the soil are affected by changes in the physical environment that follow ﬁres. Decreases in populations of soil microarthropods following ﬁres have been attributed to increases in soil temperature and decreases in soil

Michael R. WILLIG and Mark A. McGINLEY

moisture content (Bufﬁngton, 1967; van Amburgh et al., 1981; Seastedt, 1984). The small size and high surface-to-volume ratio of microarthropods may make them especially sensitive to these changes. Alterations of environmental conditions may be favorable for some species. For example, grasshopper nymphs emerged earlier than usual following ﬁres, because increased soil temperatures increased developmental rates (Warren et al., 1987). Litter acts as a food source for detritivores. Thus, populations of detritivores should decrease as a consequence of emigration or mortality immediately after a ﬁre. For example, densities of detritivorous macroinvertebrates in burned areas of Konza Prairie in Kansas were ~50% of those in unburned areas (Seastedt et al., 1985). Similarly, winter burning decreased the summer density of millipedes (Diplopoda) in an Illinois prairie (Rice, 1932). In addition, the densities of species that rely on litter for cover should decrease following ﬁres. Tester and Marshall (1961) attributed decreases in grasshopper abundance following ﬁres to a reduction in cover. Clark and Kaufman (1990) suggested that prairie voles (Microtus ochrogaster) emigrated from burned areas because the reduction of litter did not allow them to construct runways or nests. Conversely, species that prefer more open conditions may respond positively to the removal of litter. For example, Kaufman et al. (1988) and Clark and Kaufman (1990) suggested that deer mice immigrated into burned areas in response to the open vegetation structure and sparse litter cover. Changes in physical characteristics also may limit immigration into disturbed areas following a ﬁre. For example, immigration of spiders following a ﬁre that killed most of the residents was limited to those species that are capable of tolerating decreased moisture availability and the reduced availability of support structures needed for web construction (Riechert and Reeder, 1972). Biological effects: The eventual recovery of animal communities following ﬁre is affected strongly by interactions within the biotic component of the environment. Because plants affect the physical environment, act as a source of cover, or act as a source of food, the recovery of animal communities is predicated on the recovery of the plant community. Similarly, the recovery of certain animal species may be affected by the recovery of other animal species that act as their prey or their predators. Because plant

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community response depends on the season of the ﬁre (Owensby and Anderson, 1967; Bragg, 1982; Towne and Owensby, 1984; Hulbert, 1985), the response of many animal species is dependent on when the prairie burns. Moreover, the plant community may recover relatively slowly following ﬁre, and as a result, the recovery phase for animal species may be protracted. Fires in grasslands affect the phenology, species composition, and productivity of the plant community. Fire tends to increase the proportion of warm-season grasses while decreasing the proportion of coolseason grasses and forbs (Gibson, 1989). In general, burning increases the subsequent production of foliage, rhizomes, and roots of prairie grasses (Hadley and Kieckhefer, 1963; Kucera and Dahlman, 1968). Consequently, population sizes of herbivorous species that either survive the ﬁre or immigrate into burned regions should increase following ﬁres. Numerous studies have documented such an increase in populations of phytophagous hemipterans and homopterans following ﬁres (reviewed by Warren et al., 1987), although this pattern was not ubiquitous. In addition, lepidopteran densities increased following ﬁre (Warren et al., 1987), as did dipterans with herbivorous larvae (van Amburgh et al., 1981). Evans (1984, 1988a,b) studied the response of grasshopper communities to ﬁre on Konza Prairie, and demonstrated how the response of animals depends on ﬁre history. His studies documented the response of grasshoppers to spring ﬁres set on sites that differed in ﬁre history (unburned, burned every year, burned every other year, or burned once every four years). Fireinduced mortality of grasshoppers was low because most of the individuals, regardless of species, were present as eggs in the soil, which provided protection from elevated temperatures and reduced concentrations of atmospheric oxygen (Knutson and Campbell, 1976). Thus, interspeciﬁc differences in response to ﬁre were due to differences in post-hatching survival, emigration, or reproductive rates. Differences in ﬁre history affected plant-community composition and diversity which in turn affected the grasshopper community. Forbs were less abundant in annually burned regions than in less-frequently or unburned regions (Gibson and Hulbert, 1987; Evans, 1988a). Plant species richness and diversity increased with decreasing ﬁre frequency (Abrams and Hulbert, 1987; Gibson and Hulbert, 1987). These differences were mirrored in the composition of the grasshopper community, in that forb feeders became relatively less

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common as ﬁre frequency increased, and grasshopper species richness or diversity was correlated positively with plant richness or diversity. Population sizes of root-feeding arthropods should be higher following burns because of higher root production. Scarabid beetle larvae and cicada nymphs were more abundant in burned than in unburned prairie (Seastedt, 1984; Seastedt et al., 1985). However, densities of some root-feeding homopterans were not affected by burning (Seastedt and Reddy, 1991). Populations of earthworms were higher in burned than in unburned areas (James, 1982), suggesting that these organisms also responded positively to the increased production of roots following ﬁres. Populations of carnivorous species whose prey populations increase following a ﬁre should increase as well. The density of predaceous damselﬂies, dragonﬂies, and wasps increased following ﬁre (Hurst, 1971; van Amburgh et al., 1981). Few studies have monitored simultaneously the ﬁre response of predators and their prey. In summary, animals may escape the direct effects of ﬁre by ﬂeeing or seeking refuge in protected locations. Animals that live or nest below ground tend to suffer fewer direct effects than do animals that live or nest on the surface. Disturbances may indirectly affect animals by causing alterations of the physical environment. Species that feed on or live in the leaf litter may suffer negative effects following the removal of litter by ﬁre, and soil arthropods are affected negatively by the increase in soil temperature and decrease in soil moisture content that occurs following the removal of vegetation and litter. The ultimate recovery of animals following disturbance is inﬂuenced strongly by changes in the biotic community. Populations of above- and below-ground herbivores may increase following ﬁres as a result of increased production of roots and aboveground vegetation. Forest tree-falls In many forest ecosystems, tree-falls are important agents of landscape heterogeneity (Brokaw and Scheiner, 1989; Poulson and Platt, 1989; Spies and Franklin, 1989; Veblen, 1989). Depending on their size and shape, tree-falls open space in the canopy, redistribute biomass to the forest ﬂoor, and alter microclimatic conditions. As a consequence, many forests are a melange of patches that differ because they are in different stages of secondary succession

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(Brokaw, 1985; Runkle, 1985; Veblen, 1985). This has led Lieberman et al. (1989) to advocate abandonment of the treatment of forests as a mosaic of two states – gap and non-gap – in favor of a gradient approach based on a continuum of forest geometries. In either case, spatial heterogeneity should be reﬂected in the distribution and activities of animals as well as plants; consequently, it has been hypothesized that patch-generating phenomena are a diversity-enhancing mechanism in many ecological systems (e.g., Denslow, 1985; Pickett and White, 1985). Birds Tree-fall gaps affect the abundance and diversity of tropical birds in Panama (Schemske and Brokaw, 1981). Some species clearly prefer gaps (e.g., Cyanocompsa cyanoides, Cyphorhinus phaeocephalus, Dysithamnus puncticeps, Hylophylax naevioides, Threnetes ruckeri), whereas others prefer the forest matrix (e.g., Gymnopithys leucaspis, Pipra mentalis, Platyrinchus coronatus). Moreover, 19 additional species of birds, most of them rare or uncommon, were captured exclusively in forest gaps. The vast majority of gap specialists were insectivores, and none were frugivores. As a consequence of these speciesor guild-speciﬁc differences in abundance, distinct bird assemblages were documented for gaps versus the forest matrix. Nectarivorous birds in a Costa Rican cloud forest responded to the distinctive characteristics of treefall gaps as well (Feinsinger et al., 1988). Species richness as well as total density in tree-falls was higher than in mature forest, consistent with the greater density of ﬂowers there. Nonetheless, true gap specialists were not detected for the nectarivore guild. The frequent appearance, low intensity, and rapid recovery of gaps were considered to be driving conditions that allow the accommodation of generalist species to a heterogeneous landscape and diminish selective pressures that favor gap specialists. Similarly, differences at the population and community levels were detected between tree-fall gaps and surrounding matrix in the tabonuco forest of Puerto Rico (Wunderle et al., 1987). Four of 17 species were captured more frequently in gaps than in forest (Chlorostilbon maugaeus, Coereba ﬂaveola, Dendroica caerulescens, Loxigilla portoricensis), but only one of them (D. caerulescens) was not captured in mature forest as well. The distinctive species composition of gaps and forest is more a product of differences in

Michael R. WILLIG and Mark A. McGINLEY

bird density than differences in presence or absence of species, per se. In fact, it was argued that the small size and rarity of gaps in tabonuco forest were characteristics selecting against the production of gap specialists. Those species found more frequently in gaps than in mature forest were canopy species that follow the border of the canopy into and out of gaps. Moreover, differential responses by birds to gaps and forest matrix are context-dependent (Wunderle, 1995). Although new gaps and surrounding forest were statistically distinguishable based on foliage proﬁles one year after the passage of Hurricane Hugo, no signiﬁcant differences existed between their avifaunas at that time. Most birds may not respond directly to the differences in foliage proﬁles between gaps and forest matrix; rather, differential habitat use is based on the existence of distinctive food resources, and these are not sufﬁciently different between habitats after one year of secondary succession. Wunderle (1995) hypothesized that, because of the slow rate of canopy closure in tabonuco forest after a hurricane, many years may pass before gaps and forest matrix are sufﬁciently distinctive in vegetative structure and resource proﬁle to support distinctive bird assemblages. Birds in upland deciduous forest of Illinois (U.S.A.) responded differentially to gaps compared to surrounding forest (Blake and Hoppes, 1986), and did so in a season-dependent fashion. In general, more individuals and greater species richness were recorded in gaps compared to forest understory in spring and fall. Nonetheless, total species richness in gaps was not different from that in forest sites. However, a large number of species exhibited signiﬁcant preferences for gap habitats, including Carduelis tristis, Catharus ustulatus, Dendroica fusca, D. magnolia, Dumetella carolinensis, Empidonax ﬂaviventris, Regulus calendula, R. satrapa, Seiurus aurocapillus, S. noveboracensis, Setophaga ruticilla, Vireo olivaceus, Wilsonia canadensis and Zonotrichia albicollis. Most gap species were birds that foraged on the ground or in lower vegetation, and most were not following the border of the canopy into or out of gaps. Most feeding guilds (e.g., ﬂycatchers, ground insectivores, foliage insectivores, granivoreomnivores, and frugivores) were represented by more individuals in gaps than in forest understory. No guild preferred the forest understory, and only one guild (bark foragers) occurred equally in gaps and in forest. Mammals Rain-forest bats in Australia strongly differentiate

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between forest matrix and moderate-sized (0.03– 0.07 ha) gaps produced by felling (Crome and Richards, 1988). Four bat species (Eptesicus sagittula, Hipposideros ater, Nyctophilus gouldi, Rhinolopus megaphyllus) were recorded exclusively in forest matrix, whereas ﬁve species (Chaerophon jobensis, Chalinolobus nigrogriseus, Mormopterus beccarii, M. planiceps, Scotorepens balstoni) were recorded exclusively in gaps. Only three species (Eptesicus pumilus, Hipposideros diadema, and an unknown taxon, possibly Emballonura nigrescens or a close relative) were “gapincorporators” – in other words, they were detected in both gap and closed-canopy situations. Importantly, gap-incorporators accounted for little of the activity in either habitat (8% in gaps, 27% in closed canopy). This is markedly different from the situation for birds in tropical or temperate ecosystems, where most species are gap incorporators (Schemske and Brokaw, 1981; Blake and Hoppes, 1986; Wunderle et al., 1987; Feinsinger et al., 1988). Moreover, these three groups of bats exhibited distinct morphologies related to considerations of aerodynamic capability (e.g., aspect ratio and wing loading). Gap specialists are fast ﬂyers with high aspect ratio and wing loading, canopy specialists are slower and more maneuverable (low aspect ratio and wing loading), gap-incorporators are intermediate in morphology. Invertebrates Because of their small size and reduced mobility compared to volant vertebrates, one might expect invertebrates to respond more strongly to the dichotomy between gap and forest matrix. Although insects exhibit numerous adaptations to disturbance (Schowalter, 1985), few studies have addressed, in a quantitative fashion, the responses of invertebrates, at population or community level, to gaps created by tree-falls. A comparison of gastropod density in tree-fall gaps to density in adjacent areas of undisturbed understory has been undertaken in the tabonuco forest of Puerto Rico [unpublished observations by Alvarez (1991) and Alvarez and Willig (1993)]. Only three species were exclusively captured in gaps (Cepolis squamosa, Oleacina playa, Vaginulus occidentalis), and each of them was quite rare (approximately 0.1% of the total captures). Five species evinced sufﬁciently high density to allow statistical evaluation of disturbance effects. Three species (Austroselenites alticola, Megalomastoma croceum, Subulina octana) did not differ in density between gap and undisturbed understory. In

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contrast, the density of Nenia tridens was higher in gaps than in understory, while the density of Caracolus caracolla was higher in undisturbed understory than in the gaps. The higher abundance of N. tridens in gaps was attributed to increased availability of substrate (dead wood) and food (fungi and algae). The preference of C. caracolla for undisturbed areas of the understory may be related to abiotic factors and physiological limitations (microclimatic conditions reducing evaporative water loss). One of the snail species (C. squamosa), exclusively detected in gaps during pre-hurricane surveys, became a common member of the snail community within one year of the hurricane’s impact (Willig et al., 1998). As the canopy closed during secondary succession, quantities of dead and decaying branches decreased; concomitantly, abundance of C. squamosa decreased from approximately 190 individuals ha−1 in 1991, to less than 50 individuals ha−1 in 1995 (Willig et al., 1998). At the community level, a number of important differences exist between gaps and forest understory. Snail species diversity in quadrats of 8 m2 in Puerto Rico was signiﬁcantly higher in undisturbed understory than in tree-fall gaps (Alvarez, 1991; Alvarez and Willig, 1993). Moreover, differences in snail species composition between gaps and understory were consistent, regardless of season. Compositional differences were reﬂected in the greater turnover of species between gap and undisturbed forest, than between sites within gaps or between sites within forest understory. In summary, few data are available to document the direct effects of tree-falls on populations or communities of animals. Indirect effects include alteration of microclimatic conditions, especially temperature and humidity, addition of biomass to the forest ﬂoor, which increases structural complexity and the quantity of dead and decaying material, and recruitment of earlysuccessional plants. The response of animal species depends upon their trophic position, physiological capabilities, mobility, and interactions with other members of the biota.

ANIMALS: UBIQUITOUS AGENTS OF DISTURBANCE

Animals are agents of disturbances through their usual activities including movement, feeding, building, digging, burrowing, elimination, and death. Animalgenerated disturbances generally occur more fre-

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Michael R. WILLIG and Mark A. McGINLEY

Fig. 27.3. Conceptual model illustrating the mechanisms whereby the direct effects of animal disturbance on environmental conditions indirectly affect the biota.

quently, but are smaller in extent and smaller in magnitude than are abiogenic disturbances generated by climatological or geological events. However, animalgenerated disturbances can have profound direct and indirect effects (Schowalter and Lowman, Chapter 9; Oesterheld et al., Chapter 11; MacMahon, Chapter 12; Bradbury, Chapter 24; this volume) on the structure of plant, animal, and microbial communities (Fig. 27.3). In general, the importance of animalgenerated disturbances comes not from their direct effects as destructive events, but rather from the indirect effects that arise as a result of their alteration of the environment. An important effect of many animalgenerated disturbances is the creation of novel habitats that often harbor their own unique combinations of species. Movement The physical forces that derive from locomotion by animals can act as a disturbance, either by causing mortality or by altering environmental conditions (Harper, 1977). For example, trampling by grazing animals can reduce the depth of leaf litter (Johnston,

1961; Johnston et al., 1971; Langlands and Bennett, 1973; Leege et al., 1981; Naeth et al., 1991; Holland, 1994), increase proportion of bare ground (Langlands and Bennett, 1973; Bassett, 1980; Williams, 1992), or compress the soil (Johnston et al., 1971; Langlands and Bennett, 1973; Gibson, 1988). Even the coiling behavior of rattlesnakes can alter the distribution of leaf litter on the soil surface (McGinley, pers. observ.). The effects of animal movement depend on a number of characteristics, including the mass of the animal, as well as the number of animals and the frequency with which they move through an area. These likely depend on the social and spacing system of a species. For example, the effects of movement of colonial animals are more concentrated than are the effects of solitary animals. Moreover, the effect of movement may depend on the activity with which it is associated. If animals are traveling to and from foraging or mating areas, their activities usually are concentrated on trails, where the repeated passage of animals effects the removal of vegetation. However, if animals are traversing a habitat in search of food, the effects of movement are more dispersed throughout the area. Trampling may indirectly alter abiotic soil char-

ANIMALS AND DISTURBANCE

acteristics. The fragmentation and reduction of leaf litter increases the amount of insolation reaching the soil surface and consequently results in an increase in soil temperature (Weaver and Rowland, 1952; Johnston et al., 1971; Langlands and Bennett, 1973; Knapp and Seastedt, 1986) and a decrease in soil moisture content (Johnston et al., 1971; Langlands and Bennett, 1973; Knapp and Seastedt, 1986). Trampling by cattle can increase rates of litter decomposition, which in turn affects rates of loss of nutrients due to leaching (Knapp and Seastedt, 1986). Furthermore, trampling compacts the soil (Johnston et al., 1971; Langlands and Bennett, 1973; Gibson, 1988), which can reduce the amount of nutrients lost through leaching (Langlands and Bennett, 1973; Knapp and Seastedt, 1986). The relatively ubiquitous effect of activity by cattle likely homogenizes many aspects of the environment (Rusch, 1992). Digging and burrowing Many species search for food by digging, or dig burrows to provide protection, nesting sites, and secure food caches. In fact, the burrowing activity of soil micro- and macrofauna is essential for the formation and maintenance of soil characteristics (Killham, 1994). Clearly, soil animals play an important role in decomposition. Burrowing animals, especially earthworms, are responsible for mixing organic matter throughout the soil proﬁle (Killham, 1994), thereby affecting the spatial distribution of plants and microbes. The direct effects of digging and burrowing by larger animals include the death of plants by burial, and the exposure of the soil surface creating bare spaces. The effects of digging and burrowing may be concentrated when associated with colonial species such as prairie dogs (Cynomys spp.), or dispersed across a habitat for more solitary species such as pocket gophers (Geomys spp.), kangaroo rats (Dipodomys spp.), and badgers (Taxidea taxus). In general, digging activity generates heterogeneity in environmental conditions such as soil depth (Kershaw, 1959) and nutrient content (e.g., Gibson, 1988; Moorhead et al., 1988; Reader and Buck, 1991; Dhillion et al., 1994; McGinley et al., 1994). In addition, digging affects the distribution, abundance, and composition of soil microbial communities (Allen et al., 1992; Friese and Allen, 1993; Dhillion et al., 1994; McGinley et al., 1994) and opens sites for plant regeneration. The effects of animal-generated

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disturbance on plant communities have been well documented (e.g., Platt, 1975; Hobbs and Mooney, 1985; Huntly and Inouye, 1988; Moorhead et al., 1988; Peart, 1989; Cofﬁn and Lauenroth, 1990; Mun and Whitford, 1990; Dean and Milton, 1991; Reader and Buck, 1991). Interestingly, the production of mounds may also have direct effects on animal populations. For example, grasshopper abundance was correlated with the density of gopher mounds, because the mounds were used as oviposition sites by the grasshoppers (Huntly and Inouye, 1988). In addition, animal burrows represent novel habitats. The burrows of prairie dogs are used by a number of other species of animals (Hoogland, 1995). For example, the abundance and spatial distribution of the burrowing owl (Speotyto cunicularia) depends on the availability of preexisting burrows of rodents or badgers (Best, 1969; Green and Anthony, 1989). Heteromyid rodents often store food in large subterranean burrow systems (Voorhies and Taylor, 1922; Reichman et al., 1985). The warm temperature and high humidity in the burrows of the banner-tailed kangaroo rat (Dipodomys spectabilis) allow the development of a diverse and distinctive fungal community compared to that found outside the burrow (Reichman et al., 1985). Building Many birds construct nests (Welty, 1982); some mammals such as beavers (Grinnell et al., 1937), muskrats (Errington, 1963), and woodrats (Linsdale and Tevis Jr, 1951; Rainey, 1956; Finley, 1958) build houses; and some insects such as termites, wasps, and ants build mounds or nests (Borror et al., 1989). The collection of materials for house building by mammals such as beavers and woodrats directly alters the distribution of woody biomass in the ecosystem. Not only does house building concentrate wood at the house site, but because selectivity varies as a function of distance (McGinley, 1984), it also results in an unequal distribution of stick sizes across the landscape. This redistribution of woody biomass can alter patterns of nutrient cycling. The concentration of wood in houses produces “islands of fertility” which affect the subsequent growth of woody and herbaceous plants (Zak et al., 1994). Moreover, woodrat houses are inhabited by a variety of species, including snakes, lizards, toads, and small mammals (Fitch and Rainey, 1956). The building of dams by beavers results in larger

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effects than would be predicted by their size, activity, or abundance (Naiman et al., 1994). Beaver dams alter stream characteristics, alter nutrient ﬂow, modify riparian zones, and create wetland habitat (Naiman et al., 1986, 1994; Naiman, 1988; Smith et al., 1991). Such alterations affect the use of habitat by terrestrial animals, as well as altering population sizes and community composition of terrestrial insects with aquatic larvae (Clifford et al., 1993). Some termites build large mounds (Howse, 1970; Lee and Wood, 1971). Nutrients in the soil of termite mounds are unavailable for use by plants. For example, mounds in a savanna in the Northern Territory, Australia, held 2% of the total soil, 9% of the total organic carbon, 5% of the total nitrogen, and 11% of the total calcium in the system (Lee and Wood, 1971). The redistribution of nutrients by termites can affect patterns of vegetation, and large termite mounds can support distinct vegetation (Lee and Wood, 1971). The mounds of some species of termites contain fungal gardens which provide sites for a unique fungal community (Bakshi, 1962; Howse, 1970). Waste products The disposal of waste products by animals (i.e., elimination and defecation) has important direct and indirect effects. Fecal deposition by cattle may directly affect plants by acting as a cause of mortality or a barrier to the establishment of seedlings, and indirectly by inﬂuencing abiotic soil characteristics (Harper, 1977). For example, the deposition of dung by cattle can add nutrients such as nitrogen, phosphorus, potassium, calcium, magnesium, and a number of trace elements to the soil (Petersen et al., 1956b; Underwood, 1956; Barrow, 1967; Weeda, 1967; MacDiarmid and Watkin, 1972). As a result, nutrient content directly beneath cattle feces is different than that of adjacent soils. The deposition of feces by smaller herbivores, such as rabbits, also inﬂuences soil fertility. For example, the feces of jackrabbits (Lepus californicus) annually represent up to 4% of the standing crop of nitrogen in the Chihuahuan Desert (Shoemaker et al., 1973; Gist and Sferra, 1978). Defecation generates spatial and temporal heterogeneity in environmental characteristics. Because of patterns of cattle activity, feces are not distributed uniformly across the landscape (Petersen et al., 1956a). Moreover, differences in rates of release of nutrients from decomposition, and nutrient-speciﬁc differences

Michael R. WILLIG and Mark A. McGINLEY

in lability in the soil, can generate variation in nutrient content over time. For example, nitrogen and potassium under cattle dung attain a peak in concentration soon after deposition, and decrease rapidly (Weeda, 1967; MacDiarmid and Watkin, 1972), so that their effects are short-lived. In contrast, phosphorous and calcium in the soil generally increase more slowly and remain elevated for extended periods of time (Weeda, 1967; MacDiarmid and Watkin, 1972). The addition of nutrients through feces affects plant growth and community composition. In pastures, addition of cow dung increased productivity of adjacent plants (Weeda 1967; MacDiarmid and Watkin, 1972). Moreover, the alteration of nutrient content results in changes in species composition of plant communities (Norman and Green, 1958; Weeda, 1967; MacDiarmid and Watkin, 1971). The addition of dung from smaller herbivores may affect plant growth as well. Watt (1981) suggested that the growth and survival of several species were affected by decomposition of rabbit dung. The addition of feces to the environment provides a novel habitat for colonization by decomposers. From that perspective, the community dynamics during dung decomposition has been well studied (e.g., Valiela, 1969; Hanski and Koskela, 1977; Doube, 1987). In general, feces represent an ephemeral resource that is variable in space and time (Doube, 1987). Fecal communities were similar to, but only a subset of, communities found in decomposing carcasses (Doube, 1987; and see below). Interestingly, the dung community is well adapted to the characteristics of the feces of the native fauna. For example, when cattle were introduced to Australia, their dung was decomposed poorly by the native fauna. As a consequence, dung beetles were introduced to break down feces of these exotic mammalian herbivores (Waterhouse, 1974; Doube, 1987). When defecation is localized, the accumulation of feces can form the basis of unique ecosystems. Such communities are found in caves in the accumulations of guano produced by bats (e.g., Martin, 1977; Benarth and Kunz, 1981; Culver, 1982; Conn and Marshall, 1991; Whitaker et al., 1991) or crickets (Poulson and Kane, 1981; Poulson et al., 1995). The input of guano and other detritus provides the base of the food chain for these communities, described at length in Volume 30 of this Series. The amount and quality of guano differs between sites and varies over time in response to differences in climate (Poulson et al., 1995), or the diet

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and physiological state (e.g., maternity colonies versus hibernacula) of the bats (Studier et al., 1994). Such communities contain disparate feeding guilds including fungivores (e.g., acarinans, psocopterans), detritivores (e.g., coleopterans, dipterans), predators (e.g., acarinans, coleopterans, hemipterans), and ectoparasites (e.g., hemipterans, siphonopterans; Benarth and Kunz, 1981; Whitaker et al., 1991; Trajano and GnaspiniNetto, 1994). Moreover, many of the microclimatic attributes (temperature, humidity, ammonia concentration) of these cave ecosystems are a consequence of the processing of guano by the resident biota. Death The decomposition of animal carcasses alters environmental conditions in a manner that affects surrounding vegetation and soil fauna for an extended period of time. According to Bornemissza (1958), plants beneath a carcass may be killed, but the growth of plants in a zone surrounding the carcass increase. In addition, Bornemissza found that the decaying carcass interacted with leaf litter to form a crust that inhibited plant regeneration for over a year. By-products from the decomposition of carrion reduced the density of the usual soil fauna. For example, the densities of Collembola and Acari were reduced beneath a carcass during particular periods of decomposition, and some species were eliminated completely for a period of over a year. Carcasses provide novel habitats for organisms in the detrital circuit. The succession of animal species that inhabit decaying carcasses has been well documented (Putman, 1978a; Beaver, 1984; Schoenly and Reid, 1987). Putman (1978b) examined the ﬂow of energy and nutrients from a carcass during decomposition. In fact, patterns of carrion succession are so predictable that they have been used by forensic scientists to estimate the time of death (e.g., Catts and Goff, 1992; Goff, 1993). Decomposing carcasses contain diverse communities including primary consumers such as ﬂy larvae and dung beetles, and higher-level consumers, such as predatory beetles (reviewed by Doube, 1987). Guilds of primary consumers include species that consume moist ﬂesh (e.g., sarcophagid ﬂy larvae), dry ﬂesh (e.g., phorid ﬂies), skin (e.g., dermestid beetles), and dung (e.g., scarabaeid dung beetles), whereas guilds of secondary consumers include both predators and parasitoids. These taxa and feeding guilds would not otherwise occur within the terrestrial community.

647 A CASE STUDY: THE SAND SHINNERY OAK ECOSYSTEM

The effects of animals on the sand shinnery oak ecosystem of western Texas provide an illustrative example of biogenic disturbances (Fig. 27.4). Sand shinnery oak (Quercus havardii) is a low-growing (<1 m tall), clonal species that is dominant on sandy soils in the Southern High Plains (Dhillion et al., 1994). The plant community consists of an almost continuous cover of oak (>85%) with an understory of predominantly herbaceous, prairie species (Dhillion et al., 1994; Holland, 1994). Because of the abiotic factors that characterize this ecosystem (i.e., sandy soils, low rainfall, and cold winter temperatures; Dhillion et al., 1994), disturbances produced by animals may be especially important. The oak is deciduous and rates of decomposition are low, so the soil is covered by a layer of leaf litter which appears to limit plant regeneration (Dhillion et al., 1994; and unpublished observations of McGinley and Jeffery). The sandy soil is home to a number of species of burrowing animals such as pocket gophers (Geomys bursarius), cottontail rabbits (Sylvilagus audubonii), heteromyid rodents (e.g., Dipodomys ordii and Perognathus ﬂavescens), and several species of ants (e.g., Crematogaster punctulata, Pheidole dentata, Pogonomyrmex barbatus), whose digging or mound-building activities produce soil disturbances that open sites in the leaf litter through the removal or burial of litter (Dhillion et al., 1994). Grazing by cattle is a common agricultural practice in this region, so the sand shinnery oak ecosystem is disturbed by the activity of cattle as well. Dhillion et al. (1994) reported that soil disturbances produced by animals created variation in abiotic and biotic characteristics of soils. For example, soil associated with active mounds of cottontail rabbits had higher nitrogen and magnesium contents, higher organic matter concentrations, and higher densities of fungi and bacteria than did soils from adjacent nonmound areas. Soil nutrient content, organic matter content, moisture, pH, and the potential for mycorrhizal infection differed among soils associated with mounds built by the three species of ants mentioned (McGinley et al., 1994). The addition of feces by cattle and rabbits increased nutrient and organic matter, and altered the microbial content of soil (unpublished observations of McGinley and Purdom). Experiments that mimicked animal disturbance through the removal of leaf litter were sufﬁcient to alter the nutrient

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Michael R. WILLIG and Mark A. McGINLEY

Fig. 27.4. Conceptual model illustrating the mechanisms through which animal disturbances affect populations and communities of plants and animals in the sand shinnery oak ecosystem of western Texas. Effects of animal disturbance on soil characteristics appear at landscape or patch scales, which together affect plant populations and communities.

content of soils for at least three years (unpublished observations of McGinley and Jeffery). Thus, the activity of animals acted as an important source of environmental heterogeneity, and the effects of animal disturbance persisted for long periods of time. Environmental heterogeneity created by animal disturbance affected seedling establishment. Seedlings of the perennial grasses, sand dropseed (Sporobolus cryptandrus) and Lehman’s lovegrass (Eragrostis lehmanniana), survived better in greenhouse studies when grown in soil collected from gopher or harvesterant mounds than when grown in soil collected from rabbit mounds (unpublished observations of McGinley et al.). Environmental heterogeneity also affected seedling growth rates. For example, seedlings of the perennial grass, little bluestem (Schizachyrium scoparium), grown in the greenhouse in soil collected from mounds of Pogonomyrmex barbatus were twice as large as seedlings grown in soil collected from mounds of Crematogaster punctulata and Pheidole dentata. This difference likely occurred because soils from Pogonomyrmex mounds contained more nitrogen than did soils from mounds of the other two species (McGinley et al., 1994).

Not all species responded to animal-generated environmental variation in the same manner. In greenhouse studies, seedlings of little bluestem, unlike those of sand dropseed and Lehman’s lovegrass, survived equally well in soil collected from harvester-ant, gopher, or rabbit mounds. Seedlings of blue grama (Bouteloua gracilis) and little bluestem were larger when grown in soil collected beneath rabbit feces than when grown in typical soil, whereas the growth rate of seedlings of sand lovegrass (Eragrostis trichodes) was not affected by the deposition of rabbit feces (unpublished observations of McGinley and Purdom). Thus, animal disturbance affected different species in different ways. An important effect of animal-generated soil disturbances in this community was to open sites for plant regeneration (Dhillion et al., 1994). Seventy-one percent of seedlings were associated with disturbed sites even though bare ground constituted only 19% of the total area (Dhillion et al., 1994). Moreover, herbaceous plant density was positively correlated with the number of disturbances in ungrazed areas (unpublished observations of McGinley). Thus, animal disturbance played an important role in determining the

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abundance of herbaceous plants in sand shinnery oak communities. Animal disturbance also affected patterns of plant dispersion. Animal disturbances were distributed in a patchy fashion (Holland, 1994); because animal disturbance inﬂuences plant establishment, patterns of herbaceous plant dispersion should match patterns of disturbance. As expected, there were similarities in the spatial dispersion of Euphorbia fendleri, a species that can only establish in bare ground (unpublished observations of McGinley and Jeffery) and the spatial distributions of animal disturbances (Holland, 1994). Environmental heterogeneity associated with animalgenerated disturbances of soil was sufﬁciently large to alter plant–microbe interactions (McGinley et al., 1994). For example, infection by mycorrhizal fungi enhanced the growth of seedlings of little bluestem when grown in the greenhouse in soil from ant mounds (P. dentata and C. punctulata) with low nitrogen content. However, mycorrhizal infection decreased the growth of seedlings of little bluestem in soil from ant mounds (P. barbatus) with high nitrogen content. Environmental heterogeneity produced by animal disturbance shifted the mycorrhiza–plant interaction from mutualism to parasitism. Cattle activity inﬂuenced patterns of abiotic characteristics as well as patterns of herbaceous plant dispersion, abundance, and diversity in sand shinnery oak ecosystems (Holland, 1994). The activity of cattle reduced litter depth and increased the cover of bare ground. The concentration of soil nutrients other than nitrogen (i.e., calcium, magnesium, phosphorus, and sodium) was lower in grazed regions than in ungrazed regions. In addition, variation in the content of most soil nutrients (calcium, magnesium, phosphorus, and sodium; but not nitrogen) was higher in grazed than in ungrazed areas. This, combined with the fact that dispersions of herbaceous plants were more clumped in the ungrazed region than in the grazed region, suggests that cattle activity had a homogenizing effect in this environment. The greater proportion of bare ground and shallower depth of litter caused by the activity of cattle offered increased opportunities for seedling establishment in grazed regions. The density of herbaceous plants was almost twice as high in grazed as in ungrazed regions (Holland, 1994). Because species richness in a quadrat of 4 m2 is highly correlated with the number of individual plants in that area, quadrats in grazed areas contained approximately twice as many species as did

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quadrats in ungrazed areas. Thus, disturbance by cattle increased species richness at a local scale. Animal-induced changes in the physical environment, as well as modiﬁcations of the spatial dispersion, abundance, composition, and diversity of the plant community, may have indirect effects on animal populations and communities. Population sizes of herbivores may be predicted to respond positively to increased abundance of herbaceous plants associated with grazing. Grasshoppers (e.g., Arphia conspersa, Brachystola magna, Dactylotum bicolor, Melanoplus spp. and Schistocerca alutacea) were common herbivores in the sand shinnery oak ecosystem, and their densities were higher in grazed than in ungrazed regions of this community (M.A. McGinley, pers. observ.). In contrast, the activity of cattle did not affect differences between grazed and ungrazed regions in the abundance or dispersion of ant or gopher mounds (Holland, 1994). The consequences of disturbance on the abiotic environment and structure of the plant community lingered for some time. The effects of litter removal (experimental treatments differed in the amount and pattern of litter removal) were still evident three years after the initial disturbance (unpublished observations of McGinley et al.). Plots in which all litter was removed, subsequently had more bare ground and shallower leaf litter even after three seasons of leaf fall. In addition, soil nutrient content differed between plots where all litter was removed and plots that were continuously covered by litter. Moreover, seedlings of sand dropseed were larger when grown in the greenhouse in soil collected in plots that had been continuously covered with litter than when grown in soil from plots where litter was removed. These differences in seedling growth led to differences in the size distributions of adult plants in the ﬁeld. Plants grown in undisturbed plots in the ﬁeld were larger than plants growing in plots where litter was removed. Moreover, abundances and species richness of herbaceous plants were higher in plots where litter had been experimentally removed three years earlier compared to those which were undisturbed. In summary, the effects of any single animalgenerated disturbance was small and localized. Nonetheless, their cumulative effects were large because they were frequent, pervasive, and potentially long-lasting in this community. A common effect of animal disturbance was the opening of sites for herbaceous plant regeneration by removal and burial of the leaf litter. In addition, animal disturbance acted as a source

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Michael R. WILLIG and Mark A. McGINLEY

Table 27.1 Generalizations concerning the characteristics of animals which inﬂuence their responses to direct and indirect effects of disturbances Characteristic

Description

Mobility

organisms of low mobility are more likely to suffer high mortality because of their inability to avoid the deleterious effects of a disturbance by seeking refugia in situ or ﬂeeing to other areas; similarly, rates of establishment should be higher for mobile species (high colonization rate) with appropriate phenotypic attributes to prosper in a post-disturbance environment

Habitat/microhabitat

subterranean organisms as well as organisms that inhabit protected locations (e.g., caves) should experience lower direct mortality than do surface-dwelling animals; associations between habitat afﬁnity and indirect responses depend upon the peculiarities of the disturbance and speciﬁc habitat requirements of the species

Life history

different life stages (e.g., pupae versus adult) may be more or less vulnerable to direct effects of disturbance, depending on the particulars of the disturbance and the phenotypes of the life stages; in addition, organismal response to disturbance may be dependent on the timing of the disturbance with respect to the phenology and ontogeny of the species

Body size

smaller animals (higher surface to volume ratios) should be more vulnerable to alterations in temperature and humidity as a consequence of disturbance; on the other hand, small size may allow species to exploit microhabitats that provide protection from the altered abiotic environment; smaller species may also recover more rapidly because of their higher reproductive rates and shorter generation times

Trophic position

responses of animals to direct effects of a disturbance are independent of trophic status; the response of a species to changes in the biotic environment should depend on how direct and indirect effects have altered its food supply, as well as the population sizes of its competitors and predators

of environmental heterogeneity, which can potentially inﬂuence seedling growth and survival. Thus, animal disturbance had strong effects on patterns of plant dispersion, abundance, and diversity suggesting that a thorough knowledge of disturbance history and the mechanisms through which disturbances affect seedling establishment is necessary to understand patterns of community structure in this ecosystem.

CONCLUSIONS

Disturbances affect animals directly or indirectly through modiﬁcations of the physical or biological environment. Direct effects occur as a consequence of mortality or emigration; however, they are rarely documented. Although disturbances can act as sources of mortality, many species can avoid mortality by ﬂeeing the area or escaping detrimental effects in situ. The ability of animal species to survive disturbance is inﬂuenced by the timing, magnitude, and intensity of the disturbance, as well as by the phenotype and lifehistory characteristics of the animal (Table 27.1). Disturbances alter the physical environment; in terrestrial systems, this involves the removal or redistribution of vegetation. As a consequence, patches of various sizes experience increased exposure to sunlight, which results in an increase in temperature

and a decrease in humidity. These indirect effects are most frequently cited as causes of short-term responses to disturbance. Indeed, small, less mobile, or heterothermic organisms may be especially susceptible to changes in the physical environment. Animal populations and communities change in response to alterations in the plant, animal, and microbial communities that follow a disturbance. Although initial responses are species-speciﬁc, much of the variation in trajectories of recovery is related to trophic position or guild afﬁliation. Despite the intensity of many disturbances, some animal populations and communities return to initial conditions relatively rapidly (Zimmerman et al., 1996). Animals are important agents of disturbance, mostly as a consequence of the manner in which their activities alter the physical environment and create novel habitats. Compared to many abiogenic disturbances, the effects of individual animal-generated disturbances are small, but because they occur frequently they have major cumulative effects on the abiotic environment, as well as on microbial, plant, and animal communities. In rare circumstances (e.g., beaver dams), animal activity drastically alters the structure and function of communities and ecosystems, or represents the deﬁning characteristic of the ecosystem (e.g., guano communities in caves).

ANIMALS AND DISTURBANCE

One of us (MRW) was once asked after a presentation concerning the role of animals in a forested ecosystem, “What would the forest be like if it contained no animals?” The somewhat startled response was “It would not be a forest at all, just as if the system contained no trees. It is critical to see the forest beyond the trees!” Clearly this statement represents more of a belief than a fact at this point in time. Indeed, the challenge to animal ecology in particular, and systems ecology in general, is to document the situations in which animals are important agents of ecosystem structure and function, and to identify the mechanisms of the interactions and attributes of system under the regulation of animal activities.

FUTURE DIRECTIONS

Progress in understanding the complex nature of disturbance as it affects populations and communities of animals requires a multifaceted approach that differs in essence from those characterizing most past studies. In particular, we recommend approaches which integrate experimental manipulations, comparative approaches, and long-term perspectives. In addition, research assessing dynamic processes (e.g., rates of emigration, immigration, natality, mortality, and species turnover) rather than static attributes (e.g., density, diversity) should provide greater insight to mechanisms of response to disturbance. Manipulative experiments in the ﬁeld are valuable tools for addressing questions in ecology (Hairston, 1989); however, this approach has been used less in studying disturbance than in other areas of ecology. The lack of an experimental approach to disturbance ecology is not surprising, given the difﬁculty of producing effects analogous to those of intense abiogenic disturbances such as hurricanes and landslides. However, when it has been feasible to experimentally create disturbances (e.g., controlled burns, artiﬁcial tree-falls or gaps, experimental animal exclusions), it has been possible to dissect direct and indirect effects of disturbance in detail. Because of differences in the spatial and temporal scale necessary to study disturbances effectively, the experimental approach has been applied much more often to study the response of plants than the response of animals. Nonetheless, manipulative experiments are powerful tools for ﬁeld ecologists examining the responses of animals to disturbance because they allow researchers

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to (1) control the timing, intensity, and extent of disturbance, (2) document pre-disturbance conditions, (3) distinguish between the relative importance of direct responses and indirect responses, and (4) disentangle the effects of confounding factors associated with particular disturbance events (e.g., distinguish the effects of opening the canopy from those of redistributing biomass to the forest ﬂoor). Most studies concerning the response of animal populations and communities to disturbances have focused on events whose timing, location, or extent are beyond the control of the investigator. As a consequence, researchers usually have inadequate background information about environmental conditions, or population and community structure prior to the disturbance. Pseudoreplication is a pervasive problem. However, researchers conducting long-term studies have been well poised to take advantage of opportunities presented by rare disturbance events (see Walker et al., 1991, 1996) because they have considerable information about salient environmental features before the disturbance. Thus, long-term ecological studies are essential for understanding responses of animal populations and communities to disturbances that cannot be manipulated experimentally. Comparative approaches offer considerable insight to understanding the manner in which disturbances affect populations and communities. Comparisons that contrast taxonomic groups (e.g., birds versus bats) or functional groups (e.g., frugivores versus insectivores) are particularly promising. In addition, understanding the context-dependent nature of disturbance is critical (e.g., contrasting the effects of tree-falls when rare and dispersed, to the effects when common and in close proximity to each other). Finally, human activities may be classiﬁed in the same manner as those of other animals (transport, construction, waste disposal, etc.). Human populations continue to grow at alarming rates, with attendant environmental assaults on virtually all terrestrial (and aquatic) ecosystems. It is all too clear that ecology in the next millennium will become the study of disturbed ground. Clear understanding of the effects of animal-generated disturbance on natural ecosystems may provide insight to understanding the impact of anthropogenic disturbance as well. ACKNOWLEDGMENTS

We would like to thank the editor, Lawrence Walker,

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for inviting us to contribute to this volume and for helping to clarify the exposition. Many have contributed to the gestation of ideas concerning animal disturbance, especially our graduate students at Texas Tech University and our colleagues in the Luquillo Mountains Long-term Ecological Research program. In addition, J. Alvarez, R. Gee, S. Huskey, N. Mehdiabadi, M. Mills, J. Neumann, D. Nordmeyer, J. Jeffery, and A. Purdom provided access to unpublished data. D. Hall and C. Guthrie provided editorial assistance, and T. Schowalter, S. Cox, D. Yee, C. Guthrie, and an anonymous reviewer provided helpful comments on earlier versions of the manuscript. In part, this research was performed under grant BSR-8811902 from the National Science Foundation to the Terrestrial Ecology Division, University of Puerto Rico and the International Institute for Tropical Forestry, as part of the Long-term Ecological Research Program in the Luquillo Experimental Forest. Additional support was provided by the Forest Service (U.S. Department of Agriculture), the University of Puerto Rico, and Texas Tech University. Work at the Sand Shinnery Oak site was facilitated by Mr. J. Fitzgerald, and supported in part by the Howard Hughes Medical Institute, the Clarke Foundation, and Texas Tech University.

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Michael R. WILLIG and Mark A. McGINLEY Waide, R.B., 1991a. Summary of the response of animal populations to hurricanes in the Caribbean. Biotropica, 23: 508–512. Waide, R.B., 1991b. The effects of Hurricane Hugo on bird populations in the Luquillo Experimental Forest, Puerto Rico. Biotropica, 23: 475–480. Waide, R.B., 1996. Birds. In: D.P. Reagan and R.B. Waide (Editors), The Food Web of a Tropical Forest. University of Chicago Press, Chicago, pp. 363–398. Waide, R.B. and Lugo, A.E., 1992. A research perspective on disturbance and recovery of a topical montane forest. In: J.G. Goldammer (Editor), Tropical Forests in Transition. Birkhauser Verlag, Basel, pp. 173–190. Walker, L.R., Brokaw, N.V.L., Lodge, D.J. and Waide, R.B. (Editors), 1991. Ecosystem, plant, and animal responses to hurricanes in the Caribbean. Biotropica, 23: 313–521. Walker, L.R., Silver, W.L., Willig, M.R. and Zimmerman, J.K. (Editors), 1996. Long-term responses of Caribbean ecosystems to disturbance. Biotropica, 28: 414–614. Wallace, M.M.H., 1961. Pasture burning and its effects on the aestivating eggs of Halytydeus destructor (Tuck.). Aust. J. Agric. Anim. Husbandry, 1: 109–111. Warren, S.D., Scrifes, C.J. and Teel, P.D., 1987. Response of grassland arthropods to burning: a review. Agric. Ecosyst. Environ., 19: 105–130. Waterhouse, D.F., 1974. The biological control of dung. Sci. Am., 230: 100–109. Watt, A.S., 1981. A comparison of grazed and ungrazed grassland in East Anglia Breckland. J. Ecol., 69: 499–508. Weaver, J.E. and Rowland, N.W., 1952. Effects of excessive natural mulch on development yield, and structure of native grassland. Bot. Gaz., 114: 1–19. Weeda, W.C., 1967. The effect of cattle dung patches on pasture growth, botanical composition, and pasture utilization. N. Z. J. Agric. Res., 10: 150–159. Welty, J.C., 1982. Life of Birds. Saunders, Philadelphia, 546 pp. Whitaker Jr., J.O., Clem, P. and Munsee, J.R., 1991. Trophic structure of the community in the guano of the evening bat Nycticeius humeralis in Indiana. Am. Midl. Nat., 126: 392–398. White, P.S. and Pickett, S.T.A., 1985. Natural disturbance and patch dynamics: an introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Inc., San Diego, pp. 3–13. Wiles, G.J., 1987. Current research and future management of Marianas fruit bats (Chiroptera: Pteropodidae) on Guam. Aust. Mammal., 10: 93–95. Wiley, T.R. and Wunderle Jr., J.M., 1993. The effects of hurricanes on birds, with special reference to Caribbean islands. Bird Conserv. Int., 4: 1–31. Will, T., 1991. Birds of a severely hurricane-damaged Atlantic coast rain forest in Nicaragua. Biotropica, 23: 497–507. Williams, R.J., 1992. Gap dynamics in subalpine heathland and grassland vegetation of south-eastern Australia. J. Ecol., 80: 345–352. Willig, M.R. and Camilo, G.R., 1991. The effect of Hurricane Hugo on six invertebrate species in the Luquillo Experimental Forest of Puerto Rico. Biotropica, 23: 455–461. Willig, M.R. and Gannon, M.R., 1996. Mammals. In: D.P. Reagan and R.B. Waide (Editors), The Food Web of a Tropical Forest. University of Chicago Press, Chicago, pp. 399–431.

ANIMALS AND DISTURBANCE Willig, M.R., Secrest, M.F., Cox, S.B., Camilo, G.R., Cary, J.F., Alvarez, J. and Gannon, M.R., 1998. Long-term monitoring of snails in the Luquillo Experimental Forest of Puerto Rico: heterogeneity, scale, disturbance, and recovery. In: F. Dallmeier and J. Comisky (Editors), Forest Biodiversity in North, Central, and South America and the Caribbean: Research and Monitoring. Parthenon Group, New York, pp. 293–322. Woolbright, L.L., 1991. The impact of Hurricane Hugo on forest frogs in Puerto Rico. Biotropica, 23: 462–467. Woolbright, L.L., 1996. Disturbance inﬂuences long-term population patterns in the Puerto Rican frog, Eleuthrodactylus coqui (Anura: Leptodactylidae). Biotropica, 28: 493–501. Wunderle Jr., J.M., 1995. Responses of bird populations in a Puerto Rican Forest to Hurricane Hugo: the ﬁrst 18 months. Condor, 97: 879–896. Wunderle Jr., J.M., Diaz, A., Velazquez, I. and Scharron, R., 1987. Forest openings and the distribution of understory birds in a Puerto Rican rainforest. Wilson Bull., 99: 22–37.

657 Wunderle Jr., J.M., Lodge, D.J. and Waide, R.B., 1992. Short-term effects of Hurricane Gilbert on terrestrial bird populations on Jamaica. Auk, 109: 148–166. Yih, K., Boucher, D.H., Vandermeer, J.H. and Zamora, N., 1989. Efectos Ecol´ogicos del Hurac´an Joan en el Bosque Tropical H´umedo del Sureste de Nicaragua a los Cautro Meses: Posibilidades de Regeneraci´on del Bosque y Recomendaciones. Centro de Investigaciones y Documentaci´on de la Costa Atl´antica (CIDCA), Managua, 97 pp. Zak, J.C., Sinsabaugh, R. and MacKay, W.P., 1994. Windows of opportunities in desert ecosystems: their implications to fungal community development. Can. J. Bot., 73: S1407-S1414. Zimmerman, J.K., Willig, M.R., Walker, L.R. and Silver, W.L., 1996. Introduction: disturbance and Caribbean ecosystems. Biotropica, 28: 414–423.

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Chapter 28

HOW HUMANS RESPOND TO NATURAL OR ANTHROPOGENIC DISTURBANCE C.J. BARROW

INTRODUCTION

This chapter reviews human responses to ecosystem disturbances. These disturbances may be natural, or caused by human activity, or by a combination of both, and responses are often complex and difﬁcult to predict. After reviewing human response to ecosystem disturbance I discuss the need for it to be better managed and the potential of impact assessment. Some ecologists believe that, given long enough, an ecosystem will reach a steady-state with a web of interrelationships between organisms and environment that allow adjustment to all but serious disturbances (Clements, 1916). Even if that is true, at least some areas might not be in a steady state. They may be developing toward such a condition, or have been upset from one by natural events, anthropogenic change, or both (Kershaw, 1973; Park, 1980; Hill, 1987). A steady state may not be static and simple to work with; it could be cyclic, subject to erratic change about a norm, or gradually shifting. Therefore, it makes sense to learn how to deal with the fact that the majority of ecosystems are not in equilibrium and, in the majority of cases, are disturbed and/or are being disturbed. Ecosystem disturbance may have no effect on human affairs, may constrain development, or offer opportunities. Crudely, there are three human responses to ecosystem disturbance: move, adapt, or “die”. Humans have responded to cyclic change with transhumance (seasonal migration up or down a mountain or to and from areas with better rainfall), exploitation of food sources usually ignored, or various other “coping strategies”. Recurrent erratic change has resulted in opportunism, nomadism, insurance strategies, or migration. Gradual ecosystem disturbance either permits virtually unconscious adaption, results in a failure to

perceive change, or reaches a point where things are obvious and human activities must be modiﬁed. Failure to notice or accept the threat of change until it is too late to adapt has doubtless led to the decline of a number of past civilizations and could affect modern societies. Increasingly, resource exploitation and other activities, notably pollution and “free-trade” developments, are making it impossible for ecosystems to be developed in isolation. The later part of the twentieth century has witnessed increasing trans-boundary (i.e., across national borders) concern and interdependence as economies are more and more interlinked economically and as developments including global warming, acid deposition, and ozone depletion have appeared. Many animals including humans seek to maximize their exploitation of the environment through territoriality. For humans the global nature of ecosystem disturbance demands reconsideration of concepts such as sovereignty and what constitutes a nation’s “internal affairs”.

ECOSYSTEM DISTURBANCE

Ecosystems may be broadly classiﬁed by degree of human disturbance: (1) Natural ecosystem, unaffected by human interference, increasingly rare and therefore in need of conservation. (2) Modiﬁed ecosystem, affected to some extent by human activity; most ecosystems are now in this category. (3) Controlled ecosystems, human interference, by accident or design, plays a major role in maintaining

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function (e.g., intensive aquaculture, hydroponics). The most extreme example would be the “Biosphere 2” facility which, it was hoped, would operate in isolation from Biosphere 1 (the Earth’s atmosphere, climate and nutrient ﬂows) (Allen, 1991). Today, controlled ecosystems are relatively rare, but will probably increase if sustainable development (see later discussion) is pursued. Modiﬁed ecosystems can be subdivided into those established by intention, for instance, through agriculture – the major cause of ecosystem disturbance at the world level, and likely to increase as human population grows; and those resulting inadvertently (e.g. caused by pollution or ﬁre). Monitoring the development of modiﬁed ecosystems and preventing dangerous or undesirable conditions presents a major challenge. The challenge has been recognized by those concerned with sustainable development. There is much debate over the precise meaning of “sustainable development”, and it is unlikely there will ever be full agreement on the concept (Barrow, 1995a). A frequently cited deﬁnition of sustainable development is “Development that meets the needs of the present without compromising the ability of future generations to meet their own needs” (World Commission on Environment and Development, 1987). The pursuit of sustainable development requires that attempts to alleviate poverty, improve world trade, advance technology, etc., be linked to environmental concerns. Unfortunately, to develop principles and concepts is a long way from establishing and maintaining practical approaches to sustainable development that work in the “real” world. Ecosystems subject to unintentional or deliberate disturbance can be divided according to their resiliency (Holling, 1973) into: (1) Fragile, those which are easy to disturb and recover with difﬁculty. (2) Robust, those which are difﬁcult to disturb, but which recover easily. (3) Not stable but resilient, those which are easy to disturb, but which recover easily. (4) Stable but not resilient, those which are difﬁcult to disturb, and which recover with difﬁculty. The ﬁrst of these four situations poses a particular challenge, because the ecosystems are vulnerable to unintentional disturbance and require careful management to prevent and to try to rectify disturbance. Examples of such ecosystems are alpine vegetation or semi-arid savannah. Some humid grasslands or scrublands would ﬁt category (2) and temperate grasslands dominated

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by plants which seed annually fall into category (3). Category (4) includes formations where plant species are robust but face problems in recolonizing after disturbance, or where vegetation disturbance results in edaphic changes that prevent recolonization (an example of the latter are grassland areas of Vietnam where defoliation of forests led to waterlogging which prevented the regrowth of anything other than rank grassland). Situations (1) and (4) present such difﬁculties to those attempting restoration following disturbance that it is better to try and convert them to a different type of ecosystem. For example, a hillside in a humid tropical environment, disturbed by non-intensive agriculture, may degrade into unproductive land which has to be abandoned. The resultant vegetation cover is likely to be less species-rich than the original forest, making adequate recovery of a pre-disturbance situation unlikely unless the cleared area was only small. Farmers may then move on to degrade a new patch of land, thus continuing the process of disturbance and degradation. Alternatively, some types of land-use (e.g., irrigated rice terraces) may have sustained intensive agricultural production for generations, the process of ever-expanding disturbance thus being reduced or halted (Geertz, 1963). Clearly, if environmental damage is to be limited and human welfare improved, there is a need for skilled monitoring and extension services which encourage the best approach for a given situation. Such things are far from perfected in either developed or developing nations. It is often very difﬁcult to model and monitor with a view to predicting and assessing disturbance of complex systems which operate at different scales, perhaps in parallel and often with non-linear behaviour (Biswas et al., 1990; Spellerberg, 1991; Jakeman et al., 1993). It is therefore vital to model and monitor ecosystems at different scales (local, regional, national, etc.); but to do this effectively requires better indicators and improved approaches. Some problems result when one or more critical thresholds are passed, and in this case it may be practical to monitor selected parameters to give early warning before this happens. Such a simple situation is rare. Increasingly, damage results from insidious cumulative effects, some or all of them indirect (i.e., they lie at the end of a difﬁcult-to-monitor “chainof-causation”) (Barrow, 1997). Typically, planners and managers recognize disturbance after it has happened, and then seek “technological ﬁxes” which are often costly and ineffective. Monitoring leading to early

HOW HUMANS RESPOND TO NATURAL OR ANTHROPOGENIC DISTURBANCE

recognition of problems, so that preventive action can be taken, is likely to be cheaper and much more effective. In order that ecosystem disturbance be minimized through preventive action, monitoring and impact assessment should be better integrated with planning and policy-making, and become more forward-looking. If development decision-making looks further into the future and beyond the local situation, ecosystem disturbance might be reduced and human responses become more rational (Graham Smith, 1993). According to Von Liebig’s “Law of the Minimum”, whichever resource or environmental parameter necessary for survival is most unfavourable is the critical or limiting one (thus a locality may be degraded beyond human use at the point when, for example soil, water, fuelwood, or living-space becomes critical). The critical limit may be reached gradually or suddenly, once or repeatedly in a random or cyclic manner. Reaching a critical limit may be obvious or difﬁcult to recognize, and once a limit is exceeded damage may be serious and irreversible. To cope with these challenges, analysts are turning to studies of chaos and catastrophe theory, mathematical approaches which seek to recognize where a process of gradual, possibly complex, change reaches a threshold beyond which there is a sudden shift to a different state. For example, catastrophe theory may be used to model a pasture that may withstand grazing for a long time, gradually becoming weakened and depleted in species, until a threshold is reached where catastrophic soil loss following rain prevents regrowth (Barrow, 1995b). Where disturbance has been unforeseen, there may be little chance of adaptation; even moving people, biota, or infrastructure may be too hurried to be safe or satisfactory. Natural ecosystems seldom behave simply, so predicting the effects of disturbance is difﬁcult – doubly so when the effects of human population growth and the ﬁckle behaviour of people are factored in (Gilbert and Braat, 1991; Barrow, 1997). On account of human population growth, there is a world-wide decrease in the time available for developing satisfactory responses to ecosystem disturbance, coupled with a growing threat that such disturbance will happen. Added to these problems are lack of funding, and attitudes which slow or prevent responses. Study of human response to ecosystem disturbance is important; it is a ﬁeld which has been relatively neglected, and deserves top priority. Ecosystems subject to unintentional or deliberate disturbance, in addition to being divided according to

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resiliency, may be grouped into those where change is gradual or sudden, gradual and then sudden, sudden and then gradual. Faced with gradual ecosystem disturbance humans often do not respond, assume there is no need for haste, or fail to notice the change. Few of the people living around the Mediterranean today perceive land degradation, but if shown the richer forests of, say, 8000 years ago they would probably be shocked. Sudden but predictable ecosystem disturbance (recurrent storms, ﬂoods, brushﬁres, earthquakes) can be planned for, but often such planning is neglected. Unforseen disturbance (a ﬂood where none had been known before, a freak storm, etc.) is very difﬁcult to avoid. Sudden natural disturbance of ecosystems has long been of interest to catastrophists, risk and hazard assessors, relief and civil-defence bodies. Recently, attention has focused on “time-bomb effects” – processes which lead to more-or-less gradual, often insidious, accumulation of a threat (chemical or biological), the release of which can be “triggered” suddenly by a natural event or human action. For example, a fertilizer or a pesticide used on farmland may accumulate in the soil, perhaps bonded to clay minerals, and cause no real harm. A warming climate, acid precipitation, or a change of land use may break the bonding, triggering an “explosive” release of the pollutant to soil, springs and streams, where it can be harmful (Stigliani et al., 1991; Zurek et al., 1994). Incautious land use such as land drainage or clearance of woody vegetation can sometimes trigger “runaway” effects, for example: formation of drainageinhibiting pans; the accumulation of salts within the soil or on its surface; the formation of acid-sulphate soils (this happens where waterlogged soils have suffered reduction of their sulphate ions to sulphides under anaerobic conditions; when they are drained, aerobic conditions allow re-formation of sulphates, making the soil highly acidic) (Greenland and Hayes, 1981; Lal et al., 1989). Costly chemical amelioration or selective replanting may partially reverse this degradation; but such remedies are usually difﬁcult, and consequently large areas of the world remain degraded. Are some localities more prone to disturbance? Risk-assessment specialists can establish the vulnerability of a given locality to disturbance by natural hazards (Whyte and Burton, 1981; Bartell et al., 1992; Carpenter, 1995). This can be linked to mapping

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of vulnerable soils and other relevant factors. For example, regions subject to earthquake and volcanic activity, frost, drought, storm, avalanche, locust damage, tornados, or at risk from rising sea-levels and close to potentially hazardous human activities, or even vulnerable to military or terrorist activity, are relatively easy to identify. Areas of vegetation that are periodically vulnerable to brushﬁres, and regions with growing human and livestock populations, can be recognized as potential problem areas. The use of geographical information systems (GIS) has made such risk assessment easier and more reliable. A computer data-base of environmental and/or human parameters can be regularly updated from data gathered by a geographical information system. The user may interrogate the data-base to construct an image or map which displays selected parameters, sometimes updated in real time as it is watched (Jones, 1996). Less easy to predict are the impacts of possible future global environmental change, of technological innovation, and social and economic development. In some situations it is possible to determine the probability – the risk – of a particular outcome, but increasingly planners and administrators face uncertainty – they have no idea what outcomes are possible, let alone probable (Funtowicz and Ravetz, 1991, 1994). Human behaviour is particularly ﬁckle; a change in fashion may have considerable effect on environmental use; the social sciences have been notably unsuccessful in predicting social impacts during the last century! Land particularly likely to suffer degradation includes areas where vegetation is under periodic stress (e.g., seasonally dry areas, upland areas), cold regions, and areas subject to the demands and pressures associated with urban growth. Over the very long term there is unlikely to be any locality wholly secure from some form of disturbance. However, in the short to medium term, the susceptibility of localities to disturbance varies greatly. It makes sense to adapt activities to the locality, in order to minimize the risk of disturbance and degradation. Given the imprecision of predicting future threats and the “random” nature of events such as an earthquake or a catastrophic impact of the Earth with a meteorite, comet or asteroid (Huggett, 1989), it would be wise to duplicate and disperse important infrastructure or facilities for conservation of biodiversity. Clearly, development strategies and environmental management need to be adaptive (Holling, 1978).

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Poverty and ecosystem disturbance There is a growing literature on poverty (often closely linked to population growth) and environmental degradation (e.g., Leonard et al., 1989; Ghimire, 1993). This is a mineﬁeld for the incautious and for simplistic Malthusians, who argue that population growth invariably leads to environmental degradation (Dupˆaquier and Grebenik, 1983). Poverty can restrict responses to environmental disturbance, and it may trigger the disturbance. But poor people may be more willing to adapt and are more resilient than the afﬂuent; the impoverished make far less demand per capita on resources, and some poor people have developed adequate sustainable ways of life. There are regions where population is sparse and environmental degradation is rife, often at the hands of a few rich entrepreneurs, and it is possible to ﬁnd situations where the number of poor people has grown but environmental conditions have improved (Tiffen et al., 1993). Any attempt to blame the poor for environmental disruption should be treated with caution.

RESPONSE TO DISTURBANCE

Outlook and ethics help determine perception and responses to disturbance (Hardin, 1968; Axelrod, 1984; Costanza, 1987). Different groups of people may view the same process at different scales and react in different ways. Response to disturbance can be complex, and frequently involves conﬂicts of interests. People are more likely to respond to something that is manifest and understandable, than to a threat of which they are little aware. (Fear may be aroused by lack of understanding, leading people to action.) Some people may be fatalistic and hesitate to counter a threat or a problem, whereas another group will cooperate and make sacriﬁces to solve a challenge, especially if there is charismatic leadership. A few are genuinely proactive and rational. The most usual outlook is basically utilitarian – the individual and his or her family come ﬁrst. Apart from a few “deep-green” environmentalists, who have eco-centric ethics (i.e., they put nature before human needs) the likely response is anthropocentric (human welfare before environment). Behavioural psychologists, risk assessors, social anthropologists, and others have studied perception of risk and response to disturbance; but there is a very long way to go to develop reliable prediction. When

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in control of resources, people usually give priority to short-term, personal gain as against longer-term gains for themselves, their descendants, and strangers. In many cases, people in general may have little control over resources, and decisions are made by special interest groups or an elite. The poor in particular have little control over resources, and must take what they can get, giving day-to-day survival priority. It is not unusual to come across poor people forced by circumstance to degrade the environment, who are quite aware they are doing this and thus destroying future opportunities. Changing attitudes to environment and development Response to disturbance to a large extent reﬂects attitude toward environment and development. It is possible to recognize a number of historical phases characterized by different attitudes, although there are some regional and national variations. In a primitive phase covering human development up to as late as 5000 BP, humans were low in numbers and largely nomadic. Pollution was not a problem to the low numbers of mobile people. However, that there was no disruption of the environment in this phase is an “Arcadian myth” – forests were cleared, grasslands were burnt, and it is virtually certain that many animal species were exterminated by hunting (Barrow, 1995b). By roughly 5000 BP the ancient Egyptians, Chinese, and Greeks saw a transition to a classical phase, by which time control over resources and societies had increased considerably. Growing populations were less nomadic, and a number of large cities appeared. Religion and laws maintained some control over resource use, but there was little rational investigation of the world in Europe before the sixteenth century (elsewhere, for example, in China and India and in parts of the New World there had been study and exploration before the ﬁfteenth century). Unquestioning fatalism, superstition, and rule by divinely-ordained monarchs gave way to the idea that humans could determine their own destiny, partly as a result of the Renaissance (between the fourteenth and sixteenth centuries). In the modern phase (roughly from the seventeenth century to the 1960s) the pervasive attitude was to seek to understand and dominate nature through specialism-oriented (reductionist) science and organized management (whether the system was capitalist or communist). Exploitation of resources and proﬁt

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came ﬁrst, and problems that resulted were ignored or dealt with later. Many people in a range of disciplines (poets, novelists, artists, philosophers, geographers, mathematicians, etc.) recognize a post-modern phase, starting in the 1960s, or a little earlier, which is supplanting the modern phase. Many would claim that in the post-modern phase more people are aware of the need for multidisciplinary, even holistic, approaches to the environment and to development (Glaeser, 1988; Harvey, 1989). Modern holism is not universally welcomed and is difﬁcult to deﬁne, but implies acceptance that “the whole is greater than the sum of its parts”, and that to understand the totality of problems is better than to research the component parts in detail (Smuts, 1936; Atkinson, 1991). In the last few decades there has been a shift to the promotion of “grassroots” participation, rather than trust in management decisions. Ethics have been stressed that emphasize stewardship, husbandry, and sustainable development. One example is the “polluter-pays-principle” (the idea that those responsible for pollution bear the direct and indirect costs of remedy, rather than the “polluted” and/or third parties). The breakdown of traditional resource use a growing cause of ecosystem disturbance Established systems of resource use may work for long periods and then fail. For example, irrigated agriculture along the Euphrates and Tigris Rivers functioned for centuries before failing (Jacobsen and Adams, 1958); rain-fed agriculture in North Africa provided grain for ancient Rome for centuries before being abandoned (Gischler, 1979); the Maya and Inca peoples managed sophisticated agricultural and communication systems for a very long time before they were conquered (Barrow, 1987). Some of these systems may have failed as a result of natural causes, others as a consequence of social change or warfare, few have ever recovered. If stable and complex institutions are needed to manage ecosystem use it may not take much to disrupt them, either from within or through external factors (Rees, 1990; Smith, 1992). During the twentieth century there has been a revolution in human ability to exploit resources, mainly in developed countries. Many traditional responses to ecosystem disturbance which allowed relatively stable, long-term exploitation of the environment have broken down, especially in developing countries. Traditional land-use systems, and strategies for ﬁshing or coping

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with drought, which functioned for a long time have ceased to function effectively in many countries. The main problem seems to be exposure to modern “development pressures”, which include: warfare and civil unrest, opportunities for migrant labour, decline of traditional values, commercialism, new communications, restrictions on movement of people, and veterinary and health-care provisions. Unrest has damaged the environment directly through “scorchedearth” activities, and indirectly by disrupting land management, and in Africa at least it is a signiﬁcant cause of ecosystem damage. When traditional ways fail, there may be little chance of replacing them with an effective modern alternative. Throughout the world people now aspire to material living standards which, even if there were no population growth or pollution problems, would be beyond available technology to achieve (Meadows et al., 1992). This is increasingly placing the environment and societies under strain. One might say that “consumerism” (demand for material things which are not vital for adequate survival) has become a major cause of environmental disruption (Barrow, 1995b). If this is so, the problem is that people must change their aspirations and ethics to counter ecosystem disturbance, and such changes are likely to be difﬁcult. Range of choice when ecosystem disturbance happens Responses to threats, proposals, and opportunities are seldom rational or predictable, although they tend to emphasize individual survival or generation of wealth. To respond to a disturbance a person must perceive it, and the form of that perception is highly variable. Societies consist of a number of component groups, differing in age, sex, and social class. Even within these groups there may be a wide range of individual responses to disturbance. For a number of reasons, people fail to learn from repeated yet similar disturbances (e.g., storms, ﬂoods, droughts). People may fear unlikely events that are less of a threat than something they accept (the motor car is usually an under-estimated threat) and some may have views which encourage them to be over-optimistic on the one hand, or fatalistic on the other. For some disturbances there is little or no choice of response, for others several, as in the case of recurrent bushﬁres; for instance, people may maintain ﬁre-breaks around their property, may invest in ﬁre-watching

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and ﬁre-ﬁghting services, may relocate to safer areas, may employ ecologists to prepare and conduct ﬁreprotection strategies (removal of dead biomass by harvesting or controlled burning); may modify building materials and styles to give better ﬁre resistance, or may do nothing. Ownership patterns greatly affect response. Those who have no formal ownership of a resource are unlikely to be listened to, or to be compensated if there is disturbance. Those with weak ownership title or limited holdings may lack the resources to resist disturbance, and may easily succumb. In Ireland before the twentieth century, some tenants reputedly spent the ﬁrst one-third of their tenancy rectifying the previous tenant’s neglect of the land, one-third farming wisely, and the last one-third maximizing returns and damaging the land in the process, because it was unlikely the tenancy would be renewed. Security of tenure can be important in ensuring the land owner (or other land user) develops “ties” and manages for the long term. However, there are situations where resource users act unwisely and some control or even threat of loss of tenure may be needed (Powell, 1994). Good management of shallow groundwater resources is difﬁcult to enforce without some form of licensing and policing system. The ideal is to offer reward and security for good stewardship, and threats to discourage bad stewardship, and this must be accompanied by monitoring and enforcement. It is pointless having a law to prevent settlers clearing more than half their land of forest if the situation is not policed – the settlers may simply sub-let to others who then clear it completely. It is vital to understand the social roots of ecosystem disturbance and responses to it (Blaikie, 1985; Garcia-Perez et al., 1995). Greed, rather than rational behaviour, may affect actions. For example, in parts of Spain with a long tradition of sustained agriculture, farmers have been seeking subsidies from the European Union [the European Union (EU) is the expanded European Community, within which ﬁnancial controls and incentives are used to try to shape the member states’ agriculture] for certain crops, have lost pride in good land management, or have been attracted away by the lure of part-time, off-farm employment (Faulkner, 1995). Control of proﬁt or subsidy for desirable actions, and possibly farmer education, will be required to counter unwanted changes. Similar observations were made by Edmonds (1994) in a review of trends in China’s environmental degradation and remedial actions.

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Timing of response when ecosystem disturbance threatens The speed of response can be crucial. Two broad stages can be considered in terrestrial ecosystem disturbance: (1) land degradation (Barrow, 1991; Johnson and Lewis, 1995); and (2) “desertiﬁcation” (Mainguet, 1994; Thomas and Middleton, 1994). If an appropriate response is made during stage (1) it is possible to halt degradation and, perhaps, cause a partial or complete recovery, or even improvement on original conditions. If a response is not made; and (2) is reached, damage is probably too difﬁcult or costly to reverse, although a halt to the degradation might be possible. In practice, the situation may be more complex if ecosystem disturbance creates opportunities: loss of mangroves may constitute environmental degradation (reduced coastal protection, loss of mangrove-swamp biodiversity, and loss of sites for marine animals to feed and reproduce), but it might also offer the opportunity for real-estate investors or prawn-pond developers to proﬁt. Unless the conservation lobby or other supporters of mangroves have a strong voice, special-interest groups are likely to resist mangrove protection. This sort of conﬂict can be even more complex, as disruption creates further opportunities for exploitation and further disruption. For example, when road builders (with strategic motives or interests in logging or oil) improve access to forests, squatter settlers are able to penetrate along the roads and to clear land for shifting cultivation; ranchers follow, getting title for the land occupied by squatters, but waiting (with or without the latter’s agreement) for the cultivators to clear the forest enough to facilitate the establishment of pasture; an industrial corporation may follow the ranchers, buying degraded pasture and developing it as eucalyptus plantations to be used for paper-pulp production (Lohmann, 1990). Thus, there is a chain of developers, none likely to support a return to the original undegraded forest. To respond adequately to ecosystem disturbance demands a more preventive approach; to react after a problem is well developed is too late. There must be better monitoring to support improved standards by which to measure and compare. The development of environmental impact assessment (EIA) and related approaches since 1970 has some potential to help planning and decision-making, but generally needs much improvement before it is fully effective (Barrow, 1997).

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Patterns of response to ecosystem disturbance Faced with problems caused by ecosystem disturbances, people may: (1) Ignore them, perhaps hoping environmental conditions and food or commodity production will not decline too much and will then stabilize. This is probably what has happened over large areas of the land fringing the Mediterranean, where an apparent steadystate “natural vegetation” – scrubland and thin soils replaced forest or woodland with richer soils at some point before or during the “classical phase” (Dale and Carter, 1954; Thirgood, 1981). (2) Try to manage a transition to something better. Geertz (1963) explored different approaches to modifying the environment for agricultural use. He recognized a “canny imitation of nature” – shifting cultivation which mimics the continuous vegetation cover and species diversity of natural forest. This approach protected the soil against rainfall that could erode it, and gave the farmer a diversity of crops ensuring some security against pests, diseases and drought. The modern development of this approach would be various forms of tolerant forest management, where a tree canopy is maintained and useful natural species are encouraged (Anderson, 1990; Nepstad and Schwartzman, 1992; Barrow, 1995b). A similar approach in drier savannas is the rotational rain-fed cultivation of grain linked with the collection of gum arabic from acacia trees (Acacia senegal and A. seyal), still practised in parts of the Sudan (Olsson, 1984). The conversion of river ﬂood-plains to water-meadows, or areas of ﬂood-recession irrigation which make use of natural deposition of silt to sustain cropping, could also be counted as “mimic nature” approaches. Geertz (1963) recognized a somewhat different response that could be considered a “bold re-working of nature” in the form of Javanese irrigated terraces for cultivation of rice (Oryza sativa), which were a drastic change from natural forested slopes. The rice terraces could then be maintained indeﬁnitely with few inputs and no further environmental damage. Much modern agriculture has also been “a bold re-working”, but it depends on agrochemicals, fuel for farm machines and large capital investments. In the long term these inputs may be difﬁcult to provide, the agrochemicals may well cause pollution, and soil degradation often takes place. Biotechnology may create new possibilities. For

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example, fermentation tanks producing single-cell protein, vegetable oils, cocoa, etc., could substitute for ranching and plantations (Litchﬁeld, 1989) and thereby reduce soil erosion, the use of agrochemicals, and loss of biodiversity as land is converted from a natural to a farmed environment (Goodman and Redclift, 1991; Mannion, 1992). (3) Resist the disturbance and seek to maintain established ways. There are two forms of resistance: passive and aggressive. The former ranges from nonviolent protest and obstruction to group court action and political lobbying, with or without the aid of sympathetic non-governmental organizations (NGOs). A number of the native peoples of Australia and the Amazon Basin have turned with considerable success to legal action, cultivation of supportive media coverage, and representation on government bodies (Cummings, 1990). Aggressive resistance has followed ecosystem disturbance in many countries, though it is often dismissed as banditry or terrorism. A notable example recently was the protest in the Niger River delta region against government and corporate exploitation of resources which had led to environmental disruption (Moffat and Linden, 1995). (4) Abandon the land or resource being used. This can develop in a number of ways. Poor people may become trapped in a vicious spiral. Lacking resources to escape or rectify a problem, their poverty increases and they may struggle to survive, further degrading their resource base. The process can be easily triggered by a family debt, illness that incapacitates a farmer, or by drought or some other natural or human disaster. If no aid is given to help the victim break free, the result is enforced relocation or change of means of livelihood (e.g., poor farmers may become migrant labourers) or destitution (Chambers et al., 1981). Faced with a deteriorating local situation and/or attractive opportunities elsewhere (a marginalization process), able-bodied and adventurous members of a society may move to ﬁnd employment, perhaps seasonally or as a temporary measure. The result is selective depopulation, with the remaining labour pool unable to manage the land adequately, which can result in land degradation. Sometimes abandonment takes the form of withdrawing from part of the used environment such as uplands, valley-bottoms, or marginal pasture. Another 1

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possibility is the cessation of some activity. For example, those farming and ﬁshing may turn to ﬁshing only, or farmers may turn to a high-return activity requiring less labour for a given proﬁt – narcotics or tourism (e.g., renting caravan and trailer camps to tourists). Sometimes the new activity may take over, so that other farmland is neglected. In some developed countries the government has encouraged or enforced abandonment of land to counter pollution or excess production of certain crops. In Europe, set-aside programs compensate farmers for converting farmland to less intensive uses in order to reduce pollution by fertilizers or pesticides, and possibly improve sustainability of resource use (Pretty, 1990). In the United States, the Conservation Reserve Program (initiated in 1985) seeks to reduce soil erosion by encouraging set-aside. (5) Remain ignorant of the problem. There have been situations where the response to an environmental problem is ofﬁcial inactivity or secrecy. Fearing a blight on real-estate values or damage actions from residents, an authority may be tempted to neglect documentation or even conceal land contamination or the risks attached to some activity. Delayed collection and release of information makes it difﬁcult for people to react. There have been land-contamination incidents in several countries where illness has prompted people or non-governmental organizations to probe for information, at which point the authorities may begin to act. Examples include the Lekkerkerk disaster in The Netherlands or the Love Canal disaster in the United States1 (Crump, 1991; Newson, 1992). (6) Become involved in ofﬁcially-supported relocation. The relocation may be forced and assisted by the authorities or voluntary and assisted (Hansen and Oliver-Smith, 1982). Assistance may simply take the form of compensation payments or grants for relocation, or it may be a comprehensive package of planned relocation assistance. A major cause of such forced relocation has been the construction of large dams and reservoirs. Relocation can also be caused by accidents such as the partial melt-down of a nuclear power station that occurred at Chernobyl in the Ukrain (International Atomic Energy Authority, 1991). Relocation may be sponsored by governments in order to settle territory for strategic reasons, or as a development or welfare measure. This selective or

In both of these cases, domestic housing was built on land which had been previously contaminated by industrial pollutants; gradually residents became seriously ill and the contamination was ﬁnally identiﬁed, necessitating evacuation and compensation.

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voluntary relocation, with ofﬁcial support, is typiﬁed by Indonesia’s Transmigration Programme, the settlements under Malaysia’s Federal Land Development Authority (FELDA), and similar schemes in Brazilian Amazonia (in the 1970s), Tanzania, and several other countries. Some of these efforts have had markedly poor results, the ofﬁcial support failing to establish adequate sustained livelihoods for those relocated or to prevent people from becoming restless and moving on (Moran, 1981, 1983). (7) Relocate without ofﬁcial support. Some of these relocatees do so voluntarily (e.g., squatter settlers attracted by hopes of better livelihood). Increasing numbers are forced to relocate without ofﬁcial support as refugees or marginalized people. Marginalization and common resources The process of marginalization has generated controversy and a large literature (Baker, 1984). It is the movement of people into high-risk, often lowproductivity environments. They may be driven to do this by civil unrest, population growth, or the activities of large corporations. Alternatively, they may be attracted by hope for better opportunities, or have stayed in a locality which has become marginalized around them. Technical change, new demands for a product, new road or rail communications, and other factors may drive or counteract marginalization. Common resources, ecosystems, or speciﬁc resources exploited by a number of people but not owned or controlled by an individual, present particular problems when degraded. There may be a tendency for use of these common resources to continue while responsibility for remedial action is debated and the authority to take it is disputed. Hardin (1968) tried with his “tragedy of the commons” essay to establish the fate of a common resource under increasing demand (e.g., due to population growth), concluding that individual users would seek to maximize short-term beneﬁts, with the overall effect of serious resource damage. Hardin’s thesis is often questioned on the grounds that use of common resources is seldom a “free-for-all”, and that communities have generally evolved controls, are not ﬁxed in their response to challenges, and can learn from mistakes (Boser¨up, 1990; Harrison, 1992). Nevertheless, many would argue that the present-day over-exploitation of natural resources and pollution without care for future generations, ﬁts Hardin’s thesis.

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Loss of access to common resources is a worldwide problem that appears to be a growing cause of breakdown of traditional livelihoods, poverty, social unrest, and often relocation (either forced on the people by declining quality of life or required by authorities supporting those who have gained control of the common resource). An example would be the takeover of land traditionally used by people for grazing and other beneﬁts by a company seeking to grow Eucalyptus species for paper pulp. The government has granted a license and the traditional users are in a weak position to ﬁght because they have no formal, written legal rights to the land. [For an introduction to this form of marginalization see: The Ecologist (1993)]. Those forced to relocate by environmental factors (e.g., storms, or drought, soil degradation, ﬂooding, or volcanic disasters) have been termed eco-refugees. To date their numbers have been limited, although in the American mid-west Dust Bowl during the 1930s considerable numbers of people abandoned land. In the future, North Africa seems likely to generate ecorefugees as population growth places stress on available land leading to soil degradation, and as water supplies fail to meet demand. If global environmental change leads to altered rainfall patterns and rising sea-levels there may be very large numbers of eco-refugees from low-lying nations such as Bangladesh (McGregor, 1993; Myers, 1993; D¨oo¨ s, 1994). Management of response to ecosystem disturbance A group may have evolved traditional responses which are initiated, often with little thought or question on the part of the people, perhaps prompted by tribal elders or witch doctors. Some might argue that in modern society bureaucrats and scientists have taken over a similar role. Cooperative organizations may assist people in situations prone to natural disturbances, with difﬁculties in transporting or marketing products, or with ﬂuctuating market prices for products. There are societies where people have called a national referendum, or NGOs have been able to insist on the adoption of popular responses. For example, Sweden initiated a moratorium on its established nuclear-power program as a consequence of public concern. One challenge is that the management of response to transboundary and global problems, international agreements, and the practice and improvement of international law, tend to be slow.

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Soil degradation is a major threat to human welfare and biodiversity which must be addressed and managed. Serious soil degradation is taking place, and the threat is accelerating in both developing and developed countries. Regions which currently yield the produce which largely “feeds the world” are subject to this degradation. The threat is great but the response has been weak, even in the United States, where painful lessons were learned from the 1930s Dust Bowl disaster (Bonniﬁeld, 1979; Worster, 1979). The results of soil degradation have included salinization, erosion, and loss of organic matter. What has had far less attention is assessing the cause of soil degradation – why people mistreat the land, why there are problems in initiating effective counter-measures. A welcome exception was Blaikie’s (1985) consideration of the political economy of soil erosion, which examined why people failed to perceive and respond to soil damage (Blaikie and Brookﬁeld, 1987). The potential of impact assessment Impact assessment has the potential to alter policy making, planning, and environmental management so as to improve responses to problems, reduce their occurrence, and provide a framework for seeking sustainable development. So far, environmental and social impact assessment has mainly been focused on individual projects, and has tended to be initiated only after a development decision has been made. If impact assessment can be applied more at policy, programme, and planning levels, to integrate planning and environmental management, and to involve people in decision making, there might be far-reaching beneﬁts. Much has to be done to make impact assessment an effective part of planning, decision-making, and environmental management. Currently, it is more of a “bolt-on-tool”, used to convince people, government regulators, and funding bodies that the developer is environmentally responsible. To reduce the control exercised by developers over impact assessment will require greater public involvement, and improved review and audit of EIA by independent authorities. Impact assessment techniques must be improved to cope with complex situations where available data is often poor and the time for study is limited. Promising approaches include adaptable environmental assessment and management, and strategic environmental assessment. The latter is a form of tiered EIA which could link policy to projectlevel planning (and link local to national or even global

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issues) and improve responses to challenges (Therivel et al., 1995; Barrow, 1997).

CONCLUSIONS

With growing human populations in many regions, together with the increasing likelihood of transboundary impacts such as global environmental change or widespread pollution, can there be any optimism about human responses to ecosystem disturbance? The human–environment relationship is not simple, and there are situations where, in spite of population growth and environmental difﬁculties, land degradation has actually been reversed while food and commodity output has increased. Boser¨up (1965, 1981) studied situations which led her to suggest that, under the right conditions, population growth could “lift” agriculture from poor productivity causing environmental damage to much more intensive production with less environmental damage. Other studies have supported her conclusions (Little and Horrowitz, 1987; Tiffen, 1993; Tiffen et al., 1993). In Australia, where population densities are relatively low, land degradation has become widespread during the last 200 years, but agricultural output has increased (Chisholm, 1994). It is probable that in Australia and other developed countries, the use of fertilizers over the last few decades has more than compensated for soil degradation. But this response may have its limits, and those limits may be reached soon if care is not taken. There will come a point with soil degradation where a “plateau” of production is reached and further application of fertilizer leads to little or no improvement. Without a technological breakthrough output may plummet. Decision-makers have always tended to seek proof of threats before they dare act; otherwise they face the charge of being rash if the problem they tried to prepare for fails to materialize. Yet many disturbances to ecosystems are difﬁcult, costly, or even impossible to deal with once they have happened – for example, a dangerous organism or toxic compound may be virtually impossible to remove once it has dispersed. There is thus a conﬂict at the heart of management responses to potential disturbances: a desire to “wait and see” versus the desirability of pre-emptive action. This conﬂict has been discussed in relation to global change, and there is now a growing literature on

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“no-regrets” policies2 , adaptive environmental management, and sustainable development (e.g., Carley and Christie, 1992; Reed, 1995). Given the risks of natural or anthropogenic disturbance in any locality over the long term, it is wise for activities and facilities which are vital for environmental quality (and the associated well-being of humans) to be duplicated in widely different localities and, if possible, dissimilar environments. For example, conservation of a type of vegetation might rely on a number of in-situ reserves, together with gene-bank material stored in secure, “hardened” stores around the world. Those promoting sustainable development would be well advised to adapt to local conditions and to operate at a regional scale of management, with national and ultimately global coordination. There should be enough overlap between material in distant elements of the “mosaic” to provide a chance of recovery if there is a local or regional disaster (Barrow, 1995a). Given the growing disturbances of a transboundary or global nature which local or regional bodies do not have the resources to deal with, national and global coordination is vital.

REFERENCES Allen, J.P., 1991. Biosphere 2: The Human Experiment. Penguin, Harmondsworth, UK. Anderson, A.B. (Editor), 1990. Alternatives to Deforestation: Steps Toward Sustainable Use of the Amazon Rain Forest. Columbia University Press, New York, 233 pp. Atkinson, A., 1991. Principles of Political Ecology. Belhaven, London, 251 pp. Axelrod, R., 1984. The Evolution of Cooperation. Basic Books Inc., New York. Baker, R., 1984. Protecting the environment against the poor. The historical roots of the soil erosion orthodoxy in the Third World. Ecologist, 14: 53–60. Barrow, C.J., 1987. Water Resources and Agricultural Development in the Tropics. Longman, Harlow, UK, 356 pp. Barrow, C.J., 1991. Land Degradation: Development and Breakdown of Terrestrial Ecosystems. Cambridge University Press, Cambridge, 295 pp. Barrow, C.J., 1995a. Sustainable development: concept, value and practice. Third World Planning Rev., 17: 369–386. Barrow, C.J., 1995b. Developing the Environment: Problems and Management. Longman, Harlow, UK, 313 pp.

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Barrow, C.J., 1997. Environmental and Social Impact Assessment: an Introduction. Arnold, London, 312 pp. Bartell, S., Gardiner, R. and O’Neill, R., 1992. Ecological Risk Evaluation. Lewis Publishers, Ann Arbor, USA. Biswas, A.K., Khoshoo, T.N. and Khosla, A. (Editors), 1990. Environmental Modelling for Developing Countries. Tycooly, London, 166 pp. Blaikie, P.M., 1985. The Political Economy of Soil Erosion in Developing Countries. Longman, Harlow, UK, 188 pp. Blaikie, P.M. and Brookﬁeld, H., 1987. Land Degradation and Society. Methuen, London, 296 pp. Bonniﬁeld, M.P., 1979. The Dust Bowl: Men, Dirt and Depression. University of New Mexico Press, Albuquerque, USA, 232 pp. Boser¨up, E., 1965. The Conditions of Agricultural Growth: the Economics of Agrarian Change under Population Pressure. Allen & Unwin, London, 255 pp. Boser¨up, E., 1981. Population and Technology. Basil Blackwell, Oxford. Boser¨up, E., 1990. Economic and Demographic Relationships in Development. Johns Hopkins University Press, Baltimore, USA. Carley, M. and Christie, I., 1992. Managing Sustainable Development. Earthscan, London, 303 pp. Carpenter, R.A., 1995. Risk management. Impact Assessment, 13: 153–187. Chambers, R., Longhurst, R. and Pacey, A. (Editors), 1981. Seasonal Dimensions to Rural Poverty. Frances Pinter, London, 259 pp. Chisholm, A.H., 1994. Land-use changes in a changing world. Land Degradation & Rehabilitation, 5: 153–178. Clements, F.E., 1916. Plant Succession. An Analysis of the Development of Vegetation, Publication No. 242. Carnegie Institute, Washington, 512 pp. Costanza, R., 1987. Social traps and environmental policy. Bioscience, 37: 407–412. Crump, A., 1991. Dictionary of Environment and Development. Earthscan, London, 272 pp. Cummings, B.J., 1990. Dam the Rivers, Damn the People: Development and Resistance in Amazonian Brazil. Earthscan, London, 132 pp. Dale, T. and Carter, V.G., 1954. Topsoil and Civilization. University of Oklahoma Press, Norman, USA. D¨oo¨ s, Bo.R., 1994. Environmental degradation, global food production, and risk of large-scale migrations. Ambio, 23: 124–130. Dupˆaquier, J. and Grebenik, E. (Editors), 1983. Malthus Past and Present. Academic Press, London, 350 pp. Edmonds, R.L., 1994. Patterns of China’s Lost Harmony: a Survey of the Country’s Environmental Degradation and Protection. Routledge, London, 334 pp. Faulkner, H., 1995. Gully erosion associated with the expansion of unterraced almond cultivation in the coastal Sierra de Lujar, S. Spain. Land Degradation & Rehabilitation, 6: 179–200. Funtowicz, S. and Ravetz, J., 1991. A new scientiﬁc methodology for global environmental issues. In: R. Costanza (Editor), The Ecological Economics. Columbia University Press, New York.

2 These are sometimes called “win–win” policies; when a problem is suspected but unproven, the administrator is faced with a situation where he does not want to delay appropriate action, so he adopts an approach which seeks to gain advantages even if the problem fails to materialize. For example, trees planted to “lock up” carbon may still be useful for wood or amenity forest even if global warming does not become a problem.

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Chapter 29

RESTORATION OF DISTURBED ECOSYSTEMS Richard J. HOBBS

INTRODUCTION: WHY RESTORATION?

Most chapters in this volume have examined particular forms of disturbance affecting the world’s ecosystems. These range from natural disturbances such as windstorms and volcanic eruptions through various forms of human-induced disturbance. The recent history of the world has been one of a dramatic increase in the incidence of human-induced disturbances, as humans utilize an increasing proportion of the earth’s surface in some way or another and appropriate an increasing amount of the earth’s productive capacity and natural resources (Vitousek et al., 1986; Goudie, 1990; Turner et al., 1991; Postel et al., 1996). The extent of modiﬁcation and predominance of different types of disturbance in any given area is determined by the prevalent land use. These uses can be summarized in terms of their impacts on the natural ecosystem (Table 29.1). Conservation-oriented use will tend to limit the extent of human-induced disturbance, either by design or by default (for instance in remote or harsh environments). Utilization of native ecosystems is a prevalent type of land use over large parts of the world, and includes rangelands and other pastoral areas, forestry in native forest ecosystems, other extractive uses such as animal or plant harvesting and apiculture, and a growing utilization for recreation and tourism. Replacement, on the other hand, involves the removal of the native ecosystem and its replacement with a simpler system geared towards the production of particular crop or livestock species. Finally, complete removal results from destruction of the native ecosystem, such as occurs during surface mining or urban and industrial development. Each land use entails a suite of deliberate and inadvertent impacts of varying severity on the natural ecosystem. Clearly, the extent of each of these various land

uses varies geographically, and removal is a relatively small category compared to the others. Similarly, any particular area can have a mixture of these land-use categories; for instance, remnants of native ecosystems may remain (in a more or less altered state) within urban areas. Thus, while I shall initially consider the different land-use categories separately here, the interactions between different land uses, either on the same or adjacent areas, are also important. Table 29.1 Major land use categories, based on degree of modiﬁcation of the natural ecosystem 1 1.

CONSERVATION (no deliberate modiﬁcation) - wilderness areas - conservation reserves - uncommitted government-owned land - water catchment

2.

UTILIZATION (exploitation of native ecosystem) - non-plantation forestry - plant harvesting (e.g., fuelwood, medicines, etc.) - pastoralism - animal harvesting - recreation

3.

REPLACEMENT (with intensively managed systems) - agriculture - horticulture - plantation forestry

4.

REMOVAL - urban development - mining - transport - industrial development

1

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Modiﬁed from Hobbs and Hopkins (1990).

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Human modiﬁcation has led in many cases to increasing degradation of ecosystem components, resulting in a decline in the value of the ecosystem, either for production or conservation purposes. This has been met with an increasing recognition that measures need to be taken to halt or reverse this degradation. This is where ecological restoration becomes important.

RESTORATION ECOLOGY

The term “ecological restoration” covers a wide range of activities involved with the repair of damaged or degraded ecosystems (Jordan et al., 1987; Berger, 1990; Baldwin et al., 1994). An array of terms has been used to describe these activities, including restoration, rehabilitation, reclamation, reconstruction, and reallocation. Generally, “restoration” is used to describe the complete reassembly of a degraded system to its undegraded state, while “rehabilitation” describes efforts to develop some sort of functional protective or productive system on a degraded site. In addition, some authors also use the term “reallocation” to describe the transfer of a site from one land-use to a more productive or otherwise beneﬁcial use. Majer (1989a), Aronson et al. (1993a), Jackson et al. (1995), and others discuss terminology and ideas on what comprises restoration ecology. Unfortunately, a stable terminology has been slow to develop, and the above terms are frequently used interchangeably and differently by different authors. Here I shall follow Hobbs and Norton (1996) and use the term “restoration” to refer broadly to activities which aim to repair damaged systems, although the other terms are used as deﬁned above in particular examples. Restoration ecology involves a number of interconnected activities, as summarized in Fig. 29.1. These activities are discussed in more detail below. Figure 29.1 represents an idealized scheme for restoration ecology, and it should be recognized that much current ecological restoration does not conform to this. In particular, clearly stated goals and easily measured success criteria have been hard to develop in much of restoration ecology. Similarly, the identiﬁcation of the key degrading processes or factors limiting ecosystem recovery does not always precede restoration activities.

Fig. 29.1. Processes involved in ecological restoration.

RESTORATION GOALS

Hobbs and Norton (1996) have suggested that ecological restoration is usually carried out for one of the following reasons: (1) To restore highly disturbed, but localized sites, such as mine sites. Restoration often entails amelioration of the physical and chemical characteristics of the substrate and ensuring the return of vegetation cover (Collins et al., 1985; Bradshaw, 1987; Ward et al., 1990; Schaller, 1993). (2) To improve productive capability in degraded production lands. Degradation of productive land is increasing worldwide, leading to reduced agricultural, range, and forest production. Restoration in these cases aims to return the system to a sustainable level of productivity, for instance by reversing or ameliorating soil erosion or salinization in agricultural or range lands (= rehabilitation: Aronson et al., 1993a). (3) To enhance nature-conservation values in protected landscapes. Conservation lands worldwide are being reduced in value by various forms of human-induced disturbance, including the effects of introduced livestock, invasive species (plant,

RESTORATION OF DISTURBED ECOSYSTEMS

animal, and pathogen), pollution, and fragmentation. In these cases, restoration aims to reverse the impacts of these degrading forces, for instance by removing an introduced herbivore from a protected landscape. In many areas, there is also a recognized need to increase the areas of particular ecosystem types – for instance, attempts are being made to increase the area of native woodlands in the United Kingdom, in order to reverse past trends of decline and to increase the conservation value of the landscape (Ferris-Kaan, 1995). (4) To restore ecological processes over broad landscape-scale or regional areas. In addition to the need for restoration efforts within conservation lands, there is also a need to ensure that human activities in the broader landscape do not adversely affect ecosystem processes. There is an increasing recognition that protected areas alone will not conserve biodiversity in the long term, and that production and protection lands are linked by landscape-scale processes and ﬂows (e.g., hydrology, movement of biota). Methods of integrating conservation and productive use are thus required, as for instance in the Biosphere reserve and core– buffer–matrix models (Hobbs, 1993a; Noss and Cooperrider 1994; Morton et al., 1995). Both relate to the situation where a conservation reserve sits within a production landscape, but is surrounded by compatible land uses. There is thus a “core” conservation area, a “buffer” zone of compatible land use surrounding it, and the production “matrix” outside that. Restoration in this case entails (a) returning conservation value to portions of the productive landscape, preferably through an integration of production and conservation values; and/or (b) ensuring that land uses within a region do not have adverse impacts on the region’s ecological processes. Ecological restoration thus occurs along a continuum from the rebuilding of totally devastated sites, to the limited management of relatively unmodiﬁed sites (Hobbs and Hopkins, 1990). The speciﬁc goals of restoration and the techniques used will obviously differ between these different cases. In general terms, however, restoration aims to return the degraded system to some form of cover which is protective, productive, aesthetically pleasing, or valuable in a conservation sense (Hobbs and Norton, 1996). A further tacit aim is to develop a system which is sustainable in the long term. It should, however, be recognized that goals

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for restoration are likely to change as societal values and attitudes change. For instance, in many parts of the world forest ecosystems are now being valued less for production than for non-production values such as those relating to biodiversity and recreation (Williams, 1989; Aplet et al., 1993; Dargavel, 1995), and hence these factors increasingly have to be built into management and restoration strategies. Within these broad general aims, more speciﬁc goals are required to guide the restoration process. Ecosystem characteristics which may be considered when considering restoration goals include (from Hobbs and Norton, 1996): (1) Composition: species present and their relative abundances; (2) Structure: vertical arrangement of vegetation and soil components (living and dead); (3) Pattern: horizontal arrangement of system components; (4) Heterogeneity: a complex variable made up of components 1–3; (5) Function: performance of basic ecological processes (energy, water, nutrient transfers); (6) Species interactions, such as pollination and seed dispersal; (7) Dynamics and resilience: succession and statetransition processes, recovery from disturbance. This set of characteristics is complex, and often individual components are considered as primary goals. For instance, restoration of a mine site may aim to replace the complement of plant species present prior to disturbance, whereas other situations may have the restoration of particular ecosystem functions as a primary aim (e.g., bioremediation of eutrophication in lakes, or the manipulation of vegetation cover to modify water use). Unfortunately, restoration goals are often poorly deﬁned, or stated in general terms relating to the return of the system to some pre-existing condition. The deﬁnition of the characteristics of this condition has proved problematic, since it assumes a static situation. Recent commentators have noted that natural systems are dynamic (Pickett et al., 1992; Pickett and Parker, 1994), that they may exhibit alternative (meta-)stable states (Westoby et al., 1989; Hobbs, 1994a), and that the deﬁnition of what is the “natural” ecosystem in any given area may be difﬁcult (Sprugel, 1991). Indeed, the concept of “naturalness” has been the subject of much recent debate (Elliot, 1982, 1994; Maser, 1990; Gunn, 1991; Cowell, 1993).

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The question of relating restoration efforts to particular reference ecosystems has also been debated (Pickett and Parker, 1994; Aronson et al., 1995). For instance, which time period does one use for deﬁning the baseline ecosystem? A common trend in nonEuropean countries is to use the conditions prior to European colonization as a baseline (e.g., Anderson, 1991), despite the fact that the characteristics of these pre-European ecosystems are often poorly documented, and that irreversible ecosystem changes may have occurred in the meantime. On the other hand, the problem can also arise where the existing ecosystem is taken to be “natural” and used as a baseline, but is, in fact, an artifact of current management practices. An example of the dilemma confronting those wishing to establish baseline or reference systems is New Zealand over the past 1300 years (Cochrane, 1977). At each stage, the ecosystems probably existed in a more or less metastable state. Should the baseline be pre-Maori, preEuropean, or present-day? Much of the ensuing confusion could be avoided by the careful enunciation of speciﬁc restoration goals – for instance, restoration of productive capacity can be assessed relative to the productive capacity of similar, undegraded land. Restoration of compositional, structural, and other ecosystem properties can be related to the known range of those properties, either in recent history if temporal data are available, or within similar less degraded ecosystems in the area. A prerequisite for this approach is the development of a set of easily measurable indicators or ecosystem response variables which can be monitored as the restoration proceeds. Aronson et al. (1993a,b) have developed a set of “vital ecosystem attributes” which they suggested could be used to assess the status and trajectory of a system, and hence to set restoration goals more rigorously. In addition, Hobbs and Norton (1996) suggest using these attributes within the framework of ecosystem health, which is currently receiving increasing attention (Costanza et al., 1992; Rapport, 1992). Assessment of system trajectories toward the recognized range of conditions can then be used as a measure of the success in achieving restoration goals. The ﬁnal box in Fig. 29.1, which relates to monitoring, thus provides a feedback loop which checks progress against the established goals. Examples illustrating the complexity of possible pathways of degradation and restoration are given in Figs. 29.2 and 29.3. These illustrate two different situations where hypotheses have been formulated concern-

Richard J. HOBBS

Fig. 29.2. States and transitions for central Chilean ecosystems (from Aronson et al., 1993b).

ing the possible range of ecosystem states, the likely transitions between them, and potential system thresholds. In each case, a number of different pathways are possible, thresholds are apparent, and restorative methods can lead to different outcomes. In Fig. 29.2, states and transitions for central Chilean ecosystems are illustrated, as described by Aronson et al. (1993b). The “natural” state in this case is considered to be matorral, a Mediterranean-type shrubland/woodland with a high plant diversity. Overuse through grazing and other activities leads to the degradation of this system to a more open savanna-like system known as espinal. Mixed espinal retains a reasonable degree of species and structural diversity, but this declines with increasing degradation to impoverished espinal. A threshold of irreversibility is indicated as resulting from continued overuse of impoverished espinal, resulting in badly degraded espinal. A number of options are available for the restoration of degraded espinal. One is recovery to a simpliﬁed espinal ecosystem, with the potential for continued recovery back to mixed espinal. Alternatively, “reallocation” to a ley or mixedfarming system using annual legumes, or to an array of agroforestry systems, is possible. One may note that in this case, complete restoration to matorral is not considered to be a viable option. In Fig. 29.3, states and transitions for woodland ecosystems in Western Australia are illustrated, as described by Yates and Hobbs (1997). In this case, the “natural” state is an open woodland dominated by Eucalyptus salmonophloia, with a shrubby understorey. With degradation, this system degrades with loss of understorey, soil structural decline, and weed invasion. With continued degradation, loss of adult trees and lack of recruitment eventually leads to the development of a grassland system (annual or perennial). Any

RESTORATION OF DISTURBED ECOSYSTEMS

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Fig. 29.3. States and transitions for woodland ecosystems in Western Australia. Transitions D1–D9 represent system degradation, while transitions R1–R8 represent system repair. See text for details (modiﬁed from Yates and Hobbs, 1997).

of these systems can degrade further to salt land if rising water tables bring soil-stored salt to the surface. Transitions D1–D9 represent system degradation, while transitions R1–R8 represent system repair. Several thresholds of irreversibility are identiﬁed, which require active restoration management to force transitions back to less degraded states. Transitions are thought to be forced as follows: (D1) Light or short-term grazing by livestock and rabbits; (D2) Long-term or continuous overgrazing by livestock and rabbits and invasion by annual weeds; (D3) See D2; (D4) Woodland clearing, gradual senescence of trees from old age and disease, or mass mortality of trees following disturbance; (D5) Heavy livestock and rabbit grazing, or ﬁre; (D6) See D4; (D7) Deposition and concentration of salt on the soil surface as the result of a rising saline water table;

(D8) See D7; (D9) See D7; (R1) Complete or near-complete removal of grazing by livestock and rabbits; (R2) Amelioration of soil degradation, control of annual weeds, gap creation, direct seeding, and replanting of seedlings; (R3) See R2; (R4) Removal of grazing, amelioration of soil degradation, reintroduction of micro-symbionts, control of annual weeds, direct seeding, and replanting of seedlings; (R5) Removal of grazing; (R6) See R4; (R7) Abandonment of cropping and livestock grazing: transition gradual, taking many years; (R8) Replanting with salt-tolerant species. These examples indicate the importance of recognizing where thresholds of irreversibility are likely to occur which prevent the system from recovering

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without management input, and of identifying clear goals as to what the desired outcome of restoration is.

SYSTEM RESILIENCE, AND FACTORS LIMITING SYSTEM RECOVERY

A key concept relating to disturbed ecosystems is that of resilience, which describes the ability of the system to return to a pre-disturbance state following the occurrence of a disturbance or the application of a stressor (Westman, 1986; Majer, 1989b, 1992). A resistant system will show relatively little change in response to a stressor, whereas a resilient system will show a potentially large initial change followed by a relatively rapid recovery. Five properties of resilience have been recognized, namely: (1) Elasticity, or the rate of recovery following disturbance; (2) Amplitude, or the threshold level of strain (the impact of the stressor) beyond which return to the original state no longer occurs; (3) Hysteresis, or the degree to which the path of change in the stressed system differs from the path of change following removal of the stress. Hysteresis and elasticity would seem to be inextricably linked; (4) Malleability, or the difference between the ﬁnal state of recovery of the system and the predisturbance state; (5) Damping, or the extent to which the system is liable to overshoot or oscillate about an equilibrium level. An underdamped system will approach equilibrium rapidly, overshoot, and oscillate for a period. An overdamped system will take a long time to reach equilibrium, but will not oscillate. A criticallydamped system will approach equilibrium slowly, overshoot slightly and then approach the equilibrium state. In this model, resilience is associated with the concept of equilibrium, in that it relates to the return to a given equilibrium following disturbance, and implicitly assumes a stable system (Pimm, 1984). This model is, however, too simplistic for many situations. Increasingly, it has become recognized that ecosystems may exhibit non-equilibrium dynamics (DeAngelis and Waterhouse, 1987; Pahl-Wostl, 1995), and that recovery following disturbance may not entail a simple return to the pre-existing state. Reasons for this

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may include different pathways of recovery resulting from the timing of disturbances relative to the life histories of component species (Noble and Slatyer, 1980), or transitions between alternative (meta-)stable states caused by different combinations and timings of disturbances (Westoby et al., 1989; George et al., 1992; Hobbs, 1994a). Despite this, the concept of resilience is useful in considering the potential for systems to recover following disturbance. Ecological restoration is necessary only where the system’s natural resilience has been compromised – that is, where the system’s amplitude has been exceeded (a threshold of irreversibility has been crossed), or where the malleability is such that recovery does not result in the desired system, or where hysteresis results in a recovery process which is too slow. These concepts are compatible with the ideas of multiple successional pathways and alternative stable states. The idea of system amplitude also corresponds with the idea that thresholds may exist beyond which the system is unable to recover (Hobbs and Norton, 1996). In such cases, the transition back from a degraded to a less degraded condition may require considerable restoration effort, as was discussed for the examples given in Figs. 29.2 and 29.3. In essence, therefore, ecological restoration is required only where the system’s resilience has been diminished in some way, or where the normal recovery processes are too slow to achieve management goals within a desirable time frame. Restoration requires that the stressors acting on the system are removed and may also involve replacing components that have been lost during the degradation of the system (Fig. 29.4; Brown and Lugo, 1994). Degrading processes can result in a variety of ecosystem responses, depending on the intensity, duration, and scale of the impact, and on which system components and processes are affected (Fig. 29.4a). Stressors which impact the processes of resource capture by plants (e.g., soil erosion, alteration of hydrology) are likely to have much greater impacts than stressors which remove or damage plants or consumers (e.g., pathogens or unsuitable ﬁre regimes). The potential for reversal of stressor effects is also dependent on which components and processes are being affected. It will be relatively easier to remove or control stressors that impact the biotic components of the system only, than to reverse the impacts of stressors which impact the resource base and the ability to capture resources (Fig. 29.4b). Cessation of the former type of stressor may result in ecosystem recovery

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Fig. 29.4. (a) Simpliﬁed model of an ecosystem with resource inputs, major components and ﬂows, indicating where stressors (circles numbered 1–5) impact on the system. Stressors 1 and 2 have the greatest impact since they act on resource supplies, while stressors 3– 5 act on biotic components; R, respiration losses; (b) the same system indicating the actions needed to restore the degraded ecosystem (circles above the bottom row of circles). Options 1 and 4 involve removing the stressor, while options 2, 3 and 5 involve replacing lost ecosystem components. The degree of difﬁculty and cost of restoration increase from 1 to 5. Redrawn from Brown and Lugo (1994).

without further management intervention (e.g., Allen et al., 1994), whereas cessation of the latter type is unlikely to be sufﬁcient on its own (e.g., Milchunas and Lauenroth, 1995). It is also possible that removing a stressor which affects only the biotic component will not result in ecosystem recovery if components of the original system have been lost. Similarly, if the effects on the biotic component have subsequent impacts on critical ecosystem processes, removal of the stressor will not be sufﬁcient. In these cases, restoration has to aim to replace the components that have been lost, or to provide alternatives which allow the rates of ecosystem processes to reach the values set as restoration goals. For instance, it has been suggested that the re-establishment of structural characteristics of

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tropical systems, even by species not originally found in the system, can reinstate adequate nutrient-retention processes (Ewel, 1986; Ewel et al., 1991). The level of resilience to particular disturbances is likely to vary among ecosystems and regions, depending on the natural disturbance regime experienced by the system. For instance, an ecosystem subject to recurrent soil disturbance by animals may be more resilient to human soil disturbance than one in which soil disturbance is not a feature. Similarly, the history of human disturbance may inﬂuence ecosystem resilience. Hobbs and Hopkins (1990) have suggested that the ecosystems of southwestern Australia are less resilient to human disturbance than ecosystems in similar climatic zones in the Mediterranean basin. They argue that a long association with technological human intervention has removed any non-resilient components from the systems of the Mediterranean, and resulted in ecosystems which are very resilient to human-related grazing and disturbance regimes. In contrast, the systems of southwestern Australia, although experiencing Aboriginal disturbance for long periods, have not been subjected to European-style cultivation and stock grazing until the last century. They thus contain many components that are not resilient to such disturbances. Halting or reversing the loss of these components presents a problem for ecological restoration quite different from that of the restoration of degraded systems in the Mediterranean area. Aronson and Le Floc’h (1996a) have explored a similar argument when comparing the ecosystems of Chile and the Mediterranean. Factors limiting recovery Given that ecological restoration is necessary only where natural ecosystem dynamics are disrupted, restoration can then be viewed as an attempt to reinstate or “kick start” these dynamics. Luken (1990) has provided a methodological framework for “directing ecological succession” in vegetation management, which is also relevant to ecological restoration. Pickett et al. (1987) observed that the three basic factors in succession are site availability, differential species availability, and differential species performance. From this, Luken proposed that these three factors could be regulated by using: (1) “designed disturbance” to manipulate site availability; (2) controlled colonization to increase or decrease the

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availability and establishment of particular species; and (3) controlled species performance to decrease or increase growth or reproduction of particular species. Although the speciﬁc requirements for each of these actions will vary from situation to situation, they provide a broad overview of what restoration aims to do. The basic premise is to modify the substrate to provide a suitable seedbed, and then to do whatever is necessary to allow the desired set of species to re-establish and persist. This approach recognizes that disturbances are often important components of natural systems that maintain species richness and provide opportunities for regeneration (Pickett and White, 1985; Petraitis et al., 1989). In other words, disturbance can be a positive inﬂuence and may be part of the natural dynamics of a system, and hence needs to be included in the management of that system. Such disturbances differ from the types of disturbance discussed previously, which are predominantly human-caused and generally have negative impacts on the system. The need to balance the system’s requirement for some types of disturbance against the need to exclude detrimental disturbances is one of the major conundrums facing land managers today, and is explored by Hobbs and Huenneke (1992). “Designed disturbance” aims to provide the positive disturbances required to prepare the site for establishment of the desired species. In some cases, negative disturbance will have already occurred (as in mine sites), and restorative action entails the amelioration of the physical and/or chemical properties of the substrate (Munshower, 1993). In others, some form of designed disturbance may be required to force the jump from one metastable state to another. For instance, where non-native plant species prevent the regeneration of natives, the removal of the non-native species by ﬁre, cultivation, or other means may be necessary. Similarly, where soil structural decline has occurred, physical disturbance of the soil, for instance by ploughing or scarifying, may be necessary to enhance seed germination and establishment. Reinstatement of surface heterogeneity in semiarid landscapes has been shown to be important in restoring nutrient and water relations necessary for germination and establishment (Tongway, 1990; Tongway and Ludwig, 1994; Ludwig and Tongway, 1995). Simple physical structures and the creation of small-scale “microcatchments” to reintroduce surface

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heterogeneity and impede surface water ﬂow may be relatively simple methods of restoring function in degraded areas to values which then allow the reestablishment of vegetation cover (Thurow and Jou, 1995). More speciﬁc types of designed disturbance may be necessary to encourage germination of desired species. Germination of some species is stimulated by ﬁre or heat treatment, while germination of other species requires processing by passage through the guts of mammals or birds. Recent work has shown that large suites of species in Western Australia and elsewhere require contact with smoke to germinate (Dixon et al., 1995). Pretreatment of seed to provide the appropriate cue by using the actual agent or something which mimics it will signiﬁcantly increase germination success. Colonization of a site by plants following disturbance may occur naturally if a seed bank persists in the soil (Putwain and Gillham, 1990; Bellairs and Bell, 1993); but the availability, germinability, and composition of the seed bank in relation to restoration goals needs to be assessed (Noble and Slatyer, 1980). If some of the species in the area have substantial stores of seed within the plant canopy, recently-harvested material can be used as a mulch or brush-covering to allow seeds to be released and dispersed. Further colonization may occur from surrounding areas, and this can be enhanced, especially if dispersal is by birds (McClanahan and Wolfe, 1993; Robinson and Handel, 1993). If neither soil seed stores nor dispersal into a site are effective in providing the required species mix, species may have to be introduced to the site artiﬁcially, either as seeds or as seedlings. A wide range of techniques for enhancing establishment success is available (Buchanan, 1989; Leopold and Wali, 1992; Harker et al., 1993). Most involve the provision of an adequate seed bed, ensuring the correct germination cues, and minimizing herbivory and interference from invasive species. These same techniques are important during the process of controlling species performances to achieve particular mixes of species abundances. In addition to plant colonization, ecosystem restoration requires the re-establishment of faunal components, although this aspect has traditionally received less attention (Majer, 1989a). While it is generally accepted that the re-establishment of the plant component of the system is necessary for faunal recovery, it is also increasingly recognized that faunal components may be essential drivers of the system (Jones et al.,

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1994). A striking example of this is the beaver (Castor canadensis), whose reintroduction into areas of central North America from which it had been virtually exterminated led to dramatic changes in the distribution and extent of wetlands (Johnston and Naiman, 1990; Johnston, 1994). Other less striking but important faunal impacts include herbivory, pollination, seed predation, and dispersal. The soil micro- and macrofauna also have essential roles in decomposition and nutrient cycling (Parker, 1989). Also of importance are other soil biota such as mycorrhizae (Allen, 1989; Allen et al., Chapter 22, this volume), without which plant growth on restoration sites may be severely restricted. The return to the ecosystem of these functional components is thus an essential element in restoration. While it is often the aim to restore a complete community, a whole branch of restoration ecology deals with the reintroduction of selected individual species to areas from which they have been eliminated previously (Maunder, 1992; Bowles and Whelan, 1994; Olney et al., 1994; Sarrazin and Barbault, 1996). The principles behind these activities are the same as for restoration in general – that is, the recognition and treatment of factors causing the original decline or extinction, and development of techniques for the establishment of viable populations of the target species. Frequent requirements for reintroductions are the provision of adequate food sources and habitat, and the removal of feral predators. Another manipulation often considered necessary in restoration is the removal of species preventing the achievement of restoration goals. Feral predators such as foxes and cats are thought to be the major factor preventing the re-establishment of native species in Australia and New Zealand (Friend, 1990; Towns et al., 1990; Armstrong and McLean, 1995). Nonnative herbivores are known to have major impacts which prevent ecosystem restoration, but which can be alleviated following their removal. For instance, in New Zealand and Australia, examples include rats (Rattus spp.) (Allen et al., 1994), rabbits (Oryctolagus cuniculus) (Chesterﬁeld and Parsons, 1985; Allen et al., 1995), and livestock (Profﬁtt et al., 1993; Greene et al., 1994; Petit et al., 1995; Rose et al., 1995). Other feral herbivores such as pigs (Sus scrofa), goats (Capra hircus), possums (Trichosurus vulpecula), water buffalo (Bubalus bubalis), horses (Equus caballus), and camels (Camelus dromedarius) are also problems in some areas. Considerable emphasis on methods

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of controlling non-native predators and herbivores is provided in restoration and management guides from these countries (e.g., Breckwoldt, 1983; Porteous, 1993). Also in the category of species which need to be removed or controlled to achieve restoration goals are weedy plant species (Berger, 1993; Hobbs and Mooney, 1993). In many instances, weedy species prevent the establishment of native species (Allen and Knight, 1984; Hobbs and Atkins, 1991; Menke, 1992), or completely alter the structure of the system (Humphries et al., 1991; Humphries, 1993; Cook and Setterﬁeld, 1996; D’Antonio et al., Chapter 17, this volume) or its dynamics, particularly in relation to the ﬁre regime (Whisenant, 1989; D’Antonio and Vitousek, 1992). Control of weeds can be achieved by mechanical, chemical or, in some cases, biological means, although it has recently been suggested that in many cases the weeds are symptoms of broader management problems, rather than problems in themselves (Hobbs and Humphries, 1995), and that careful management to prevent the establishment of weeds in the ﬁrst place is the most effective option to pursue. Non-native species have frequently been used in restoration to provide rapid recolonization, but in some cases this has proved detrimental to the establishment of native species (Wade, 1989; Wilson, 1989). The choice of whether or not to use non-native species depends on the goals of the restoration – in some cases, the system may have degraded to such an extent that the species originally present on the site can no longer establish. In other cases, non-native species may provide a more rapid improvement in the functional aspects being restored, such as water use or shelter, or may act as nurse species for slower-growing native species. The main concern in using non-native species for restoration activities is that their use may create a bigger problem later if the introduced species spread into other areas and become a degrading inﬂuence there.

LARGE SCALE RESTORATION

Most of the information and methodologies on ecological restoration centre on individual sites. However, site-based restoration has to be placed in a broader context, and is often insufﬁcient on its own to deal with large-scale restoration problems (Hobbs and Saunders, 1993; Saunders et al., 1993; Vos and Opdam,

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1993). Landscape-scale or regional-scale processes are often either responsible for ecosystem degradation at particular sites, or alternatively have to be restored in order to achieve restoration goals. I discuss this further in relation to fragmented ecosystems and large-scale aquatic systems. Fragmented ecosystems In fragmented systems the major impacts on vegetation fragments often arise from changes in the surrounding matrix (Hobbs, 1993b, 1994b). In particular, hydrological changes in the surrounding landscape (e.g., rising or falling water tables) can have dramatic impacts on remaining patches of native vegetation (Rowell, 1986; Barendregt et al., 1995; George et al., 1995). Here, the regulation of landscape-scale processes is an important goal for restoration. These goals include efforts to redress hydrological imbalances and to ameliorate ﬂuxes of materials and nutrients by reducing soil erosion (Ryszkowski and Kedziora, 1987; van Buuren, 1991; van Buuren and Kerkstra, 1993). For both of these goals, restoration requires the insertion of system components such as deep-rooted vegetation or windbreaks. In addition, buffer zones can be developed to alter ﬂuxes between landscape elements. For instance, restoration of riparian buffers can protect rivers and wetlands by acting as traps for sediment and nutrients (Muscutt et al., 1993; Burke and Gibbons, 1995; Vought et al., 1995). Also important are attempts to provide landscape features which allow fauna to persist and move across the landscape. Here, restoration may attempt to increase the amount of habitat available, provide buffers around important remnant areas or provide linkages or corridors between isolated remnants (Fig. 29.5; Hobbs, 1993a). While such activities can be guided by basic descriptive principles (e.g., bigger is better, connected is better), there are few quantitative guidelines to indicate when a landscape is in need of restoration in terms of habitat availability and connectivity, or exactly how much is required. It is likely that the requirements will be set by individual species, and that different answers will be required depending on which target species are selected. There is a need, however, to examine landscape geometry in a more general sense to ascertain whether landscape descriptors can be developed which are useful indicators of overall landscape condition. An initial attempt has been made by Aronson and Le Floc’h (1996b) to provide some

Fig. 29.5. Options available for landscape-scale restoration for retention or reintroduction of fauna in extensively fragmented systems (from Hobbs, 1993a).

general descriptors, or what they term “vital landscape attributes”. A further focus of revegetation is the aesthetic impact on the landscape, and principles of landscape design have been developed which centre around visual and aesthetic considerations (Bell, 1993, 1995). There has apparently been little investigation of how these principles based on visual criteria match up with the functional criteria discussed above. Clearly, this is an important area for further study, especially if mismatches are likely between aesthetic and functional prescriptions. Aquatic systems Restoration of many aquatic systems must tackle not only the degradation within the system itself, but the factors causing the degradation which enter from elsewhere. Examples include eutrophication and pollution of lake and wetland systems due to land-based inputs (Francis and Regier, 1995), and rises in saline water tables due to vegetation clearance (Froend et al., 1987; Froend and McComb, 1991). The need for restoration of large-scale processes is illustrated well by recent initiatives to restore the Kissimmee River in southern Florida; these aim to restore about 70 km of river channel and 11 000 ha of wetland which was signiﬁcantly modiﬁed when the river was channelized during the 1960s (Cummins and Dahm, 1995; Koebel, 1995). Central to the restoration project is the goal of dechannelizing the river, and thus re-establishing historical hydrologic conditions (Dahm et al., 1995; Toth et al., 1995). The modiﬁcation of the timing and duration of ﬂoodplain inundation should improve dissolved oxygen conditions, change substrate conditions, and allow germination and establishment of

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Fig. 29.6. Relative timing of the recovery of different ecosystem components projected for the restoration of the Kissimmee River system; (A) represents the time required for the return of ﬂoodplain, channel margin and in-channel vegetation, physical features, and sediments. The time between (A) and (B) represents the projected time for natural habitat development to allow invertebrates to return to the channel. The time between (B) and (C) is estimated from the generation time of the longest lived ﬁsh species in the river and the estimated effects of stochastic environmental variation on recruitment success (from Trexler, 1995).

wetland vegetation and the redevelopment of benthic invertebrate populations (Harris et al., 1995). This should in turn allow the redevelopment of ﬁsh and waterbird communities (Trexler, 1995; Weller, 1995). The ambitious Kissimmee River restoration project depends on the restoration of hydrological processes over a large area, and could not be contemplated on a smaller scale. The studies mentioned have attempted to use existing information to develop a clear conceptual framework within which the restoration project can be carried out and assessed. There is a recognition that restoration is a long-term process, and that different system components will take different lengths of time to recover (Fig. 29.6; Trexler, 1995). Clearly, restoration efforts often need to focus at a scale larger than an individual site. A larger-scale focus is also needed for connected systems such as waterways and wetlands, again because the degrading inﬂuence may originate in an adjacent system, or even one that lies some distance away. For instance, recent initiatives to restore the Everglades in southern Florida recognize that restorative measures need to incorporate changes in water and land use in adjacent agricultural and urban areas, in order to restore important aspects of the hydrological regime, and reduce nutrient inputs and other harmful impacts (Walters et al., 1992; Cohn, 1994; Davis and Ogden, 1994; Light et al., 1995).

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The Kissimmee River, discussed above, forms part of the larger hydrological unit which feeds into the Everglades. A similar picture emerges for other largescale systems such as Chesapeake Bay and the Great Lakes in North America (Costanza and Greer, 1995; Francis and Regier, 1995), and the Baltic Sea in Europe (Jansson and Velner, 1995). In addition to the requirement for a regional approach, other common features of these and other regional systems include the prevalence of numerous myths concerning management and restoration requirements, the occurrence of numerous management crises, and the importance of social and institutional factors which act as strong barriers to effective restorative actions (Gunderson et al., 1995). Large-scale ecological restoration thus involves not only ecological but social, political and institutional changes. Ecological restoration needs to be part of the overall strategy of ecosystem management – it can rarely take place as an independent activity, especially at the larger scales discussed in this section.

CONCLUSION

As ecosystems become more subject to human-induced disturbance, ecological restoration will become an increasingly important component of ecosystem management. Possibilities of restoring degraded systems will only be as good as current understanding of the systems and their functioning. Restoration is generally a costly process, and the complexity of most ecosystems and their responses to multiple environmental factors means that the outcomes of particular restoration measures cannot be predicted accurately. It therefore makes sense to try to limit the extent to which restoration is necessary by ensuring that the management of ecosystems for conservation and production aims to maintain or enhance their present status. Prophylaxis is always preferable to and less costly than therapy, and a strategy of careful use of the world’s ecosystems would be in the long-term interests of humanity.

ACKNOWLEDGMENTS

I thank James Aronson, Lawrence Walker, and an anonymous reviewer for constructive comments on the draft chapter, Tien McDonald for discussions on resilience, James Aronson, Sandra Brown and Colin

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Yates for permission to use previously published material, and ﬁnally, Lawrence Walker for his patience and enthusiasm.

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Chapter 30

ENVIRONMENTAL POLICIES AS INCENTIVES AND DISINCENTIVES TO LAND DISTURBANCE Gregory E. ECKERT and C. Ronald CARROLL

INTRODUCTION

In this chapter we evaluate human inﬂuences on ground disturbance through the implementation of policies, public and private. We consider policy to be any operating guidance of government agencies, individuals, industries, environmental and developmental organizations. Policy includes legislation, regulations, and technical guidance which lead, directly or indirectly to land degradation, conservation, or restoration. While natural disturbances such as ﬁre or ﬂoods play an important role in ecosystem dynamics, and hence make up a substantial part of this volume, because of the obvious connection to policy, we will focus on human disturbances on the land. We consider “ground disturbance” to be physical or chemical disruption of soil, and alteration of structure and diversity of terrestrial communities and ecosystems. Although we emphasize forestry and the management of public lands, other activities, including agriculture, invasions by exotic species, mining, military activities, urban and suburban development, and transportation corridors all have large impacts on the land (Barrow, 1991). Policies often have far-reaching implications beyond their intended target. Policies directed at water quality have land implications and vice versa, and general economic and ﬁscal policies may have a greater inﬂuence on land disturbance than any land conservation programs. Beyond short-term disturbances, we also evaluate how policies that permit resource degradation to develop slowly can, through cumulative effects, make ecosystems more susceptible to acute disturbances. Additionally, chronic degradation processes can make ecosystems more difﬁcult to restore following major disturbances. Many organizations inﬂuence land policy, and consequent land management activities, at large and small

scales, and from national to individual actions. The principal players in promulgating environmental policy are the state and national public agencies that implement public law. Public and private credit-lenders informally shape environmental policy through their lending requirements. Multi-national organizations, such as the United Nations, the European Community, the Andean Pact Countries, etc., also inﬂuence policy through regional agreements and treaties. A notable example of multi-national policy development is Agenda 21, a series of global environmental goals that arose out of the 1992 Rio Conference (United Nations Conference on Environment and Development held in Rio de Janeiro, Brazil, 1992). The European Community is currently in the process of adapting the goals of Agenda 21 to meet trans-national environmental concerns of European countries. Additional players in policy development are larger international nongovernmental organizations (NGOs), such as CARE International, Catholic Relief Services, and Plan International, which indirectly shape policy by encouraging sustainable land-use practices. The World Bank, the major multi-national credit lender, has historically encouraged large projects, many of which have had negative environmental impacts. In response to external criticism and its own internal assessments, new Bank policies incorporate environmental impact assessments and funding for environmentally focused programs through its Global Environmental Facility. The evaluation of all policies, and their interactions, are beyond the scope of this chapter. As appropriate to this volume and all land-management policies, we present the utility of analyzing land policies in the context of ecological concepts. The role of ecosystem characteristics and dynamics, including interactions,

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Table 30.1 Deﬁnitions of terms as they are used in the text Ecosystem

the inclusive set of interactions among populations and between those populations and their physical environment; the pattern of an ecosystem is described by its composition (e.g., number and kinds of species) and structure (the relative proportions of the species); the dynamics of an ecosystem are typically described by process rates (e.g., nutrient cycling) and trophic linkages; the spatial and temporal boundaries of ecosystems are determined by the nature of the investigation – that is, the essential attribute is the “system” that characterizes patterns and dynamics, not the locality; thus, a comprehensive lake-ecosystem study might include the lake itself along with its catchment basin, nutrient-loading from development, and distant transport of pollutants into the lake

Ecosystem management

an approach to maintaining or restoring the composition, structure, and function of natural and modiﬁed ecosystems for the goal of long-term sustainability; it is based on a collaboratively developed vision of desired future conditions that integrates ecological, socioeconomic, and institutional perspectives, applied within a geographic framework deﬁned primarily by natural ecological boundaries (from Meffe and Carroll, 1997)

Adaptive management

an interactive approach for managing complex systems that are inherently indeterminate; because a simple management decision may produce an array of possible responses, it is necessary to monitor decision processes continuously and correct those that led to undesired responses; in the context of managing renewable resources, the response of the resources to management decisions must be monitored and the management modiﬁed if objectives are not met

Stakeholders

in the most general sense, stakeholders are people whose lives may be affected in any way by decisions made in the management area; in practice, stakeholders are usually people who live in or near the management area, have economic activities associated with the area, or who have management or research responsibilities in the area; stakeholders may include public and private organizations as well as individuals

linkages, and scales as affected by policy-driven actions, will be analyzed using case studies of the United States National Forest Service and National Environmental Policy Act. Finally, we will examine how understanding of these concepts (through research and practice) can be better used to develop policies that are more comprehensive and anticipatory and less exercises in crisis management. Throughout this chapter we use terms, such as “ecosystem management”, that have been burdened with many deﬁnitional nuances. In order to provide consistency we provide a glossary of key terms in Table 30.1. UNINTENDED CONSEQUENCES AND INAPPROPRIATE SCALES

Policy is seldom a simple link between social goals and reality. Policy can have unintended consequences, as when price controls on food lead to undervaluing farm products and thereby act as a disincentive for farmers to invest in long-term erosion control measures that have low returns in the short term. Policy can also be applied at the wrong temporal or spatial scale. Policy applied too broadly risks being maladaptive to local conditions. At the other extreme, policy that is written solely for

one component of a landscape, such as for management of an area that is protected for its biodiversity, can have a negative impact on other parts of the landscape, as for instance by increasing rates of deforestation on adjacent, unprotected lands and thereby further isolating the protected forest. Temporally, policy that is based on limited historical information may miss the mark over long time horizons. In this section we expand on the following dictum: simple and rigid policies seldom work as intended in complex, changing environments. Examples of such policy failings are found worldwide. In developing countries, priorities such as establishing territorial claims, servicing foreign debt, and maximizing foreign exchange have contributed to land degradation. Policy trends in Latin America related to these broader national goals and which have caused land disturbance include the following: • designation of deforestation as a “land improvement” to establish land tenure • promotion of large-scale export-oriented projects in livestock, forestry, and mining • colonization of frontier areas by settlers, who use inappropriate methods for managing their land. Using Brazil as an example, once government

ENVIRONMENTAL POLICIES AS INCENTIVES AND DISINCENTIVES TO LAND DISTURBANCE

road construction projects opened dense forest areas, government policies encouraged settlement of the Amazon frontier through land speculation, in which land clearing gained a colonizer title to that piece of land (this title was then used to obtain bank loans to develop the land further). Land tenure was ensured through pasture establishment, ofﬁcially classiﬁed as a land improvement. Permanent occupation of frontier regions was encouraged through subsidies and tax relief which hid the full cost of production (Fearnside, 1989). For example, until recently, colonizers could receive investment tax credits – up to 50% credit against federal income tax (Mahar, 1989). Such ﬁscal incentives and subsidized credit lines were intended to encourage land uses such as cattle-raising which has allowed a relatively small population to have a large impact on the rainforest. In Brazil (Reid, 1989), these generous tax and credit incentives led to the conversion of nearly 9×106 ha of forest into cattle ranches; many have since been abandoned and the rest are only marginally proﬁtable. Reid noted that, in the long term, such policies lead to “ . . . commodity prices and trade policies that depress farmers’ proﬁts, reduce the demand for labor and lower agricultural land prices. Consequently, the return on investments for land improvement and soil conservation is reduced, and erosion and lower soil fertility are the outcome.” Generally, these problems have arisen on soils inappropriate for this type of management (Hecht, 1993). In Ecuador, as elsewhere in Latin America, organizations support colonizers who occupy and clear forested areas, even though land titles may be already held by others. After the land has been cleared, colonizers sell their parcels to the supporting organization which realizes substantial land acquisitions at a meager cost, and the original landowners are left with no recourse (Southgate and Whitaker, 1994). Resource problems in Ecuador are common to many other Latin American countries, yet farmers use few soil conservation measures. One problem is that, to maintain urban population support, governments have subsidized food prices by both their control of domestic grain prices and the absence of barriers to the importation of inexpensive foreign foodstuffs. Farmers react to this competition by either planting inappropriate crops or by failing to make long-term investment in their land. Thus, social policy to satisfy urban populations has contributed to land degradation in frontier regions, and the choice to secure food in the short term undercuts a country’s ability to achieve sustainable land use.

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Chronically high inﬂation devalues long-term credit for conservation practices. Furthermore, farmers are less likely to make long-term investments in their land because land tenure is insecure due to inadequate and poorly managed title programs. With such disincentives acting against the need to manage for sustainable returns, it is not surprising that many farmers opt to “mine” whatever beneﬁts they can from their land before ﬁnal abandonment. If not planned in a large landscape-policy context, even the establishment of protected areas can lead to rapid land degradation in developing countries. The establishment of Burkit National Park in Sumatra provides a particularly well-documented case of this counter-intuitive effect of policies that were designed to protect biodiversity, leading instead to more deforestation. Before the establishment of the park, local people practiced a unique form of agroforestry known as dammar, which resembled natural forest cycles of gap formation and succession. Initially, a small forest plot would be cleared and planted mainly to dryland rice (Oryza sativa). After a year or two the rice ﬁelds would be interplanted with shrubs such as coffee (Coffea arabica) and cloves (Syzygium aromaticum). Later, fruit trees and the native Shorea japonica (a source of commercial resin) would be added. Over a period of several decades the plot would come to resemble an old second-growth forest, but consisting of species that had renewable economic returns. Throughout the management area, the forest was a mosaic of a few rice parcels and economic forests in various stages of development. The establishment of Burkit National Park removed a critical percentage of the forest that had been traditionally used for shifting dammar agroforestry. As a result, insufﬁcient land remained to permit agroforestry, and more and more of the remaining forest was converted to upland rice and, to a lesser extent, coffee and cloves. Thus, the formation of Burkit National Park had the unfortunate consequence of hastening forest destruction outside the park’s boundaries (Mary and Michon, 1987). The United States has experienced some of the same land degradation problems in its own effort to maximize the extraction of food, mineral, and timber resources from the land. The infamous “Dust Bowl” disaster in the 1930s provides a striking example of how poor land-use policy led to chronic land abuse, which in turn set up conditions of extreme environmental vulnerability. Decades of above-average precipitation gave the normally much drier Plains states

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a misleadingly lush appearance. During this period, row crops were encouraged through federal farm-assistance programs largely to satisfy the interests of railroads and the grain market. Row-crop practices, suitable for the moist eastern climate, led rapidly to severe wind erosion and the “Dust Bowl” when the Plains climate returned to much drier conditions (Eckholm, 1976). More recently, long-standing policies to control river ﬂow in the United States are being questioned in the light of recent large-scale ﬂooding. Here, a previous lack of understanding of river and wetland dynamics, and the lack of historical information on the timing of such massive ﬂoods, led to disasters causing damage to homes, farms, and industry costing millions of dollars. In the more arid western states, United States water law follows the principle of “prior appropriation” for riparian water within state boundaries. Basically, the earliest users of the water were given priority rights over later users. In order to maintain water rights, users had to withdraw their water allocations, even if the water allocated exceeded their needs. Ground water has generally been treated as a “commons”, a resource free for the taking. For both riparian and ground water, public policy acts as a disincentive to the conservation of water resources (National Academy of Sciences, 1992). Furthermore, continuous application of riparian water to ﬁelds in environments with high evaporation leads to increasing salt concentrations downstream and to salinization of downstream ﬂoodplains. Some advocates for changing western water-rights policy have argued that farmers who use water-conservation practices should be able to sell unused portions of their water allocation to other users, especially urban users. Although this approach would encourage more efﬁcient agricultural use of water, by simply changing the composition of the end users, it would do little to reduce total water use. More effective, though politically unpalatable approaches would include the removal of public subsidies of large irrigation water projects so that water pricing would reﬂect the real costs of extraction, storage, and distribution of irrigation water; possible reverse pricing of water so that annual decreases in water use, by any end user, would result in lower water charges per unit; and reduction in annual and estate tax rates on homes and businesses that practice water conservation. Governments play a large role in land use through revenue programs. Not only are activities destructive to the land undertaxed, these activities often are rewarded through subsidies, as noted with the case

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of large water projects in the western United States. Revenues that can inﬂuence land-use decisions include sale taxes, payroll taxes, and income taxes; and taxes and subsidies for natural-resource use, such as logging and ﬁshing. Resource “mining”, and other harmful land-use activities, have been encouraged by governments either through direct subsidies or lack of requirements that operators should internalize the costs of environmentally harmful practices, and that governments and operators should recognize the future value of resources when establishing discount rates in net-present-value assessments. This lack of foresight is evident in the number of sites for hazardous wastes, the loss of wetlands, and depletion of ﬁsheries. Public and private policies related to human population have been insufﬁcient to couple human population policies with land-use policies. Giampietro (Chapter 32, this volume) provides historical perspectives and a model of population–resource interaction and urges greater action in developing policies on human population management. Pulliam and Haddad (1994), in calling for greater involvement of ecologists in studying human population and the earth’s “carrying capacity”, have warned that increases in human population would lead to continued environmental degradation. Because this issue is addressed in this volume, we limit our recognition of this critical policy concern to the past failure of policies not linking poverty to greater dependence on short-term resource use, or linking afﬂuence to greater overall use of resources. Clearly, the many examples of social, economic, and landuse policies inﬂuencing land degradation have given rise to new approaches to bring ecological processes into policy and management. We explore one of these options, ecosystem management, below.

ECOSYSTEM MANAGEMENT: INCORPORATING ECOLOGICAL PRINCIPLES INTO LAND USE

Policies often fail to recognize the spatial and temporal patterns of natural processes that affect land. Furthermore, environmental policy seldom recognizes the natural variability of ecosystem patterns and processes or that precautionary principles need to be linked to an understanding of the limits to resiliency that are speciﬁc to the particular ecosystem and stresses. Natural disturbances are an integral part of these patterns and processes, but impacts are scale-related. While these concepts excite the research ecologist, dealing

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with this complexity can be intimidating to the policy writer, and a nightmare to the regulated community. In order to bring these three groups together to develop practicable land-management policies, it is useful to begin with a simple model (Fig. 30.1). Loosely adapted from Holling (1995), Fig. 30.1 depicts time-lines to emphasize the dynamic nature of ecosystems and the importance of the frequency of signiﬁcant disturbances. Following Holling’s terminology, the time-line includes a “Conservation” stage, represented by the public’s normal perception of an intact ecosystem (e.g., an old-growth hardwood forest). Holling described this as a stage of energy storage and increasing biotic linkages. This energy as well as the suite of inorganic and organic molecules cycling through the system are “Released” following signiﬁcant disturbances such as hurricanes or ﬁres. In the dynamics of natural ecosystems, water, energy (in the form of carbon compounds), and mineral elements such as nitrogen, phosphorus, potassium, and sulfur are “Reorganized” largely by microbial processes. “Exploitation” by organisms that colonize disturbed areas begins during the reorganization phase. As these organisms interact with the environmental conditions, they are replaced by other groups of organisms, leading again to the regeneration of the organisms and associations of the “Conservation” stage. The temporal and spatial aspects of these dynamics vary by the suite of organisms that are involved. For example, one may consider the following extremes: mite populations are inﬂuenced by moisture ﬂuctuations in a single cubic centimeter of soil within days, whereas the structure of many square kilometers of Chilean forests are remnants of disturbances that occurred 400–500 years ago (Veblen, 1985). In natural conditions, with disturbances such as ﬂoods or ﬁres, the disturbance can add nutrients, and provide conditions for seedling germination and protection from herbivores. In this case, recruitment of species, nutrient changes, and the disturbance itself would be part of the natural suite of characters for that particular system. The reorganization phase occurs at a time period after disturbance sufﬁcient for plant colonization to re-establish appropriate energy sources, moisture, and temperature conditions for soil biota, which in turn recapture and recycle nutrients released as a result of the disturbance. The time necessary for this reorganization is related to the degree, extent, timing, and type of the disturbance. An important point is that the suite of species involved in each stage and the duration of each stage will vary, because succession is

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Fig. 30.1. Ecosystem dynamics and temporal disturbances. Many ecosystems undergo repeated cycles of nutrient and energy release following a signiﬁcance disturbance, followed by reorganization, exploitation, and conservation (see text for explanation of these phases). Under normal disturbance regimes, the result is a landscape that contains ecosystems in various phases of the cycle, as shown for the example of a forest on the middle time-line. When disturbances become too frequent (as shown in the bottom time-line) the ecosystem is unable to regenerate and begins to degrade (indicated by the darkgrey bar). Adapted from Holling (1995).

not a deterministic process. However, biologists would clearly recognize these differences as normal variations on a theme and, in the case, for instance, of a deciduous forest of eastern North America, would view the mosaic nature of these stages as a normal part of forest pattern and dynamics. However, as shown in the bottom time-line on Fig. 30.1, if the frequency, intensity, or extent of the disturbances become too great the forest ecosystem will become degraded and may even change into something else (e.g., a shrubby grassland, in the case of forests near the prairie ecotone). Managed ecosystems respond similarly to man-made disturbances. The “Release” in Holling’s model occurs when soil is plowed, timber is harvested, or crops are subjected to pathogens or large-scale herbivory. If disturbance occurs during the “Conservation” stage, the system will respond by secondary succession through organismal stages that have had stored propagules of “desired” species, or those species found within the suite of characters for that system (what would

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be viewed as the normal successional trajectory). In managed ecosystems, the “Reorganization” phase is critical to continued use of the land. The success of this phase is related to another ecosystem property, resiliency, meaning the range of variation that ecosystem processes and patterns can undergo before the system changes (i.e., “degrades” in anthropocentric terms) into a new system with its own characters. An example would be over-harvesting of timber (i.e., a high degree of “Release”) on semi-arid lands, which deﬂects the system to a grass–scrub ecosystem. If a major disturbance occurs during the “Reorganization” stage, or is repeated at a greater frequency than that to which the system has adapted, the propagule base will be decreased, leading to recruitment of exotic species, or even an association of species that may slow, or even stop successional dynamics through either physical barriers or regulation of nutrient ﬂow (Connell and Lowman, 1989). Often, disturbances result from policies that focus on one attribute of land management. For example, in the Paciﬁc Northwest, ﬁre has been used after tree harvesting to clear land of debris. While this may have facilitated replanting by foresters, in some cases the burning reduced the amount of mycorrhizal fungal inoculum in the soils, and resulted in decreased seedling establishment. Similarly, removal of all hardwood has led to higher levels of the soil actinomycete Streptomyces sp., which is known to inhibit mycorrhizal formation (Amaranthus and Perry, 1994). There are also examples of single species inﬂuencing ecosystem properties (Vitousek and Hooper, 1993). For instance, as observed by Lesica and DeLuca (1996), the widespread use of the introduced wheatgrass (Agropyron spp.) as a soil conservation measure results in greater areas of exposed land surface; moreover, it does not contribute as much root biomass, detritus, or exudates to the belowground system as native vegetation. Also, having a greater carbon to nitrogen ratio, wheatgrass inﬂuences soil microbial dynamics and promotes lower aggregate stability and thereby higher potential for erosion. Land-use policies have not recognized, or at least not incorporated, the need to prevent “release” from exceeding the resiliency of the ecosystem. Environmental policy should prevent levels of human exploitation that exceed resiliency, but when it is exceeded, there should be a policy triggering an appropriate restoration effort (see Hobbs, Chapter 29, this volume). An example is seen in the United States in policy mandated by the National Forest Management Act, discussed

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in detail below. In this Act, timber harvesting is prohibited on certain kinds of land, usually steep slopes, that are vulnerable to degradation from tree harvesting. In order to harvest vulnerable areas, the timber plan must include a workable restoration plan and allotted funds for restoration. However, because ecosystems are dynamic, restoration must be viewed as a continuing experiment with long-term monitoring and accountability. Environmental policy seldom deals well with the long term, and may limit accountability for the operation and maintenance of restoration projects to a few years. While protection of single species or speciﬁc site conditions has driven policies in the past, broader ecological attributes are now being incorporated into management. Along with the suite of ecosystem characters described by Hobbs (Chapter 29, this volume), Orians (1990) has advocated the identiﬁcation of “valued ecosystem components (VEC’s)” as the basis of assessment and management decisions. These “valued ecosystem components” include single species, interspeciﬁc interactions, species richness, aesthetic limits, and ecosystem services, such as soil fertility, prevention of soil erosion, detoxiﬁcation and recycling of waste products, regulation of hydrologic cycle and gaseous components of atmosphere, control of pests, pollination, and preservation of the genetic “library.” One way in which such ecosystem components can be incorporated into policy would be to consider (van Wilgen et al., 1996): (1) For a particular ecosystem, how should components be prioritized for preservation? (2) Which components are negatively affected by a particular activity on the land? (3) What fraction of designated areas should be devoted to each component so as to maintain all components in acceptable proportions? The development of these and similar concepts has resulted in the current effort to promote ecosystem management as deﬁned in Table 30.1. Ecosystem management (albeit sometimes under politically more acceptable labels) as a new and more holistic approach to conservation is being embraced by natural heritage organizations, including public agencies in the United States such as the National Forest Service, Bureau of Land Management, Environmental Protection Agency, and the Biological Resources Division of the US Geologic Survey, as well as by private organizations such as The Nature Conservancy. Ecosystem management is well established in many countries. One

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example is the International Model Forests network, with programs in Canada, Mexico, the United States, Southeast Asia, Europe, and Russia. If the major public and private land-use agencies are embracing ecosystem management as a policy mandate, one needs to be clear about the meaning of ecosystem management. We view ecosystem management as a process that consists of three essential components. First, it takes the ecosystem as the appropriate scale for the management unit and for science-based decision-making. Second, a ﬂexible and on-going model of adaptive management is used throughout the life of the project. Third, stakeholders in the management area are brought into the process of decision-making, from establishing research priorities and problem identiﬁcation through implementation of management decisions. In the following paragraphs, we discuss the meaning and signiﬁcance of each of these components. The ecosystem as a management unit In biology textbooks, ecosystems are conventionally deﬁned as a community of organisms interacting among themselves and with their physical environment. The meaning of ecosystem in a management context is considerably different (see deﬁnition of Ecosystem Management in Table 30.1), and there are attributes of ecosystems that will stretch the ingenuity of policy makers if they are to be incorporated in environmental policy: (1) Ecosystems have structure (a range of characteristic species and abundances) and processes (generally biologically regulated ﬂuxes of energy and cycling, and transformations of nutrients, each having its own range of variability). (2) Ecosystems are complex and non-linear, hence it is difﬁcult to predict their response to perturbations. (3) Ecosystems are dynamic, not static entities – that is, they are constantly changing within boundaries of resiliency. (4) Ecosystems have emergent properties; they are not simply the sum of their parts, but exhibit novel properties through indirect effects. (5) Ecosystems are open; they receive and export energy, nutrients, and waste products (including heat as degraded energy). In ecosystem management, the ecosystem concept takes on a spatial meaning. All of the attributes of ecosystems, deﬁned above, still apply but, additionally, one needs to add a spatial component. The spatial scale

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should encompass area and landscape heterogeneity sufﬁcient for populations to maintain themselves in spite of temporal variation in resource abundance. As a concrete example, grizzly bears (Ursus horribilis) depend on lipid-rich nuts, among other foods, from high-altitude pines, and on ﬁsh spawning in creeks. An ecosystem that would contain grizzly bears would need to be large enough to encompass these resources. Therefore, an ecosystem that is the focus of research may be of any size; it will depend on the nature of the research questions. An ecosystem, when it is a holistic management unit, ideally is large enough to contain the critical natural resources that are needed to maintain biodiversity (food, habitat, etc.) as well as non-degradative activities of stakeholders; and should be enclosed by natural boundaries (as in the case of a large catchment) that provide some control over external threats (such as pollution of streams). In this sense, “ecosystem” retains its biological deﬁnition, but is placed within the context of a landscape in order to meet the goals of management. By choosing the appropriate conﬁguration of critical resources and ecological boundaries, it is possible to sustain the biodiversity, the ecological processes and patterns, and the activities of stakeholders for the long term with minimal intrusive management. Examples of ecosystem management There are many examples of at least the partial application of ecosystem management. In South Africa, efforts to fund restoration are linked to water economics. The costs of keeping shrublands free of exotic tree species can lead to long-term cost savings in water resources for dry areas of South Africa. The native vegetation is adapted to drought, binds soils, conserves water, and promotes low-intensity ﬁres, and management focus is on the whole shrubland ecosystem to ensure high water yields and low impacts on soil from periodic ﬁres (van Wilgen et al., 1996). Coxhead and Jayasuriya (1994) have also followed the course of integrating quantiﬁable measures of ecosystems and economics with social networks. These authors modeled prices, production, and inﬂuence on degradation in both lowland and upland systems to show, in a broader analysis of the economic landscape, that policies promoting lowerosive tree crops have a greater social beneﬁt (i.e., overall agro-region production improves). In a more complete example of ecosystem management, the protection and restoration of the Chesapeake Bay in the

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mid-Atlantic region of the United States has involved participation by public and private organizations from three states and the federal government. Land-use practices (forestry, agriculture, industry, residential) have been evaluated with respect to their impact on the Chesapeake Bay ecosystem and improvements have been both suggested and locally regulated (Meffe and Carroll, 1997). Adaptive management as a ﬂexible model A major consequence of the dynamic nature of ecosystems is that their management must also be dynamic. Management that responds in creative and innovative ways to changes in the system is known as adaptive management (Table 30.1). The inherent dynamics of change are a salient feature of ecosystems, and make management in a prescriptive fashion difﬁcult and ineffective. Consequently, management must take a different tack: to be ﬂexible, adaptive, and predictive. This means accepting a range of possible outcomes of a management action, with some probability of each outcome occurring; approaching management as an ongoing experiment, learning from results, and modifying future actions; and personal and institutional willingness to admit to mistakes and learn from them. Historically, public agencies have followed a much more rigid “top-down” approach to management, what Holling and Meffe (1996) refer to as a pathology of natural-resource management. Adaptive management places more responsibility on local managers to make good science-based decisions, and to accept accountability for their decisions. This, in turn, implies that new training programs will be needed and that adaptive managers will be supported. Stakeholder participation If policy is to be locally adaptive (to site-speciﬁc conditions), then decision-making in ecosystem management should be highly participatory. In the working deﬁnition of ecosystem management, people are explicitly included as integral parts of the ecosystem. People are included for a suite of reasons including: their large and pervasive populations, ethically or religiously based call to be environmental stewards, and the role of people as “keystone species” in the sense that they modify most ecosystem processes and patterns. It should be clear that long-term successful management of ecosystems requires the cooperation of

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stakeholders (Eckert et al., 1995). Here, we use the term “stakeholder” in a broad and inclusive sense to include private citizens and public and private institutions with an interest in the management area. Principal stakeholders are those people whose livelihoods or residences are connected to the management area, and those institutions, public and private, with activities in the management area. Principal stakeholders should play major roles in the development, implementation, monitoring, and evaluation of the ecosystem management plan. Participation by minor stakeholders is best served through forums for public commentary. Even in situations that call for ecosystem management, operating policies are usually interpretations of broader mandates of legislation or general goals of an organization. Lack of integration with other goals and policies can lead to conﬂicts in management decisions. We consider some of the difﬁculties in implementing ecosystem management within current policy frameworks, using the National Environmental Policy Act in the United States as an example of legislation that has lost much of its substantive powers, and the National Forest Management Act as legislation that is strongly inﬂuencing resource management in the United States National Forests. The National Environmental Policy Act: the power of disclosure Previous environmental policies have played an important role in the development of ecosystem management. Critical to the success of adaptive management and informed decision-making by stakeholders are the analysis of potential impacts of an action and the development of action alternatives. These requirements were in part derived from their use in the National Environmental Policy Act of 1970 (NEPA) in the United States. This Act requires that an environmental impact statement (EIS) be prepared for federal actions that may have negative environmental consequences. A plan must be prepared showing how adverse environmental consequences will be mitigated. In ecosystem management, this approach can be used to analyse which ecosystem characters would be stressed and to what degree, and how the proposed action would match the stability and resiliency of the land affected. Alternative actions to the one proposed must also be presented and their relative environmental costs and beneﬁts presented. Except for cases of national security and emergencies, no federal agency is exempt

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from the requirements of the National Environmental Policy Act. During the 1970s, the Act was interpreted as a substantive statute, but from the 1980s onward courts have interpreted it as merely a procedural statute (Weinberg, 1994). To explain this, in a 1989 Supreme Court case (Robertson and U.S. Forest Service v. Methow Valley Citizens Council) the environmental impact statement identiﬁed possibly serious impacts on a population of mule deer (Odocoileus hemionus) from a proposed ski resort. But, in the opinion of the court, if the National Forest Service determined that the “beneﬁts” of the ski resort outweighed the “costs” of damaging the deer herd, or even completely destroying it, the Service could proceed with the ski resort project. That is, “ . . . NEPA merely prohibits uninformed – rather than unwise – agency action (Platter et al., 1994).” In other words, a federal agency has to consider the environmental consequences of its actions and analyse alternative actions through a formal process, but the National Environmental Policy Act does not require the federal agency to change its original proposed actions. If a federal agency fails to comply with the Act’s requirements by failing to prepare an environmental impact statement for a major project, the courts may issue an injunction that delays the project until an adequate environmental impact statement is prepared. However, in recent years, courts have issued far fewer injunctions than were issued in the early years of the Act, thereby further reducing its substantive nature. If the courts have weakened the National Environmental Policy Act, does this mean that it is a failed environmental public law? We propose that the Act as a substantive policy has remained effective for two reasons. First, the public disclosure requirement of the Act has informed the public better, enabling them to exert pressure on agencies to select environmentally sound alternatives. This has led to strong “citizen suit” provisions in most Federal laws in the United States. These provisions state that private citizens may sue another party which is in violation of a particular environmental law if the government has failed to do so. The provisions also allow citizens to sue the United States Government in limited circumstances. Second, the National Environmental Protection Act has had a substantive effect, notwithstanding the courts, beyond the federal sphere. At least 24 states have requirements for environmental impact assessment (Platter et al., 1994), and several states, notably California, New

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York, and Washington, have modeled environmental legislation after the Act, but with much more substantive enforcement powers (Weinberg, 1994). Furthermore, the National Environmental Protection Act has affected international environmental policy. As of 1992, 20 countries, mainly in Europe, had adopted legislation requiring environmental assessments similar to those required by the National Environmental Protection Act, especially where transboundary issues were involved (Cooper, 1993). Because the major multi-national development banks (the World Bank, the African Bank, the Asian Bank, and the Inter-American Development Bank) require procedures similar to environmental impact assessments before initiating projects, the international effect of the Act in developing countries has been substantial. Similarly, the United Nations Development Program frequently adopts World Bank environmental assessment guidelines. Because the United States Agency for International Development (USAID) falls under the federal mandate of the National Environmental Protection Act to prepare environmental impact assessments, the Act has also inﬂuenced individual USAID projects and has led to changes in the Agency’s programmatic priorities. For example, environmental assessments by the Agency of the use of pesticides to control locusts in Africa led to changes in pesticide use by other donor countries involved in locust control in Africa (Cooper, 1993). The broader implications of this environmental assessment, which was critical of undue reliance on chemical control of pests, contributed to USAID’s decision to create a new program, the Integrated Pest Management Cooperative Research Support Program.

NATIONAL FORESTS IN THE UNITED STATES AS A CASE STUDY IN ECOSYSTEM MANAGEMENT

The Multiple-Use Sustained-Yield Act of 1960 represents the ﬁrst legislative modiﬁcation of the Organic Act of 1897, which established the national forest reserves in the United States to provide a continuous supply of timber. The 1960 Act mandates the protection and management of recreation, wildlife, ﬁsh, and rangeland resources. Timber was to be managed on a “sustained yield” basis. The Act is now regarded more as a statement of principles and required disclosure than as a substantive policy directive (Tuholske and Brennan, 1994). Many aspects of management on National Forest lands in the United States derive

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from the Forest Rangeland and Renewable Resources Planning Act (1974; amended in 1988 as the National Forest Management Act). The Congressional ﬁndings state that this act should “ . . . promote a sound technical and ecological base for effective management, use, and protection of the Nation’s renewable resources (§1600).” The National Forest Management Act (NFMA) was passed largely as a result of public concern over clearcutting which was perceived to be excessive, and damage to streams and rivers. This Act requires the development of management plans (“Forest Plans”) for each of the nation’s National Forests. The plans are developed by interdisciplinary teams, and require public participation and commentary. The Act authorizes budget requests to “replant and otherwise treat” an area equal to the amount of the annual harvest, plus any backlog of national forest land needing treatment. A technical advisory committee of specialists (the Committee of Scientists) not members of the Forest Service provides oversight review on the Forest Plans. Each forest is divided into different management areas in which land-use priorities and restrictions are deﬁned. For example, an area with good timber-site characteristics might be primarily designated for timber harvest, while another area with steep slopes and vulnerable wetlands might be designated primarily for wildlife management and timber removal might be limited or even entirely prohibited. Thus, through the forest planning process, management areas can protect the more vulnerable forest lands from disturbance. Furthermore, the process leading to restoration of the original plant cover must be established within ﬁve years of tree harvesting. The Act also prohibits logging on areas that are unsuitable for sustained-yield practices, and requires the protection of biological diversity through habitat management. A critical feature of the Act is that courts have interpreted the statute to say that wildlife is given co-equal standing with timber in forest management (Tuholske and Brennan, 1994), although there remains some ambiguity about how diversity is to be protected in management decisions affecting only small portions of a management area (e.g., timber sale areas). There are two other potential weaknesses in the protection of forest ecosystems provided by the Act: biodiversity may be simply measured as the number of species present; or the measure of biodiversity may be deemed too technical for the courts to interpret. In the latter case, the courts would traditionally defer to the agency

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(rather than to a plaintiff ’s expert, for example in a court case) to provide a measure of biodiversity. This is not necessarily bad, because agency biologists are often the most knowledgeable about local biodiversity and critical ecosystem processes. The local appropriateness of the biodiversity measure is more problematic, however, when the selection of biodiversity measures is made at regional or national levels within the agency. Another potential weakness in biodiversity protection would occur if simple numbers of species were used to measure biodiversity. In this case, clearcuts might support more total species than would occur in a similar-sized patch of uncut forest, but only because the majority of the species would be habitat generalists and “edge species.” One way in which bias by the inclusion of generalist species is avoided is through the use of “management indicator species”, as mandated by the National Forest Management Act. These indicator species are chosen to reﬂect the characteristic biodiversity of the management area, and to be sensitive to management practices. For example, the northern spotted owl (Strix occidentalis) is a management indicator species in the Paciﬁc Northwest because it is a characteristic species of old-growth forest and is sensitive to forest-clearing practices. Thus, biodiversity would include two components: management indicator species and numbers of native species. The courts have not adequately resolved the issue of how biodiversity is to be measured. The effects of management decisions are monitored throughout the life of the forest plans, usually 10– 15 years, and revisions to the plan may be made at any time that monitoring reveals signiﬁcant negative changes to timber productivity. The primary intentions of the monitoring provision are to identify lands that (a) can be moved from the category of “unsuitable” to “suitable” for timber harvesting, or (b) that are losing their ability to support timbering activities. Monitoring may incidentally reveal negative changes to Management Indicator Species or evidence of environmental damage. Tuholske and Brennan (1994), in their discussion of suits brought against the National Forest Service by the Citizens for Environmental Quality in 1989 and by the Sierra Club in 1993, pointed out that, where monitoring may reveal negative environmental consequences of timber-management practices, the courts have not required the impacts to be amended. Thus, because the requirement for monitoring is not tied to a requirement for amending any environmental damage disclosed, the National Forest Management

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Act in this instance is subverted from an act of substance to one of procedure only. One may note, however, that, if species listed as federally threatened or endangered occur in the forest management area, they must be listed as Management Indicator Species. In this case, the Endangered Species Act provides additional protection. Where federally listed species are concerned, the management plan of the Forest Service may not place the listed species at further jeopardy, and the Service must designate and protect critical habitat for federally listed species. Here is a case where protection against undue land disturbance is achieved by the intersection of two separate legislative acts. The National Forest Management Act contains strong and explicit language that precludes timber harvesting on lands where even the best management practices would create signiﬁcant risks of severe soil erosion, of physical damage to streams, lakes and wetlands (including ﬁsh habitat), or of degrading catchment conditions. If the Forest Service wishes to remove timber from lands that are classiﬁed as unsuitable, the courts have insisted that the Forest Service must show that technology exists to restore the land and aquatic ecosystems damaged by the harvest to their original state, and that this technology will be made available. On lands classiﬁed as unsuitable for timber harvest, the high cost of implementing restoration technology, even if it existed, would likely prevent many timber harvests on unsuitable lands. Unfortunately, while physical protection of land seems to be interpreted as a substantive statute, the protection afforded waterways and ﬁsh habitat is more ambiguous – again a case where a portion of the act has been interpreted by the courts more as a procedural process than as substantive. Disincentives to ecologically sound forest management The United States Department of Agriculture Appropriation Act of 1908 directed that 25% of the proceeds of timber sales be returned to local communities. The original intent was to stabilize local communities so that timber could be reliably harvested. However, tying community resources to local timber sales acts as an incentive for local constituencies and their political representatives to favor timber sales over environmental protection (Goodman, 1994). Historically, the United States Forest Service has acted in a self-aggrandizing manner to maximize its

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budget, and this institutional behavior has created signiﬁcant environmental risk. Two of several examples of this behavior cited by O’Toole (1988) are road development, and below-market pricing of timber. Before the tenure of Forest Chief Jack Ward Thomas (1993–1996), the Forest Service coupled the costs of building roads with timber values. Roads, however, create two kinds of environmental risks, one resulting from increased erosion and siltation of waterways and the other from increasing fragmentation of forest habitat. In both cases, the risks are not simple increasing linear functions of the number of roads that cross the forest, but are, importantly, step functions. For example, at some point, an ever-increasing rate of silt input to rivers covers spawning beds of salmonid ﬁshes that exceeds the rate at which the silt can be removed by the river’s current. When that point is reached, the net deposition of silt causes the local collapse of ﬁsh reproduction. With regard to an increasing number of roads, two conservation principles are well established. First, some animal species seldom venture across roads and other species, for instance cougars (Felis concolor) in the Rocky Mountains even avoid the proximity of logging roads. Secondly, because roads constrain the mobility of some species, they serve to fragment forest habitats, and continual increases in habitat fragmentation eventually lead to rapid population extinctions for many species (Meffe and Carroll, 1997). Current National Forest policy does not encourage new road construction to the extent that occurred under previous administrations. However, there remains a curious pressure to build roads into previously roadless areas. The Forest Service is required to consider roadless areas for possible designation as “wilderness”, and during the lengthy period of evaluation the roadless area is protected against road building and timber harvests. Thus, there is pressure from timber interests to encourage new road construction into intact forest blocks simply to maintain the land for multiple use, including timber extraction. In pricing timber, the Forest Service essentially uses tax revenue as a subsidy of their production costs, and sets timber prices below those of a free market. A consequence of below-market pricing is greatly to increase consumer demand to harvest timber from public lands and to further fragment the forest habitat through clear-cutting or, more commonly, to degrade the forest habitat through selective harvesting.

700 POLICY OPTIONS AND PROPOSALS

Just as many issues need to be considered and coordinated to minimize land disturbance as shown in our example of the National Forest Management Act, policy makers need to improve their ability to foresee and adapt land-conservation needs to conditions which change rapidly with the demand for commodities generated by land management. For example, increased emphasis on recycling in post-industrial economies will see fewer hectares disturbed by forestry, although requirements for pulp may be replaced by increased demand for grain crops as the world population grows. Our review of the literature has yielded many sound policy recommendations for various management types. We present some of these, and summarize the important principles of these recommendations. For agriculture, a 1993 report by the United States National Research Council recommended policy options to improve the adoption of conservation measures. They are: (1) More research and development, and technology transfer from research and development programs to land-owners, with subsequent voluntary adoption of practices such as contour plowing, forested waterways, reforestation of fragile land, and conservation tillage. (2) More ﬂexibility for site-speciﬁc implementation of policy. This would remove areas on marginal soils from programs for crops inappropriate to these soils. (3) Introduction of market-based incentives for conservation, such as taxes on leaching or runoff of agrochemicals, or permits to purchase and trade in quantities of agrochemical runoff. General recommendations for forest policy in the 1993 report include the development of agriculture policies which reﬂect the role of forests as a source of environmental services essential to food security, and energy policies which include fuel efﬁciency and alternative energy sources (these proposals themselves would need to be evaluated for impacts on non-forested lands). Other recommendations include development of non-market beneﬁts of forest services, realistic development and valuation of non-timber products, charging stumpage fees (the price based on standing timber) and land rents as set by competitive bidding, levying taxes on timber exports, and provisions in long-term leases for sustainable utilization of tracts. In particular, the following ﬁnancial measures are suggested for

Gregory E. ECKERT and C. Ronald CARROLL

improved forest management in developing countries (Poore and Sayer, 1991): taxes which discriminate among species, taxes based on the area of forest harvested, licenses based on the merchantable timber in the tract instead of the volume removed from the tract, incentives payable to logging crews who meet, within narrow limits, the planned harvest rate and protection prescriptions, and leases to logging companies which are at least as long as the cutting cycle. These proposals emphasize stakeholder participation, socioeconomic issues, site-speciﬁc ﬂexibility in management practices, and ecosystem characters such as species and nutrient and energy ﬂows. Another policy approach that should not be neglected in managed areas is to include non-disturbed areas in the management matrix. As human populations grow and shift, better planning will be necessary to minimize impacts of residential, retail, and industrial development on land. To date, this has generally been an issue addressed by local governments in the United States. Several zoning options exist for these bodies to preserve land from development. They include promotion of lot-size reductions, increased density development, transferable development rights, designated urban-growth boundaries, and impact fees for land disturbance – all of which encourage developers to minimize the amount of land actually covered over by buildings and asphalt. The outright dedication of “open spaces” before any proposed development, is also an option at all levels of government to facilitate long-term change in land use. These areas can be turned over to governments, Land Trusts, or Homeowner Associations, and managed under a Conservation Easement Program, in which an agreement not to develop the land is exchanged for reduced tax burdens on the land owner. In some cases, as in Vermont, this would simply keep land in farm activities, or could be used for management of speciﬁc farm areas such as wetlands or riparian areas. Easements could easily complement current voluntary land-management programs, or be further developed from the structure of the Conservation Reserve Program of the United States Department of Agriculture. Another example of “open-space dedication” is the identiﬁcation of key landforms. The Georgia Planning Act of 1989 deﬁnes “Vital Areas” including wetlands, water-supply catchments, groundwater recharge zones, river corridors, and high elevation areas as having particular importance from an environmental standpoint. These areas receive special protection from

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land development, but unfortunately this protection is grossly inadequate, as individual areas designated for protection are small (Kundell et al., 1989). Georgia has made efforts to increase protection of such areas through the Regionally Important Resources provision of the Georgia Planning Act. This provides opportunity for multijurisdictional planning to protect natural, cultural, and historical resources of regional value. Other areas protected under this type of designation in the United States are the Adirondack Mountains, the Florida Keys, the New Jersey Pine Barrens, and the California and North Carolina coasts. As another option to conventional local-level zoning, Oregon has developed a goal-based centralized plan for local governments to follow. An important characteristic is that the interaction between state and local governments is dynamic: the level of state oversight is related to the scale, or impact, of the project (Armstrong and Jacobs, 1996). Finally, federal and state planners can consider bioregional projects. Slocombe (1993) described this as ecosystem-based management of environmental protection and economic development at a regional scale. As examples, he recorded successes in protected areas, such as Glacier National Park/Waterton Biosphere Reserves, and the Australian Alps, in which several governmental bodies coordinate their efforts to manage vast areas, and the more complex Prince William Sound region of Alaska, in which many groups representing governmental, tribal, environmental, logging, and ﬁshing concerns are working together to resolve economic and environmental conﬂicts. Other examples of bioregional projects range from River Basin Commissions to multi-state and national projects such as “Cascadia” located in the Paciﬁc Northwest of the United States (D. McCloskey, pers. comm.). Several generic recommendations to improve environmental policies also are found in the literature. One common call is to improve integration with other policies affecting land use. The World Commission on Environment and Development stressed: “The ability to anticipate and prevent environmental damage requires that the ecological dimensions of policy be considered at the same time as the economic, trade, energy, agricultural, and other dimensions. They should be considered on the same agendas and in the same national and international institutions” (World Commission on Environment and Development, 1987). A second common recommendation is the use of economic forces in environmental management.

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Besides the obvious need for internalization of the costs of land degradation and other forms of environmental degradation, Roodman (1995) called for taxation of the exploitation of resources, and of any modiﬁcation of land and incorporation of the harm to current and future generations – that is, to acknowledge externalities in economic systems, which otherwise would be paid indirectly through health costs and projects for environmental remediation. Tax modiﬁcations need to be developed in conjunction with other types of policy. For instance, higher taxes on automobile emissions need to be developed in step with efforts to discourage zoning policies resulting in housing developments requiring the use of automobiles for daily activities. Southgate and Whitaker (1994) called for reliance on free-market forces, with negative environmental practices internalized, as the alternative to current conditions in Ecuador and other developing countries. One market-based incentive that is currently being evaluated is the open sale of emission permits. These sales result in higher “taxes” for heavy polluters, and savings for industries which pollute less, either through modiﬁcations in process engineering (waste minimization) or investment in treatment technologies for end-of-pipe wastes. “Unlike regulations, market signals would impinge less on personal freedoms, giving individual polluters, or resource depleters, the freedom to decide how much to alter their practices, and how to go about doing it” (Roodman, 1995). Gradually, the amount of total permits available, proportional to the total allowable emissions, would be decreased. Market forces would then encourage the development of alternative, “cleaner” inputs and processes. These efforts, of course, would not replace the need for regulations that prohibit certain activities, such as the manufacture of extremely toxic chemicals, or the setting of the level or total number of pollutant allowances that would be marketed. Strong liability and penalty provisions in environmental laws should not be weakened in the effort to incorporate free-market activities into environmental management. Responsibility and liability for environmental damage should continue to include those who perform the damage, and also those parties who inﬂuence, through contract or production of contaminants, the activities of those who perform environmental damage. As an example, meat processors often have great inﬂuence over farmer land-management decisions because of contractual production obligations, and these obligations may take precedence over conservation programs

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(National Research Council, 1993). These types of policies can facilitate the internalization of the costs of direct and indirect effects of production and landuse practices. Although considered by economists as a cheaper solution to achieving the environmental goals than standard-based compliance programs, market incentives have mainly focused on emission abatement, not on overall land use and disturbance. This has recently been attempted in Wisconsin and Colorado, in which land managers have been allowed to sell permits to discharge phosphate to point-source dischargers in return for improved management of phosphate runoff from non-point sources. Actual monitoring costs, and ethical questions of using the permit system, have yet to be addressed (Kelman, 1981). A recent example of the controversy related to this approach is an offer by the State of New York of pollution credits, which it had purchased and stored, to a company in an effort to attract more industry to the State (New York Times, 1997). While this will not change the total amount of pollutants allowed in a particular air district, the effort has raised many concerns among citizens and environmentalists as to the State’s interest in protecting public health and the environment. These concerns also apply to valuation and internalization of costs in cost/beneﬁt analyses. We feel that monitoring and adaptive management provisions of ecosystem management can help to address these issues. Interestingly, while the development of an environmental, or land ethic is seen as an educational issue, there are calls to clarify the value placed by society on land resources for long-term use by declaring landowner responsibilities and rights. Thus, there would be a legal responsibility of landowners and land-users to manage their lands in ways that do not degrade soil and water quality, as codiﬁed by federal and state laws (National Research Council, 1993). Although this may require an enormous shift in long-held attitudes towards private property rights, and the “utility-based” perception of land (Barrow, 1991) the National Research Council argues the advantages of shifting the burden of land protection from the government to the land-owner. Another common recommendation is the call for improved use of the existing scientiﬁc base to address issues of land management and degradation before they require more of the standard “regulatory ﬁx.” While ecosystem management involves stakeholders, the policy arena needs to improve interactions with another group, the scientiﬁc community. This need is

Gregory E. ECKERT and C. Ronald CARROLL

especially important in developing countries. Understanding how to manage agricultural lands optimally within their ecological limits for the food and ﬁber needs must be addressed, or there will be no incentive for these people to value the long-term conservation of neotropical forests (Southgate and Clark, 1993). International research centers have the capacity for research related to both policy development and the introduction of new land-use technologies, such as (1) to improve understanding of why farmers and communities degrade resources; (2) to increase incentives and investment into the adoption of sustainable technologies and practices; (3) to establish priorities in a long-term growth strategy for the ecological systems of a given area, including non-agricultural opportunities (Sivakumar and Wills, 1995). As with any restorative activity, policy implementation must go beyond patching symptoms. Research into the underlying causes of land degradation can yield important information to policymakers, and later assist workers charged with implementing policy. In the case of Madagascar, Larson (1994) learned that, while burning plant stubble on slopes appeared to have a negative inﬂuence on the land, it was a management practice designed to increase water supply to valuable lowland crops. This type of information would hopefully save extension workers the embarrassment of ordering farmers to cut off an important crop input without having ﬁrst developed an alternative appropriate to both soil and irrigation needs. This example again highlights the importance of including those affected by policies in the process of policy development. These groups can often supply speciﬁc information for policy development, and their involvement in turn increases the likelihood that a recommended practice will be adopted. As responsibility for environmental programs is shifted to an adaptive management strategy at site-speciﬁc levels (i.e., away from centralized planning), the need for technical expertise will become greater. Links between state and local governments, or individual regional parks and academic researchers will need to be strengthened. Appropriate models for this exist, such as short-term assignments within government agencies to allow research personnel to work with site program managers or headquarters policy-writers. Examples also exist of government personnel spending up to a year with private ﬁrms. Finally, a promising sign of improved interaction between science and policy in the United States is the establishment of an Environmental Futures Committee

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made up of members of the Science Advisory Board of the Environmental Protection Agency. The mandate of this committee is to develop means for the Agency to anticipate environmental issues of the future better, and to recommend actions which would reduce the severity of such issues or to avoid them altogether. We recommend that Congress and State legislatures in the United States work with groups like the Ecological Society of America in the same vein, in an attempt to go beyond the tradition of reacting to environmental crises on a case-by-case basis.

CONCLUSION

We believe that environmental policy, to be successful in the long term, must reﬂect biological and socioeconomic realities. The ecosystem, in the sense in which we deﬁne it in Table 30.1, is the management unit that best reﬂects these realities. The concept of the ecosystem as a life-support system emphasizes interactions, linkages, and scale. An ecosystem approach seems most appropriate for conserving biodiversity, sheltering threatened and endangered species, including relevant human concerns, valuing environmental services, and managing impacts that originate outside the ecosystem. It also facilitates integration of issueoriented policies and the internalization of the effects of land-use activities, especially over long time periods. Of the many characteristics of ecosystems, we believe that the following ﬁve are most relevant to environmental policy: (1) Ecosystems are dynamic, not static, and contain many indirect effects and non-linear processes. These give rise to critical biological thresholds and unexpected responses to stress or management interventions. (2) Ecosystems are open and therefore linked to the larger biophysical landscape. (3) There is no single spatial or temporal scale that is appropriate for all ecosystem patterns and processes. Rather, there are suites of appropriate scales, where selection of the appropriate ones depends on particular management goals. (4) Natural ecosystems are usually small relative to the magnitude of land areas used to meet human economic needs and desires. Hence, external threats are often large. (5) Ecosystems are resilient to stresses, and these resiliency properties differ according to the type

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and scale of the stresses. Resiliency properties can be weakened by cumulative effects of small but chronic stresses. In our view, environmental policy should reﬂect and support these ﬁve attributes that characterize ecosystems. Therefore, environmental policy, whether derived from public law or from non-governmental sources, should make use of an adaptive management model that emphasizes the following seven perspectives: (1) Management is viewed as an experimental, ﬂexible process that reﬂects site-speciﬁc needs, constraints, and opportunities. (2) Long-term monitoring to validate management decisions is essential. (3) Experiments at an appropriate scale designed to deﬁne the resiliency properties of ecosystems should be facilitated. Policy should be periodically revised to incorporate new information about resiliency properties. New empirical and theoretical information on resiliency should be used to deﬁne precautionary principles for speciﬁc ecosystems and suites of similar stressors. (4) Public law should provide context for environmental policy, but community participation should give shape to the implementation of policy at regional or local levels. But accountability, perhaps through oversight committees, is needed to ensure that local implementation properly reﬂects the state and national context of environmental policy. In other words, local implementation should not violate the spirit and intent of public law. (5) Cumulative effects should be included in monitoring programs. (6) Restoration should be viewed as part of the continuum of adaptive management. (7) Adaptive management and monitoring plans should be treated as peer-reviewed proposals, much like those submitted to the National Science Foundation in the United States. Ideally, the composition of regional peer-review panels should be a mix of stakeholder interests including scientists from government agencies, university researchers, nongovernment agencies, and private citizens. A ﬁnal caveat Ultimately, public policy that expresses the public’s will is the product of strong social capital, the capacity of people to work collectively towards shared goals. When social capital is weak, the development of public policy

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becomes the exclusive domain of the government. When social capital is vested in a few privileged groups, then public policy is unduly shaped by the privileged and is a distortion of the public’s will. As we have argued throughout this chapter, public policy guiding natural-resource protection, management, and restoration should be grounded in a rigorous scientiﬁc understanding of ecosystem dynamics, and applied at appropriate scales. We have also argued that, in order for ecosystem management policy to last, it must also reﬂect the many varied viewpoints of the stakeholders. However, some commentators have called attention to the decline of social capital, at least in the United States (Putnam, 1995), and this decline does not bode well for stakeholder participation in the formation of public policy. It is easy to identify some of the factors that weaken social capital by creating disincentives to collective action. (1) Time is limiting as families become increasingly involved in competing social activities or in the need to generate two incomes in order to meet rising economic demands. (2) Intolerance towards other viewpoints weakens social capital by preventing the evolution of coalitions. For example, antagonism between hunters and “environmentalists” diminishes their effectiveness in working towards policy that would protect natural areas from destruction although this is likely a strongly shared goal. (3) In a similar vein, the news media often polarize public opinion about environmental issues by presenting colorful quotations from those who represent extreme positions, and by failing to meet the more difﬁcult tasks of providing context, background, and analysis. Polarization among stakeholders further diminishes the possibility of building social capital through coalitions. For example, the news media largely portrayed conﬂict over logging of old-growth Douglas-ﬁr (Pseudotsuga menziesii) in the Paciﬁc Northwest as a battle between loggers and the protectors of the northern spotted owl, generally ignoring the root causes of the declining regional timber-based economy (Sherman and Carroll, 1997). (4) The failure of the academic community to communicate effectively with the public contributes to public apathy (through a sense that the issues are dauntingly complex and frustratingly ambiguous)

Gregory E. ECKERT and C. Ronald CARROLL

or, as a reaction, encourages people to embrace overly simplistic solutions. Whatever the causes, the continued erosion of social capital is likely to become the single greatest threat to the development of sound environmental policy.

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ENVIRONMENTAL POLICIES AS INCENTIVES AND DISINCENTIVES TO LAND DISTURBANCE Lesica, P. and DeLuca, T.H., 1996. Long-term harmful effects of crested wheatgrass on Great Plains grassland ecosystems. J. Soil Water Conserv., 51: 408–409. Mahar, D.J., 1989. Government Policies and Deforestation in Brasil’s Amazon Region. World Bank, Washington D.C., 56 pp. Mary, F. and Michon, G., 1987. When agroforestry drives back natural forests: A socio-economic analysis of a rice–agroforest system in Sumatra. Agrofor. Syst., 5: 27–55. Meffe, G.K. and Carroll, C.R., 1997. Principles of Conservation Biology, 2nd Edition. Sinauer Associates, Inc., Sunderland, Massachusetts, 728 pp. National Academy of Sciences, 1992. Water Transfers in the West: Efﬁciency, Equity, and the Environment. National Academy Press, Washington, D.C., 300 pp. National Research Council, 1993. Soil and Water Quality: An Agenda for Agriculture. National Academy Press, Washington, D.C., 516 pp. New York Times, 1997. New York offers pollution permits to lure companies. New York Times, May 19, 951: 1. Orians, G.H., 1990. Ecological concepts of sustainability. Environment, 32: 11–15; 34–39. O’Toole, R., 1988. Reforming the Forest Service. Island Press, Washington, D.C., 247 pp. Platter, Z.J.B., Abrams, R.H. and Goldfarb, W., 1994. Environmental Law and Policy. American Casebook Series. West Publishing Co., St. Paul, Minnesota, 651 pp. Poore, D. and Sayer, J., 1991. The Management of Tropical Moist Forest Lands: Ecological Guidelines. World Conservation Union (IUCN), 69 pp. Pulliam, H.R. and Haddad, N.M., 1994. Human population growth and the carrying capacity concept. Bull. Ecol. Soc. Am., 75: 141–157. Putnam, R.D., 1995. Bowling alone: America’s declining social capital. J. Democracy, 6: 65–78. Reid, W.V.C., 1989. Sustainable development: Lessons from success. Environment, 31: 6–9; 29–35.

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Roodman, D.M., 1995. Public money and future purpose: The future of taxes. World Watch, 8: 10–19. Sherman, E. and Carroll, C.R., 1997. Newspaper coverage of the Spotted Owl/Old-Growth forest controversy: information or polarization? In: G.K. Meffe and C.R. Carroll (Editors), Principles of Conservation Biology, 2nd Ed. Sinaver, Sunderland, Massachusetts, pp. 560–561. Sivakumar, M.V.K. and Wills, J.B. (Editors), 1995. Combating Land Degradation in Sub-Saharan Africa: Summary Proceedings from the International Planning Workshop for a Desert Margins Initiative. ICRISAT, Nairobi, 47 pp. Slocombe, D.S., 1993. Implementing ecosystem-based management. BioScience, 43: 612–622. Southgate, D. and Clark, H.L., 1993. Can conservation projects save biodiversity in South America? Ambio, 22: 163–166. Southgate, D. and Whitaker, M., 1994. Economic Progress and the Environment: One Developing Country’s Policy Crisis. Oxford University Press, Oxford, 150 pp. Tuholske, J. and Brennan, B., 1994. The National Forest Management Act: Judicial Interpretation of a Substantive Environmental Statute. Public Land Law Rev., 15: 53–134. van Wilgen, B.W., Cowling, R.M. and Burgers, C.T., 1996. Valuation of ecosystem services: A case study from South African fynbos ecosystems. BioScience, 46: 184–189. Veblen, T.T., 1985. Stand dynamics in Chilean Nothofagus forests. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, New York, pp. 35–51. Vitousek, P.M. and Hooper, D.U., 1993. Biological diversity and terrestrial ecosystem biogeochemistry. In: E.-D. Schulze and H.A. Mooney (Editors), Biodiversity and Ecosystem Function. SpringerVerlag, Berlin, pp. 3–14. Weinberg, P., 1994. It’s time to put NEPA back on course. N. Y. Environ. Law J., 3: 99–116. World Commission on Environment and Development, 1987. Our Common Future. Oxford University Press, Oxford, 400 pp.

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Chapter 31

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS S.T.A. PICKETT, Jianguo WU and M.L. CADENASSO

INTRODUCTION

The purpose of this chapter is to present a framework for patch dynamics and to indicate how it clariﬁes and uniﬁes several of the important themes in this volume. Although disturbance is a familiar concept, there have been signiﬁcant reﬁnements to the concept since its widespread adoption in the 1970s (e.g., Loucks, 1970; Levin and Paine, 1975; Pickett and Thompson, 1978; Pickett and White, 1985). In addition, as ecologists have accumulated data on more and more kinds of ecological systems, spatial scales of study, long time spans, and kinds of disturbance events, new insights and hypotheses have emerged (Wu and Levin, 1994). This chapter will point to illustrative examples of studies that have led to new or reﬁned insights, and combine the empirical and conceptual approaches taken, to present a contemporary view of patch dynamics which is capable of putting the ecosystems of disturbed ground in their most complete spatial and dynamic context. The differences between the contemporary perspective we present and the earlier ideas and understanding of disturbance and its effects are sometimes subtle. However, the emerging appreciation of the rich and varied role of ecosystems of disturbed ground is based on just that subtlety. This chapter relies on the depth of empirical examples presented in the book, and emphasizes the framework rather than presenting a complete literature review. The main themes we address are: (1) that disturbance necessarily has a spatial context in which the ecology of disturbed ground can be understood; (2) there are limits to employing the intuitive concept of disturbance that ﬁrst emerged in community

(3)

(4)

(5)

(6)

(7)

ecology at other scales and in other kinds of ecological systems; added generality and comparability of disturbances and studies of disturbed systems result from a more rigorous and scale-independent concept of disturbance; structural models of ecological systems are crucial for discriminating disturbance from other kinds of alterations of ecological systems; spatial heterogeneity in disturbed ground and systems at a variety of spatial scales is functionally important; the contemporary patch-dynamics concept accommodates the foregoing insights in addition to others presented in the chapter, to enhance unity and dynamic scope in the study of disturbed ground ecosystems; and ﬁnally, we indicate the signiﬁcance of the hierarchical, patch-dynamic approach to disturbed-ground ecological systems, which allows their potential contribution to stability of more extensive ecological systems to be evaluated and used in management.

DISTURBANCE AS A SPATIAL CONCEPT

Here we deﬁne disturbance and give a few examples to ﬂesh out the abstract deﬁnition. Because disturbance is a physical force or event that disrupts the physical or biological structure of an ecological system (Fig. 31.1), there will be a geographic location where that disruption has occurred. Hence, the concept of disturbance is subtle because it implies an event, a location or affected system, and a result. At some spatial resolution, it is possible to delimit the disturbed site from the adjacent undisturbed areas. For instance, a population that has

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Fig. 31.1. Three crucial features to be speciﬁed in understanding the process of disturbance. The scale domain refers to the temporal and spatial extent and grain by which the system of interest is described. The event that may potentially disturb the system arises from outside of the system and its speciﬁed scale domain. Additional features are described later in association with the construction of a systems structural model to assess disturbance.

experienced disturbance may show a gap with a density of zero, surrounded by sites in which the population density remains high. Or a desert stream may exhibit some patches in which the bed has been scoured or the banks have collapsed, removing patches of stream habitat or the riparian vegetation that formerly occupied those sites (Fisher et al., 1998). Or a large tree may be blown down in a closed-canopy forest, leaving a gap in which the vertical structure of the community, from the rooting zone upwards, is now altered. As a ﬁnal example of disturbance, a portion of a landscape in which humans exert energy to change the temporal pattern or frequency of ﬁre can come, over a short time, to exhibit an altered structure. All of these are examples of disturbance that affects a recognizable physical location, which is the basis for the focus of this book on ecosystems of disturbed ground. On the time scale at which dynamics and persistence of the system are of interest, locations have been affected by some force external to that location interacting with the structure of the system to alter the existing structure and create something new. In those new, disturbed patches, ecological processes and species composition may differ from the predisturbance conditions. In order to understand ecological systems that have been subjected to disturbance, the alterations of composition and processes by disturbance must be known. Understanding change in composition and processes has practical implications for management, conservation, sustainable use, or restoration of such systems (Hobbs, 1987; Boeken and Shachak, 1994; Arnold, 1995). Because disturbance has spatial extent and limits, related concepts are needed to understand disturbance

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and the ecology of disturbed ground. These will be introduced, deﬁned, and exempliﬁed throughout the chapter. Here we indicate how they will be related to one another and to the central concept of disturbance. The auxiliary concepts include heterogeneity, landscape context, and patch dynamics. Disturbance creates patches recognizable at certain scales. Disturbance acts with various intensities, and hence creates heterogeneity in space. Therefore knowing the landscape context of disturbance is crucial for determining what events are or are not disturbances, how different potentially disturbing forces act and interact, how disturbance may appear through space, how ﬂuxes of organisms and materials contrast between disturbed and undisturbed sites, and how different parts of landscapes respond to disturbance. In addressing these issues, the concept of landscape is taken in the most general sense as a criterion of observation (Allen and Hoekstra, 1992). A landscape is a heterogeneous spatial array of ecological patches at any scale. The emphasis is on heterogeneity rather than on a particular spatial scale (Turner, 1989). All of these aspects of systems that experience disturbance suggest that the emerging understanding of disturbance is more subtle than the original concept, despite the successes of that rather intuitive understanding. The ﬁnal concept to be discussed below will be patch dynamics, which will be used to integrate the other concepts. Beyond the intuitive Disturbance is a deceptively simple and intuitive, yet important, idea. It is relevant to the ecology of the full range of Earth’s biomes, but the concept originated with the experience of ecologists at the community level of organization (Watt, 1947). Community ecologists noticed sudden disruptions of community structure by powerful natural forces originating outside the community (Sousa, 1984). The impacts of disturbance were obvious because the scale of the observations at the community level are close to the human scales in time and space of decades and meters. Ecologists could literally enter the aboveground structure of many communities, especially forests and woodlands, while in other systems, such as grasslands or mussel beds, it was easy to comprehend the structure from above. The condition of the system before the disturbance, the magnitude and direction of the structural and compositional changes caused by the disturbance, and the extent of the reorganization that ensued, were

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS

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Fig. 31.2. Hierarchies of patches created by different kinds of disturbances. Patchiness is divided into physical and biological categories. Within biological patches, vegetation processes and the effects of consumers can generate patchiness. Each of the speciﬁc types of patchiness can be generated by a variety of mechanisms. The process of disturbance is not listed separately since it acts through the other speciﬁc phenomena listed. From Wu and Loucks (1995).

all obvious and intuitive. Ecologists recognized that the closed architecture and layering of organisms was functionally important, and controlled the access of different species or layers to limiting resources (Horn, 1974; Allen and Forman, 1976). Furthermore, ecologists could visit the same systems year after year, or even after decades, and observe little change. In such situations, sudden disruption of community structure was conspicuous and extreme (Falinski, 1978; Canham et al., 1990). Physical disruption of some forests,

grasslands, or mussel beds, after long periods in more or less the same structural and compositional state, intrigued ecologists and required a conceptual label. This led to the adoption of a common language term – disturbance – in a specialized, technical sense (Pickett et al., 1989). The commonsense, intuitive connotations have unfortunately persisted. When ecologists began to apply the intuitive concept of disturbance beyond the usual temporal and spatial scales of community ecology (Fig. 31.2), the need for

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increased rigor became apparent. If disturbance is to be truly valuable in unifying knowledge about a wide variety of ecological systems and processes, such as the breadth encompassed in this volume, the rigorous and exact aspects of the concept must be emphasized over the intuitive (Rykiel, 1985). Furthermore, disturbance must be put in a framework that shows (1) how it can interact with ecological systems observed by different criteria (sensu Allen and Hoekstra, 1992) or at various hierarchical levels; (2) what role it plays at various ecological scales, including large and small heterogeneous landscapes; (3) how it connects with the processes of succession; and (4) how diffuse disturbances work and how they relate to the distinct concept of stress. Although the concept of disturbance has been most widely used at the scale of meters and years on which communities are usually studied, smaller systems can also experience disturbance (Cofﬁn and Lauenroth, 1989; Wu and Loucks, 1995). One reason to seek increased rigor and subtlety in the concept of disturbance is that ecological systems of vastly different spatial extents and grains experience disturbance. If generality and contrast are to be discovered in the understanding of disturbed ground, the concept must be applicable across scales. One may recall that the basic deﬁnition of disturbance is the sudden disruption of the structure of an ecological system. Old-ﬁeld vegetation, though not so intimately experienced by most ecologists from beneath its canopy as forests, can also undergo canopy-gap formation (Goldberg and Gross, 1988), soil turnover by frost heaving or animal activities (Korn, 1991), and so on. On still smaller scales, the structure of the canopy of a single plant can be disturbed. For example, if the nipping of buds to consume spring sap by arboreal mammals reﬂects outbreak levels of the herbivore population, its sudden onset and episodic pattern may make it ﬁt the basic deﬁnition of disturbance (Pickett et al., 1989) for the array of species – the ecological system in this case – comprising a network of host-speciﬁc species or other organisms depending on that plant individual. Such alteration may make opportunities available for new epiphytes, or change the resource ﬂux to organisms beneath the disturbed canopy. On smaller scales still, damage to a leaf by a herbivore constitutes a disturbance to the system of resident fungi. The effects of such events can also spread throughout a more extensive network of interacting species, leading to a change in structure of larger

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ecological systems of which individual plants or small sites are a component. The structural changes resulting from small disturbances at the individual plant level can alter resource ﬂux in the larger system and redistribute resources. Small, inconspicuous disturbances at the individual level can propagate so that they lead to disturbances to larger systems. Keeping the concept of disturbance tied to the scale of meters and years on which ecologists typically study communities satisﬁes the intuitive view of disturbance, but it does not permit broad hypotheses about the generality of occurrence, control, or effect of disturbance to be tested. Toward generality and comparability If the ecology of systems on disturbed ground is to be fully understood, the relationships of disturbance at different scales and in different kinds of ecological systems also must be understood. Moving beyond the intuitive connotations of the concept of disturbance requires ecologists to examine how disturbance applies to other scales and kinds of systems (Petraitis et al., 1989). Disturbed ground can support all the kinds of systems exempliﬁed in this section. It is important to recognize that focus on disturbed ground does not presuppose that one of these kinds of systems rather than the others is the object of study. Populations are among the ecological units other than communities that can experience disturbance. The sudden alteration of population density, size, or age structure is an example of disturbance. Often patches of reduced density or with other structural changes result. The outbreak of a disease, say in a population of oaks (Quercus spp.), or the mortality of a band of spruce trees (Picea spp.) on a mountain side are population disturbances. Of course, these can also be expressed at the associated community or ecosystem realms. The mortality of seedlings smaller than a certain size threshold due to frost heaving, ﬁre, or browsing, is a common disturbance expressed in the population realm. One may note that many disturbances in the population sphere may be seen by population and evolutionary ecologists as strong selection episodes (Endler, 1986). However, these events still encapsulate the essence of disturbance. Ecosystem disturbances are those physical events that affect the pathways by which matter or energy ﬂow in a system. Focusing on this criterion of ecological observation, patches appear as a result of some of the

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS

same kinds of events that drive disturbance in communities. However, the ecological structures and processes of interest in the site differ from or extend beyond those typical of community ecology. Catchment alterations, such as the experiments at Hubbard Brook (Likens et al., 1978; Likens, 1984; Likens and Bormann, 1995), change ecosystem structure in a patchy fashion, altering export of nutrients and particulates in stream or lake water. Wind-throw creates patches in coniferous-forest ecosystems in the Paciﬁc Northwest of the United States, changing the amount of coarse woody debris on the ground and altering the decomposition rates in the ecosystem (Harmon et al., 1986). If some of the coarse woody debris created by such a forest disturbance accumulates at certain spots in a stream channel, it can act as a dam, suddenly alter the form of the channel, and concentrate sediment and organic matter, which in turn alters the ecosystem carbon budget (Hedin, 1990). An outbreak of grazing animals converts biomass in the canopy to biomass in litter, and shifts the relative roles of the detritus and grazer food chains in speciﬁc spatial locations in an ecosystem (Risley and Crossley, 1988). The landscape criterion is one in which spatial heterogeneity is the underlying principle. One approach to landscapes is to consider them to exist primarily at physical scales where human populations, institutions, and settlements are obvious structural components. Hence, Forman and Godron (1986) deﬁned landscapes as being kilometers wide in extent. However, that scale is only one instance of landscapes, which are spatial arrays resolvable as patches that can exist on any spatial scale. Therefore, in order for landscape theory to apply, spatial heterogeneity, expressed as a variety of patches, must be considered, as well as the processes that create and connect the patches. To ask whether a landscape has been disturbed, therefore, is to ask whether the physical structure of heterogeneity is changed. The issue is not whether a ﬁre creates a new patch, or whether succession converts one patch type to another, which could represent the community realm. Rather, the concern is whether some event has altered the mixture of patches or the kinds of connections among them. Disturbances that can wreak such changes are alterations of the patterns or agents of patch creation, or alterations of the spatial conﬁguration of patches. Landscape disturbances, because they deal with mosaic structure, are likely to be of two sorts (Forman and Moore, 1992). One is the creation of barriers within the landscape. The other is the alteration of

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the processes that create the kinds of patches making up the landscape. The latter requires some sort of physical force originating climatically, geomorphically, or anthropogenically, and thus requires expenditure of energy or displacement of materials at various locations in the landscape (Swanson et al., 1988). Because the forces that generate patches often also appear on the community or ecosystem level as disturbances, when a large-scale landscape is considered to be composed of communities or ecosystems, then the alteration of the structure of the landscape – disturbance of the landscape – is a larger-scale disturbance made up of the pattern of the smaller-scale events that might themselves be deemed disturbances on the ﬁner scale of communities or ecosystems nested within a landscape. The important reﬁnement for applying the concept of disturbance to scales and criteria beyond the community is to recognize that whether an event acts as a disturbance or not depends both on the scale and intensity of the event, and on the nature of the system. We explore this relationship in the next section. System structural models Two important ideas emerge from the examples of disturbance at various scales and ecological realms. First, what is considered a disturbance depends very much on the nature of the system under consideration (Rykiel, 1985). Therefore, before disturbance is evaluated and studied, the system of concern must be clearly described and a model of the system structure explicitly stated (Pickett et al., 1989). A model of the system relevant to understanding whether the concept of disturbance applies to that system and event tells several important things: (1) What are the components of the system, including physical structures and interactions among them? (2) What temporal and spatial scale does the system occupy? (3) What is the nature of the boundaries in the system – that is, how permeable are they, and what ﬂuxes move across the boundaries? It should be noted that we are referring to a static, structural model of the system to make the role of events that might be disturbances unequivocally apparent. Other kinds of models of the same system can be constructed for other purposes. Often arguments and uncertainties about what is or is not a disturbance, and whether the disturbance emerges from inside or outside the boundaries of a system, result

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Fig. 31.3. The structure of a nested hierarchy. Because disturbance is a direct disruption of structure of a particular hierarchical level, a hierarchical conceptual model is required to assess what events are disturbances. Agents that operate with different frequencies and rates must necessarily act at different levels of the hierarchy. Stress at a lower hierarchical level may cause the physical disruption of a higher-level structure, and hence generate disturbance at that higher level. Modiﬁed from various concepts in O’Neill et al. (1986), and Urban et al. (1987).

from the error of beginning to evaluate disturbance before the model of system structure is speciﬁed and clearly stated. Whether the system model is complete, and whether it is appropriate to the research question and goal at hand, are always legitimate questions, of course. But the criteria for evaluating disturbance emerge from combining the model of the structure of the system with the fundamental deﬁnition of disturbance as a physical disruption of the structure of a system. Therefore, we can now reﬁne the deﬁnition of disturbance by stating that it is the physical disruption of the structure of a system as speciﬁed by an explicit, if preliminary, model. A second idea also emerges from the examples. What constitutes a disturbance at one scale or for one ecological phenomenon, may not constitute a disturbance for others (Pickett et al., 1989). Hence, disturbance for a population may be invisible to the community of which that population is a part. Disturbance to an ecosystem may be negligible to the large landscape of which the ecosystem is a part. Disturbances at ﬁner scales or lower levels, or more specialized phenomena, often become part of the structure of the larger scales, higher levels,

or more general phenomena. This phenomenon is called incorporation of disturbance (O’Neill et al., 1986). Arguments about whether a disturbance event is normal (i.e., has been incorporated) or not, often arise from a careless mixing of scales, criteria, or levels. Keeping such important fundamentals clear is another task that an explicit system structural model can accomplish. Indeed, linking systems models as parts of a nested hierarchy is a helpful strategy for assessing disturbances in closely related systems (Fig. 31.3).

STRUCTURAL SOPHISTICATION IN EVALUATING DISTURBANCE

Because the purpose of this chapter is to provide a conceptual framework for synthesizing key insights about the ecology of disturbed ground, we have so far emphasized ongoing conceptual clariﬁcation and rigor in applying the concept across scales and kinds of systems. In this section, we emphasize some of the insights that emerge from the empirical study of disturbed ground. We rely on the substantive review

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS

provided by the foregoing chapters, and cite only a few illustrative cases here. Clements (1916), early in his pioneering synthesis of succession, introduced the need to understand disturbance. He called the process “nudation”, and took it to be the initiation of the complex process of succession. Unfortunately, Clements’s term can connote that nudation creates a uniform, blank slate (e.g., Oosting, 1956), and is therefore of interest only as the starting point for the phenomenon of restoring the vegetation climax in an area. One important feature of Clements’s (1916) idea of disturbance was the recognition that events of great severity could initiate primary succession in which a biotic and soils legacy was lacking, while more modest disturbances would generate secondary successions that began with some biotic or soils capital. Of course, contemporary theory recognizes that these two processes are ends of a continuum (Miles and Walton, 1993). This section highlights insights from the concentrated study of disturbance since its widespread recognition beginning roughly in the 1970s (Loucks, 1970). That empirical effort has shown that the structure of disturbed sites is both complex and functionally signiﬁcant. In general, because heterogeneity is so important to the structure and functioning of ecological systems (Chesson, 1986), especially in the maintenance of biodiversity (Huston, 1994), it is crucial to understand the pattern created, ampliﬁed, or reduced by disturbance. In fact, the heterogeneity of disturbed ground may be one of its most important ecological features. Heterogeneity Heterogeneity affects the role and impact of disturbance in ecological systems. Contrary to the simplicity and uniformity connoted by Clements’s concept of nudation, most disturbed patches are themselves heterogeneous. For instance, burned areas at all scales, from the Yellowstone Plateau on the coarse scale (Knight and Wallace, 1989), through the pitch pine (Pinus rigida) plains in the New Jersey Pinelands (Forman and Boerner, 1981), to the small prairie remnant (Anderson, 1990; Collins and Wallace, 1990), are patchy. Owing to a variety of climatic and topographic factors, and to differences in ﬁre behavior and fuel loading, there are spots that do not burn at all, or that burn with only low intensity, as well as patches in which all biota are consumed (Christensen et al., 1989). The post-ﬁre soil conditions often reﬂect

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the heterogeneity of the burn. A particularly telling case of heterogeneity is the enormous variety of conditions and substrates created by the 1980 eruption of Mount St. Helens in Washington State (Del Moral, 1993; see also Chapter 5, this volume). The textbook deﬁnition of primary succession led many ecologists to assume that Mount St. Helens would straightforwardly illustrate primary succession. On the contrary, the area was immensely complex. There were areas of intact vegetation, mudﬂows, pumice plains, lake overwashes, and many other types of disturbance. The contribution of legacies and survivors was notable (MacMahon, 1982). In other volcanic situations, the ﬂow of lava from a speciﬁc eruption rarely coats an area uniformly. Wind disturbances are also diverse in their patchiness. The damage to a forest canopy by a hurricane may vary over a matter of tens of meters only (Boose et al., 1994). The uprooting of a single tree by wind will leave a pit from which the roots were wrenched, a mound of soil wasting from the exhumed roots, and fallen trunk and branches, often with leaves concentrated in speciﬁc spots. The form of the pit and mound, and hence their suitability for different colonists, can depend on exactly how the tree was uprooted (Beatty and Stone, 1986). Other wind-throws, such as long tracks created by severe tornadoes (Peterson and Pickett, 1995) or fans created by down-drafts (Dunn et al., 1983; Nelson et al., 1994) can be quite large. Depending on the topography they cross, and the condition of the forests they encounter, there can be pits and mounds of differing sizes, fallen trunks, snapped trees, tangles of crown debris and patches of leaf litter, patches of surviving shrubs, broadleaved herbs, tree seedlings, or ferns (Peterson and Pickett, 1990). Animal-generated patches can also be heterogenous. A porcupine (Hystrix indica) digging in a Middle Eastern desert creates a pit, and a small mound of earth nearby (Boeken et al., 1995). Some organic matter, and even some surviving bulbs, appear in spots within the pit. Wallows of grazing animals are also not uniform across their extent. The heterogeneity generated by disturbance and persisting in disturbed ground is important for a variety of ecological processes as mentioned above (e.g., Walker et al., 1991). Here we expand on that idea. Invasion of new propagules may concentrate in certain patches within the disturbed site (Fox and Fox, 1986). Regenerating seedlings may survive better in debris piles. Resources may only become available in some sites, such as hummocks left after wetland disturbance,

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which serve as safe sites for establishment. Likewise, resources may concentrate or persist only in certain sites, such as the depressions created by tree-fall. Such within-gap heterogeneity is crucial to seedling performance (Nu˜nez-Farfan and Dirzo, 1988). Flows of limiting resources may be profoundly inﬂuenced by the patch structure created by disturbance. The debris dams in streams already mentioned are key examples, but on various scales, the susceptibility of streams to disturbance (Reice, 1994) results in great heterogeneity. The array of sand bars, reed beds, and bank types is sensitive to disturbance, and is important for use of arid-zone riparian and stream systems by wildlife (Rogers, 1997). Stream-side habitats have conservation value for rare but relatively fugitive plants in temperate regions as well (Menges, 1990). Diffuse disturbance So far, we have focused on the patchiness created by disturbance. There is another way that disturbance may be manifested. Rather than being discrete and patchy at a particular scale, disturbance may have diffuse results. For instance, in certain sites, a hurricane may not topple any trees, yet leave the forest canopy considerably more open than before, and result in a new resource regime in the understory (Walker, 1991). Similarly, an unusual early wet snow on a deciduous forest canopy may break many branches and thin the canopy throughout the forest. Or a severe drought in a grassland may kill many plants and leave the survivors smaller than during wellwatered periods (Weaver and Albertson, 1943). All these examples – driven by such factors as wind, snow, ice, and drought – are cases of diffuse disturbance. Disturbance that increases structural heterogeneity at a particular scale without creating discrete gaps at that scale are called diffuse disturbances. Of course, at a coarser scale of observation, the disturbed areas that are diffuse at the ﬁner scale may appear as patches distinct from the adjoining lands. The phrase “diffuse disturbance” emphasizes that, like so many ecological concepts, disturbances form a continuum. Disturbance, at a particular scale, ranges from discrete to diffuse. The same agents of disturbance may have diffuse or discrete effects depending on where in geographic space they occur, and the degree and spatial pattern of susceptibility of the system to the agent. An important insight to arise from the study of diffuse disturbance is that stress can sometimes be the cause of disturbance. We ﬁt these cases to the

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basic deﬁnition of disturbance because the initial cause arises as a coarser-scale event triggered from outside the area of interest resulting in structural change. For example, the structural disturbance caused by droughtinduced mortality is sudden in terms of the longterm status of the grassland, and was initiated as severe stress to individual plants. This is an example of transformation of stress into disturbance. Stress refers to a factor that directly alters the function of an ecological system. The root of the stress concept in biology comes from physiology, where stress is a factor that alters the metabolic rates or physiological processes of an organism. Keeping the useful dichotomy between structure and function in place among various ecological criteria of observation, not just that of the individual organism, suggests that disturbance affects structure directly, whereas stress directly affects function. Altered structure can affect function at other scales or levels of organization, and altered function can affect structure at other scales or levels. Stress and disturbance are related, but they must not be confounded. Landscape context Putting disturbed sites in their landscape context indicates that disturbed patches are not isolated systems. They have the potential to interact with other patches, both neighboring and distant ones, in the landscape. In order to understand the potential for patches to interact in landscapes, their spatial relations must be described. Patches have size and shape, and these features along with what patches they abut, how far they are from similar patches, and the diversity of patch types in the landscape, determine the function of patches (Saunders et al., 1991). For example, patches of a particular composition may exchange dispersing organisms. The organisms of interest may not use patches that have some other structure. Therefore the status of that organism in the landscape depends on the spatial conﬁguration as well as the sizes of their preferred patches. Old-ﬁeld or grassland birds in primarily forested regions are a case in point (DeGraaf and Miller, 1997). Their regional demise may reﬂect a decrease in the number of disturbed patches and the distance between them. Likewise, movement of materials through patch mosaics can depend on the structure of the mosaic. How materials move through patches having vastly different resistances will determine the ﬂux of those

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS

materials through the landscape. This is generally referred to as connectivity. Connectivity may involve recognizable corridors deﬁned by patches of the same or similar type across a landscape. Or, more generally, it may be an abstract description of the ability of a patch mosaic to facilitate or retard ﬂow of a material from point to point in a landscape. When connectivity for one process and direction is achieved by the existence of a corridor, a ﬂow perpendicular to it, or the ﬂow of another material or organism, may be resisted. Hence a corridor in one direction may act as a barrier between patches arrayed in a different way in a mosaic. These phenomena illustrate the importance of accounting for the actual array of patches in a landscape in order to evaluate their function and relationship to one another. Three-dimensional patch bodies Patch interaction should not be thought of as though patches were two-dimensional features on maps. Although patch arrays are most often represented as simple, ﬂat maps, patches in real ecological landscapes and mosaics have three dimensional structures (Breen et al., 1988). In fact, the contrast between neighboring patches may often be as much a matter of architecture as of species composition. Forest plantations versus forest stands that arose from unmanaged seeding are two such contrasting patch types (Bradshaw, 1992). The three-dimensional structure of patches extends both above and below the substrate. Forest and ﬁeld patches may differ in the soil as well as in their vegetation canopies. The hyporhoeic zones of pools and rifﬂe patches in streams may differ as much as the water column. The links between the two systems through upwelling and downwelling of nutrients act to enhance the stability of the larger system comprising both zones (Valett et al., 1994). The movement and retention of organisms, energy and momentum, water and water vapor, and nutrients and pollutants may be conditioned by the three dimensional structure of disturbed patches among their undisturbed neighbors. The boundaries between patches are areas of special concern for the interaction of ecosystems of disturbed ground with neighboring patches. Apparently assuming that homogeneous areas promote the spread of disturbances, Forman (1987) stated that increasing heterogeneity in a landscape would decrease the permeability of the mosaic to disturbance. Fire is perhaps the best example of a disturbance agent that requires a homogeneously well-fueled matrix. Disease

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agents or pests with low contagion are another example. However, whether a heterogeneous or a homogeneous landscape will better promote the spread of disturbance depends on the particulars of the agent, its mode of spread, and the interaction between those features and the landscape structure. The interfaces between neighboring patches are most often referred to in the literature as “edges”. The connotation of this term is of a contact representable by a simple line on a map. A different reality is emerging from empirical studies, however (Gosz, 1993). Edges, like the patch bodies they bound, are complex, threedimensional zones that can extend well into each of the neighboring patches (Cadenasso et al., 1997). Different factors of the physical environment show different spatial patterns across edges (Chen et al., 1995; Cadenasso et al., 1997). The physical structure and biotic composition of the ecosystem also exhibit gradients or step functions across edges (Geiger, 1965; Ranney et al., 1981; Williams-Linera, 1990). Together, the abiotic and biotic gradients and structures can act as either ﬁlters or pumps, or be neutral to ﬂows across an edge (Pickett and Cadenasso, 1995). Much work remains to be done on this topic, but it is clear that the nature and structure of the edge between disturbed ground and its neighboring patches affects the behavior of those sites (Draaijers et al., 1994). Combining the various insights about structural heterogeneity outlined above, one may conclude that descriptions of ecosystems and communities of disturbed ground can no longer be silent about the structure of the patches or the nature of the landscape array in which disturbed ground appears.

PATCH DYNAMICS AND DISTURBANCE: CHANGE IN DISTURBED-GROUND ECOSYSTEMS

All the patterns and processes discussed above show the richness of inﬂuences that affect the structure in systems of disturbed ground. The value and utility of the knowledge of disturbed-ground ecosystems can be further increased by bringing it together in a single perspective that examines changes within disturbed ground ecosystems and the larger matrix in which they occur (Veblen, 1992). This section summarizes the current state of that perspective – patch dynamics. There are two major causes of change affecting ecosystems of disturbed ground. First is the creation of new disturbed patches, and second is the alteration of

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existing patches. These two processes must be considered together to understand the patterning and function of disturbed-ground ecosystems. It is important to recognize that some of the other patches with which a disturbed patch may interact have arisen or been altered by processes other than disturbance (Frelich et al., 1993). Some patches may be relatively permanent, established by resource hot spots and geomorphic features, while others may be more ephemeral and dynamic. The ﬂuxes from these different kinds of patches may differ, and for some may be relatively permanent, while in others the ﬂuxes may change through time as a result of patch change. In any event, the entire landscape structure must be understood in order to determine how systems of disturbed ground function. The landscape context suggests that in an array of patches, disturbance can be an important agent for creating new patches and altering the array. Because patches themselves can be dynamic and disturbance can act at various locations in a mosaic, the entire array may change (Remmert, 1991). This is the concept of patch dynamics, which refers to the changes in a spatial mosaic of patches over time, regardless of the scale of the mosaic, or the origin of the changes. This section will focus on disturbance as a major cause of patch dynamics in spatial mosaics, and the reciprocal effects between disturbance and patch mosaics. The temporal and spatial pattern of the creation of new patches by disturbance is the disturbance regime. A disturbance regime describes the various kinds of disturbance agents, the size, shape, and dispersion of disturbed patches, and the frequency with which the different kinds of patches are created. More detailed characterizations of disturbance regimes can include the internal heterogeneity of patches, and their relationship to particular geographic or landscape positions. In general, disturbance regimes and patch dynamics are spatially explicit concepts associated with the landscape viewpoint of ecology. Disturbance regimes are the spatial and temporal patterns of disturbance in speciﬁc areas. They are a key determinant of patch dynamics. Ecologists can use disturbance regimes to compare landscape dynamics across environmental gradients, or between contrasting sites. One kind of persistent spatial driver of differences in disturbance regime is the existence of local gradients of soil depth. In small desert catchments, patches disturbed by animals or better supplied with water are more common than in shallower soils elsewhere

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(Boeken et al., 1995). An example of the kind of comparison that becomes possible through being able to describe contrasting disturbance regimes is the different canopy structure, root architecture, and hence ecosystem functions in tropical forests exposed to different intensities and frequencies of hurricanes (Lugo and Scatena, 1996). Although few disturbance regimes have been completely characterized, especially their large, infrequent components, the recognition that a full understanding of any ecological system must include assessment of its disturbance regime is a signiﬁcant advance (Denslow, 1980; Tilman, 1996). At least some sizes and intensities of disturbance occur in virtually all ecological communities and ecosystems, and there remains the need to determine how the appearance of disturbed-ground ecosystems relates to physical or biotic patterns at a coarse to medium scale. The complexity of patch dynamics One of the most important open questions about the landscape conﬁguration of disturbed patches is whether the presence and conﬁguration of disturbed ground enhances the spread of disturbance in the landscape. Cases where disturbed ground acts as an inoculum for further disturbance in an area are the propagation of insect and disease outbreaks (Nothnagle and Schultz, 1987; Knight, 1987), the spread of ﬁre (Turner et al., 1989), and the increasing fetch around gaps that promotes wind-throw (Li et al., 1993). Sites that have been disturbed can sometimes act as ﬁrebreaks or windbreaks, however, depending on the structure of the surrounding landscape. The distribution of disturbed-ground ecosystems is determined by the intersection of different disturbance probabilities across geographic space. The creation of patches by disturbance is intimately linked to ecological heterogeneity. At the coarser level, patch creation depends on the heterogeneity of the matrix in which disturbance acts. For example, in a large landscape, the intensity and frequency of ﬁre depend on slope position (Hadley, 1994; Turner and Romme, 1994). More generally, speciﬁc agents that can act to disturb different ecological systems are conditioned by gradients and geographic location (Harmon et al., 1983). Earthquakes in fault zones can act as important agents of disturbance both to low-elevation estuary systems, for example the coastal Valdivian rainforests of Chile (Veblen, 1985), and in steep areas with high rainfall or where human or natural defoliation

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS

or deforestation is severe. This last situation can occur with logging or with tropical hurricanes in mountainous regions (Garwood et al., 1979; Walker et al., 1996). Differential sensitivity to disturbance in space causes certain types of disturbed ground to have characteristic geographic distributions, as illustrated by various chapters in this volume. Internal patch change Patches experience internal change. The most obvious internal change in patches in ecological landscapes is community succession. The arrival of different species in a patch, the growth and interaction of organisms in the patch, the alteration of resource pools and nutrient processing, and the variety of direct and indirect feedbacks among these organismal, community, and ecosystem processes change patches dramatically over time (MacMahon, 1981). It is beyond the scope of this chapter to review the richness of the literature on patch succession (Miles, 1979; Glenn-Lewin et al., 1992). However, it is one of the most powerful agents of change in disturbed patches, and hence in the larger mosaic in which disturbed patches are a part. The successional status of a patch, and of its neighbors, can determine its susceptibility to subsequent disturbance. Human management also causes patch change. In fact, management in most patches is essentially equivalent to altering the rate or trajectory of succession in patches (Luken, 1990). Hence, most human management is a manipulation of the disturbance status and subsequent successional status of ground disturbed at some point (Pickett and Rogers, 1997). Often the disturbance associated with management will be frequent and intense. But even in less intensively managed landscapes, disturbance will often be the tool that is brought to bear, or which is prevented or altered in order to produce systems with particular characteristics (Rogers, 1997). Finally, patch change now has a large human dimension in many landscapes (McDonnell and Pickett, 1993). Because human tenure and occupancy has spread so widely, the social and economic cycles that humans generate are now important engines of patch change (Turner et al., 1990; Cronon, 1991). On the global scale, the local intensiﬁcation and spread of agricultural and extractive land uses, the spread of urban–industrial inﬂuence, and indeed the increasing spread of urban areas, including suburbs, second homes, resorts, and the like, are all generating disturbed

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ground over vast regions. These patches change as management tactics and intensity change, as land uses change, and as more or different people come to occupy them. These three major classes of patch change – succession, management, and human settlement – together generate complex and dynamic mosaics of patches experiencing or recovering from disturbance, or merely moving from one type or intensity of disturbance to another. Disturbed ground is now in close proximity with less or rarely disturbed sites, and the interaction is one of the principal features of global environmental change. The dynamics within patches are ubiquitous and rapidly evolving. Changes in disturbance regime Disturbance regimes are not necessarily ﬁxed through time. Like climatic regimes, they are multivariate and subject to gradual or abrupt change through time (Clark, 1988). In fact, shifts in climate likely account for most natural changes in disturbance regimes. For example, there are apparently long-term cycles in the El Ni˜no–Southern Oscillation that alter drought and ﬂooding intensities and intervals over much of the globe (e.g., Leighton, 1986). These climatic cycles are associated with disturbances as disparate as ﬁres, ﬂoods, landslides, and windstorms. Migrations of species (Brown and Heske, 1990; Johnston, 1995), especially animals capable of “ecological engineering” (Jones et al., 1994), or geomorphic change are other potential causes of change in disturbance regimes. Disturbance regimes are also commonly altered by humans. Humans often interfere with natural disturbance regimes by inserting new kinds of disturbance – say, ﬁre in systems formerly predominantly disturbed by wind. Alternatively, people and institutions can alter the frequency of disturbance agents that remain active, as in the case of changing ﬂooding frequencies and intensities (Sparks, 1996). One of the most widespread modern alterations of disturbance regimes is the increase of energy subsidies that results from the replacement of non-industrial or locally subsidized human societies by industrial societies that draw on global markets and fossil-fuel subsidies. A complete understanding of disturbance patterns in landscapes requires knowledge of the role of various kinds of human subsidy in modifying or maintaining certain disturbance regimes. In spite of the transient nature of disturbance regimes on longer time scales, it is

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valuable to recognize that certain kinds, intensities, and distributions of disturbance events may, on a shorter time scale, be relatively predictable as key features of many ecological situations. Some implications of patch dynamics for disturbed-ground ecosystems Disturbed ground can be part and parcel of the functioning of persistent ecological landscapes. The stability of the system is the dynamic stability of a shifting mosaic of patches (Bormann and Likens, 1979). For some period of time while the climate and other forcing functions, including anthropogenic ones, remain relatively constant, the mosaic can act as a shifting steady state. This suggests that the interaction among disturbed patches, and the descendant patches of various successional states that replace them, is a key component in understanding how disturbance ﬁts into ecological systems, and how those systems function. Focus entirely on individual patches of disturbed ground can give a misleading picture of their role in larger ecological processes or landscapes. A hierarchical approach is required to evaluate the role of patch dynamics and metastability in system persistence. “Metastability” refers to the situation in which a system at a certain scale is stable whereas component parts of the system in fact change (Hanski, 1995). Many systems have metastability as a key feature of their persistence as a result of the widespread occurrence of disturbance regime and the patch dynamics they engender (Turner et al., 1993). Systems do not persist in the face of change by simply always resisting disturbance at a given level. Rather, many systems respond to the creation of disturbed patches through altered ﬂuxes of materials, energy, and species (Likens, 1984). Indeed, many species take advantage of the disturbance per se or the changes that follow disturbance and the initial response of certain organisms and ecosystem components to it. A system that is hierarchically divided into dynamic patches is not necessarily stable. One may witness the metapopulations that are in a transient state, with satellite populations being sinks rather than equal partners in periodically supplying colonists to other subpopulations in the system (Pulliam, 1988). An example is the bay checkerspot butterﬂy (Euphydryas editha bayensis), which has a large core population supplying a series of satellites (Noon et al., 1997).

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However, the question for any subdivided, hierarchically patchy system must be asked: is the system metastable? A ﬁnal value of considering disturbed ground in terms of disturbance regimes is a practical one. Because so many systems now or in the past experienced natural and non-industrial anthropogenic disturbance, disturbance is one tool available for managing wild or traditional human-dominated ecological systems (Harrison and Fahrig, 1995). Disturbance may be manipulated in many ways, including mimicking the heterogeneity and conditions created by natural disturbances, and modifying the resistance of systems to disturbance (Turner and Dale, 1990; Tilman, 1996). At the scale of large landscapes, such manipulations can affect the movement of species and materials across edges, and the function of mosaics of patches. For example, Franklin and Forman (1987) evaluated the role of different kinds of forest harvesting regimes in structuring a forest landscape, and their sustainability for resource extraction and provision of ecological services. Thus, disturbed ground and its ecological characteristics can be harnessed for restoration, production, sustainable management, and a host of other speciﬁc management goals (Pickett et al., 1997).

CONCLUSIONS

Ecological understanding of systems of disturbed ground has been enhanced by many conceptual and empirical developments over roughly the last decade. Initial research into the ecology of disturbance focused on demonstrating that the process of disturbance was ecologically signiﬁcant, and on determining how widespread and frequent it was. Appropriately, the most obvious and intuitive examples, and the most highly impacted systems, were exploited in those studies. However, as knowledge has increased, and more systems have been examined, the understanding of disturbance has become more subtle and better placed in context. Disturbance is now recognized to be a concept of much broader applicability than when it was ﬁrst introduced. Rather than being restricted to a particular spatial scale or criterion of ecological observation, it can be generalized to large and small systems, and can apply to the individual, population, ecosystem, and landscape. Such generality requires recognizing the core deﬁnition of disturbance as being an event that

PATCH DYNAMICS AND THE ECOLOGY OF DISTURBED GROUND: A FRAMEWORK FOR SYNTHESIS

interacts with the structure of an ecological system to change directly the structure of the system. In order to apply this more general deﬁnition to speciﬁc situations, a model of the system is required that speciﬁes the components, interactions, and temporal and spatial scales used to describe the system. Disturbances can be arrayed along gradients of intensity, and can have both discrete and diffuse effects at a particular scale. The gradients of intensity interact with gradients of structure in ecological systems to produce heterogeneity within disturbed sites. Heterogeneity produced by disturbance is a major cause of the variety and nature of the postdisturbance responses of ecological systems, and is proving to be a ubiquitous feature of disturbed-ground ecosystems. This realization contrasts greatly with the simplifying assumptions of early disturbance studies. Placing disturbances and disturbed sites in an explicit landscape perspective recognizes both the role of heterogeneity within disturbed sites, and the coarserscale heterogeneity which inﬂuences sensitivity and response to disturbance. The contemporary view of patch dynamics summarizes this richness of spatial effects on and of disturbance. As longer time series have accumulated in the study of disturbance, the variability of disturbance regimes with climate and human-accelerated environmental change has become apparent. Not only is this knowledge important in understanding the ecology of natural systems, but it is also crucial for effective management and conservation of natural and anthropogenic systems. In fact, the distinction between natural and anthropogenic systems is hard to maintain, so that management and restoration now commonly require disturbance regimes to be understood, their effects replaced, mimicked, or compensated for. Taken together, the insights summarized in this chapter show that disturbed ground is an important, vital, and dynamic part of ecological landscapes at all spatial scales.

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S.T.A. PICKETT, Jianguo WU and M.L. CADENASSO Walker, L.R., 1991. Tree damage and recovery from Hurricane Hugo in Luquillo Experimental Forest, Puerto Rico. Biotropica, 23: 379–385. Walker, L.R., Brokaw, N.V.L., Lodge, D.J. and Waide, R.B. (Editors), 1991. Ecosystem, plant, and animal responses to hurricanes in the Carribean. Biotropica, 23: 313–521. Walker, L.R., Zarin, D.J., Fetcher, N., Myster, R.W. and Johnson, A.H., 1996. Ecosystem development and plant succession on landslides in the Caribbean. Biotropica, 28: 566–576. Watt, A.S., 1947. Pattern and process in the plant community. J. Ecol., 35: 1–22. Weaver, J.E. and Albertson, F.W., 1943. Resurvey of grasses, forbs, and underground plant parts at the end of the great drought. Ecol. Monogr., 13: 63–117. Williams-Linera, G., 1990. Vegetation structure and environmental conditions of forest edges in Panama. J. Ecol., 78: 356–373. Wu, J. and Levin, S.A., 1994. A spatial patch dynamic modeling approach to pattern and process in an annual grassland. Ecol. Monogr., 64: 447–464. Wu, J. and Loucks, O.L., 1995. From balance of nature to hierarchical patch dynamics: a paradigm shift in ecology. Q. Rev. Biol., 70: 439–466.

Chapter 32

ECONOMIC GROWTH, HUMAN DISTURBANCE TO ECOLOGICAL SYSTEMS, AND SUSTAINABILITY Mario GIAMPIETRO

Editor’s note This volume deals primarily with how both natural and anthropogenic disturbances alter ecosystems, and with the responses following removal of the disturbance. The disturbance-altered ecosystems range from relatively pristine polar regions to heavily-impacted urban ecosystems. The disturbances themselves range in severity from lava ﬂows and pavement to tree-falls and herbivory. The present chapter takes a different approach. It addresses the consequences of human population growth on the entire planet. Many fear that a human population of more than 10×109 – likely during the next century – could imply that, in the future, re-establishment of natural ecosystems on the planet would be virtually excluded. The gross types of permanent disturbance which result are, in a sense, extreme cases of the milder types of disturbance, instantaneous or shorter in duration, that are covered elsewhere in this volume. When does environmental degradation by human activities reach a point where restoration of natural ecosystems is no longer practicable? What are the implications of the replacement of natural ecosystems with bricks and mortar, agriculture and agroforestry on net primary productivity and world food supplies? What options do world governments have to confront such global losses of resources? This chapter provides a bridge from the more speciﬁcally biological concerns raised in previous chapters to the social, ethical, and policy issues that humans must address as they continue to alter radically their natural environment. Lawrence R. Walker INTRODUCTION

In recent decades, the dramatic technological progress of humankind has translated into a rapid population and economic growth worldwide, that in turn has raised concerns for consequent environmental degradation (Brown, 1988–1996). Of course, the concern about the unavoidable friction generated by the continuously soaring demand for natural resources occurring in a ﬁnite planet is not at all new. Malthus (1803) posed in practical terms (population/natural resources) the theoretical problem of the innate instability of selfreplicating systems. Leopold (1949) made the point that, even in a situation of relative abundance of natural resources per head, as in the United States in the ﬁrst half of this century, there is an ethical and “value-dependent” view of the environment which does not coincide with an economic view of it. The

possibility of long-term development, in fact, requires the introduction of a culturally-driven set of rules curtailing the expansion of human activity. Meadows et al. (1972) gained world attention with a controversial attempt to simulate biophysical constraints on world economic growth. More recently, Vitousek et al. (1986) estimated that humans already were appropriating (directly or indirectly) 40% of the potential terrestrial net primary productivity. Given that terrestrial ecosystems provide roughly 99% of the world’s food supply (Food and Agriculture Organization of the United Nations, 1996), two main questions arise. First, will it be possible to produce sufﬁcient food – or, in technical terms, appropriate a sufﬁcient amount of terrestrial net primary productivity – to feed a world population of a size between 8 and 12×109 people? Second, what will be the nature and extent of environmental impact after

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a further enlargement of the fraction of net primary productivity appropriated by humans in the biosphere? When one adds to the problem of food security the problem of poverty in many areas of the world, the scale of the challenge becomes more severe. Given the new minimum acceptable standard of living that developed societies are spreading around the world the question becomes: is it possible to have a sustainable model of economic development, able not only to feed between 8–12×109 people in an ecologically compatible way but also to guarantee world-wide an “acceptable” and “fair” material standard of living? The two words “acceptable” and “fair” clearly indicate a cultural dimension which is crucial in deﬁning the terms of the problem. In spite of such a crucial role, the cultural dimension is often ignored by many “hard scientists” used to working only on the technical aspect of sustainability (i.e., matching requirement with supply). It is surprising to observe that when dealing with such a crucial issue, the scientiﬁc community has not provided a unanimously accepted evaluation of the sustainability of the current path of development of humankind. In fact, scenarios about the future of present civilization tend to be polarized into two main schools: (1) optimists, with very little concern about possible risks (the so-called “cornucopians” or “technological optimists”); and (2) pessimists, who foresee in the next century an unavoidable collapse of current models of technological development (the so-called “neo-Malthusians” or “prophets of doom”). There are scientists who see no problems in the future of world food security (e.g., Council for Agricultural Science and Technology, 1994; Alexandratos, 1995). Among these optimists (mainly agronomists) there are those who see no problems in feeding a population of 10×109 or even 12×109 , sparing at the same time some land for preservation of biodiversity (Council for Agricultural Science and Technology, 1994). On the other hand, there are pessimists about our future food security (e.g., Ehrlich et al., 1993; Kendall and Pimentel, 1994). Some of the pessimists (mainly ecologists) argue that the size of human population that could be fed without relying on fossil energy and without inducing negative consequences on natural ecosystems is about 2×109 (Pimentel et al., 1994). There are two schools of thought about the role of the present population in future development:

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(1) The orthodox (Blanchet, 1991) or neo-Malthusian theory (Chesnais, 1992) which considers population growth to be a negative factor in economic development. This view is concisely exempliﬁed by a statement of P. Bukman: “A development policy without a population program is like mopping the ﬂoor with the water turned on” (World Development Forum, 1989). (2) The transition theory (Blanchet, 1991) or stagnation theory (Chesnais, 1992), which views population growth as a positive factor in economic development. “The ultimate resource is people – skilled, spirited, and hopeful people – who will exert their wills and imaginations for their own beneﬁt and so, inevitably, for the beneﬁt of us all” (Simon, 1998). Looking more closely at these contrasting indications it is possible to see that these controversies are more apparent than real. Experts coming from different ﬁelds simply say different things since they are looking at the world from different perspectives (they are looking at the same system, but using different space–time scales). About food security, optimists say that it is technically possible to produce food for 10×109 people, and in saying that they are right; the pessimists, on the other hand, make the point that such a production, under present technology, would be totally dependent on disappearing stocks of fossil energy, and would not be ecologically compatible in the long term. In making this point, they are right too. Referring to the role of population in enhancing or hampering the ability of a society to make a better use of natural resources, both schools are right. When population size is not related to the environment in which society is operating, it is impossible to assess whether an increase in population size is a positive or negative event. For example, in the United States between 1800 and 1900, doubling of the population deﬁnitely improved technological capital and the ability to make better use of the abundant natural resources available, resulting in a better standard of living. On the other hand, a doubling of China’s current population would undoubtedly result in a major ecological disaster and most probably in a collapse of its socioeconomic structure. Within this framework, biophysical analyses linking human societies to ecological processes tend to have a bias against population expansion. Used to looking at the world on a large space–time scale, they tend to be mainly worried about the stability of current boundary conditions. An increase in population is

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perceived mainly in terms of a reduction of natural resources per head. What they tend to overlook is the prodigious ability of humans to adapt quickly to changes imposed on them. On the contrary, economic analyses tend to have a bias in favor of population expansion. Used to looking at the world with a smaller space–time scale (5–10 yr; country level), they tend to be concerned mainly with increasing the speed at which society can improve its internal economic characteristics. An increase in the size of the economic process is perceived mainly as a better possibility of specialization and economies of scale. What economic analyses tend to overlook is that human technology is totally irrelevant when compared to biospheric processes (e.g., the total ﬂow of energy controlled by humans is about 12×1012 W, whereas the biosphere, just for cycling water, uses about 44 000×1012 W). There is nothing that human technology can do if natural processes stop providing a life-support system. Generalizing these ﬁndings, one can say that scientists specialized in a certain perspective are unavoidably biased by the window of observation (the space– time scale) chosen to look at the world (Giampietro, 1994a). The very concept of “disturbance” when extrapolated outside its original context (disruption of a deﬁned community structure; Pickett et al., Chapter 31, this volume) becomes quite elusive. Across scales, disturbance can assume a positive or negative value; for instance, the very same forest ﬁre can have the effect of de-stabilizing certain ecological processes at a small scale, whereas it plays a role in stabilizing other ecological processes on a larger scale (O’Neill et al., 1986). Complex cross-relations among scales – typical of ecosystems and socioeconomic systems, where everything depends on everything else but on different scales – are at the root of the failure of scientiﬁc reductionism to deal with the issue of sustainability. The consequent lack of unanimous scientiﬁc indications about what should be considered “good” or “bad” for the sustainable development of humankind is generating, in turn, an impasse in decision making about natural-resources management. Sound policies require a parallel check of the economic viability, ecological compatibility, social acceptability, and technical feasibility of the different options considered. Unfortunately, this would require a comparison of different and parallel descriptions and assessments related to processes operating on widely different space–time scales (Giampietro, 1994a). According to the reductionistic paradigm usually used for

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decision making (collapsing the complexity of the real world into a single indicator of cost/beneﬁt, possibly economic value) this translates into a “mission impossible syndrome”. Indications coming from economic cost/beneﬁt analysis (e.g., describing the interaction of humans and ecosystems on a 5-year time scale) are neither consistent nor comparable with indications coming from biophysical analyses attempting to deﬁne ecological cost/beneﬁt (e.g., describing the interaction of humans and ecosystems on a 500-year time scale) (Giampietro, 1994a,b). To make things worse, conﬂicting interests, insufﬁcient and contrasting scientiﬁc information, and above all a diffuse uncertainty about the relationship among parts and the reliability of future scenarios on which policy makers are called to make decisions, are more and more coupled to a dramatic increase in the urgency of decisions and relevance of stakes (Funtowicz and Ravetz, 1994a,b). In this framework, sustainability of human development should be discussed in terms of integration at a large scale of human and ecological disturbances, such as the long term coevolution of humans and the rest of the biosphere. Therefore, this chapter addresses two main issues: (1) the link between economic development and disturbance of ecological systems on our planet; and (2) the mechanism of controls operating within human societies that should make possible such a coevolution. This chapter is divided into two parts. The ﬁrst part explores the link between development – seen as a change in the socio-economic structure of human societies – and environmental stress. In particular, the issue of food security is used as an example to discuss the ecological implications of a general increase in demographic and socio-economic pressure worldwide (both types of pressure increase the degree of disturbance in human-dominated ecosystems). The second part discusses the mechanisms generating Jevons’s paradox – that increased efﬁciency in using a resource leads to increased use of it; acknowledging the existence of an innate drive toward unsustainability of human societies is the starting point from which to get rid of several myths hiding the real nature of the problems faced today. Humankind can successfully respond to the challenge of sustainability only if it is able to re-evaluate the current set of wrong “default” assumptions about technical development.

726 THE LINK BETWEEN ECONOMIC DEVELOPMENT AND ENVIRONMENTAL STRESS

The general framework of analysis Within every human society there is a “bio-economic pressure” – generated at the hierarchical level of individuals – which is aimed at improving the material standard of living in that particular society (Giampietro, 1997a). The term “bio-economic pressure” focuses on the fact that changes in socioeconomic variables – such as income, work load per worker, life expectancy, percentage of work force in the service sectors – have a direct, biophysical effect on throughputs of matter and energy in the economic process. In fact, economic development, described by changes in the above mentioned set of variables, results in a dramatic acceleration of throughputs of energy and matter within the economic process (producing and consuming more goods and services per head). The direct link between economic and biophysical variables is determined by the need of keeping in different economic sectors the congruence between the two proﬁles of: (1) demand and supply of energy; and (2) demand and supply of human time (Giampietro, 1997a; Giampietro et al., 1997). That is, the huge increase in the level of consumption of raw materials, including food, consumed by a developed society is coupled to a dramatic shrinking of the amount of working time allocated to the productive sectors of its economy. This can be achieved only through an acceleration of throughputs of matter and energy in the productive sectors (e.g., agriculture, energy and mining, manufacturing). This, in turn, implies that more inputs per head are taken from the ecosystem in which the society is embedded, and more wastes per head are dumped into it. The amount of natural resources required per head (used to get needed inputs or to absorb wastes) deﬁnes the level of environmental loading per head (the disturbance of local ecosystems). At the hierarchical level of ecosystems, any increase in environmental loading per head generates an opposite pressure, which tends to counter an indeﬁnite expansion of societal activity. In fact, a continuous growth of economic activity would imply the need to (1) extract resources at a higher rate from a shrinking endowment; and (2) dump increasing wastes into increasingly ﬁlled sinks.

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Society can adjust to these negative feedbacks from the environment by changing its technology. In this way, humans can make up for decreasing quality and quantity of certain resources, and can attempt to control pollution for those wastes dumped into saturated sinks. However, such a human response has a biophysical cost (more energy and human time has to be allocated to activities than were necessary before). A continuous growth of environmental loading can transform free resources such as clean air and water into “expensive” resources in biophysical terms (requiring energy and technological investments to make them available). The sustainability of economic development will eventually depend on the difference between two rates of change: (1) the rate at which technical innovations generate room for expansion of the scale of the economic process; (2) the actual rate of expansion of the scale of the socio-economic process which is determined by the product: population×activity per head. Put in another way, when the environmental loading generated by the activity of the socioeconomic system exceeds a critical level (when it is no longer compatible with the stability of the ecological processes) the society will clash against biophysical constraints. This critical threshold will depend on: (1) the degree of alteration of natural patterns of ﬂow of matter and energy induced by human activity in the exploited (disturbed) ecosystems; (2) the nature of the disturbed ecosystem; and (3) the scale of the economic process (ﬂows of input and output) compared to natural resources (the size of available stocks and sinks). Changes of socio-economic structure induced by development The process of self-organization of human society can be seen as the ability to stabilize in time and space a network of ﬂows of matter and energy representing what is called the economic process. In familiar terms, this translates into the ability to stabilize and continuously improve the process of production and consumption of goods and services. To be sustainable, such a process has to be: (1) compatible with the aspirations of the humans belonging to the society; (2) compatible with natural and human-managed ecosystems; and

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Fig. 32.1. Age structure of societies at different levels of socioeconomic development (Giampietro, 1997a); y axis: the range of age values in years; x axis: the number of individuals in each age class. Numerical values on hatch marks have been removed. Due to the large differences in size between the 4 societies, the graphs are based on different scales to better focus on the difference of proﬁles of distribution of population over age classes.

(3) technically feasible. Technological development of a society can be described in terms of a de-coupling between the proﬁle of allocation of human time and that of energy allocation in different sectors of the economic process (Giampietro, 1997a). In fact, technological development implies: (1) an increase in average consumption of energy per head – from less than 5 GJ yr−1 in poor developing countries to more than 300 GJ yr−1 in the United States, for instance; (2) a dramatic reduction in the ratio (human time allocated to work in the productive sectors of the economy)/(total human time) – from 0.1 in poor developing countries to 0.04 in developed countries (Giampietro et al., 1997). Such a reduction is due to an increase in the level of education, a progressive aging of the population, lighter work loads for the working force, and an

increasing fraction of the working force (up to 60%) that is absorbed by the service sectors of the economy. The combination of these changes in socio-economic characteristics implies that, in modern societies, a smaller and smaller fraction of total human time is used for running the productive sectors of the economy (e.g., food security, energy and mining, manufacturing), whereas the material throughput in these sectors has dramatically increased. To visualize the effect of development on socioeconomic variables through a practical example, one may consider the population structure of four different societies (Fig. 32.1): (1) a pre-industrial-type society in which the population is stabilized by high mortality and fertility rates (Yanomam¨o, in the Brazilian Amazon);

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Fig. 32.2. Allocation of human time in societies at different levels of socioeconomic development (Giampietro, 1997a). Working in productive sectors = human time used in the procurement of food, energy and other raw materials, in mining and manufacturing; working in services = human time used to perform the ensemble of activities typical of the service sector of modern economies; non working = sleeping and leisure time of the working population, and the entire time of population not economically active (e.g., children and retired).

(2) a developing society with a fast growing population (Burundi); (3) a developed society with a population still growing but at a decreasing pace (U.S.A.); and (4) a developed society in which the population is stabilized at low mortality and fertility rates (Sweden). The ratio between the economically active and nonactive populations (the ratio between the shaded and non-shaded area in the population pyramid) is clearly different for these societies. These ratios are closely related to the proﬁle of allocation of human time between “labor” and “non-labor” activities. In fact, the allocation of human time to labor and non-labor activities is affected by the fraction of the population that is economically active and the work load (number of labor hours per year). In turn, these two parameters are determined by several other factors with social signiﬁcance, such as distribution of population among age classes (affected by life span); minimum age that is socially acceptable for entering and leaving the labor force; and the fraction of the potential labor force that does not work for whatever reason (e.g., unemployment, education, illness). When one looks at the different patterns of human time allocation for the same four societies (Fig. 32.2), one ﬁnds that developed societies not only have a

smaller fraction of their human time allocated to work, but also have only a small fraction of their working time allocated to productive sectors of the economy. The level of metabolic energy (energy input – food – transformed into useful energy within the human body, often called endosomatic energy) is more or less similar in these societies. It ranges from 7 MJ day−1 per head in Yanomam¨o tribes to about 10 MJ day−1 per head in Sweden (Giampietro et al., 1997). However, large differences occur between these societies in the level of exosomatic energy (energy inputs transformed into useful energy outside the human body – for example, machine power) consumed per head, which ranges from about 35 MJ day−1 in Yanomam¨o tribes to about 860 MJ day−1 per head in the United States. The ratio “exosomatic/endosomatic energy” can be used as an indicator of how much the “societal metabolism” depends on machine power and technology rather than on human labor (Giampietro, 1997a; Giampietro et al., 1997). This ratio ranges from a value of 5/1 in preindustrial societies to a value higher than 90/1 in developed societies (Giampietro, 1997a; Giampietro et al., 1997). This rationale has been validated by studying the correlation between the acceleration of energy throughputs in socio-economic systems – assessed by the bio-economic pressure (BEP) – and 24 traditional

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indicators of physiological, economic, and social development, over a sample of 107 countries covering more than 90% of world population (Pastore et al., 1996). In this analysis, bio-economic pressure has been calculated as the total energy consumed by a society in a year divided by the total amount of working time used in the productive sectors of its economy in the same year. All 24 indicators of standard of living considered show a good correlation with bio-economic pressure (Table 32.1). A sample of the type of results obtained is shown in Fig. 32.3. In conclusion, technological and social development depends on a huge increase in labor productivity in the productive sectors of the socioeconomic process (energy sector, food production, mining, and procurement of other raw materials). In a rich society, the productive sectors of the economy must not only be able to generate an adequate supply of goods for the society, but also to do so while absorbing only a negligible fraction of labor time. What are the implications of this trend for ecological disturbance? Disturbance induced by food production Three simple observations make evident the crucial link between food security and disturbance of natural ecosystems worldwide: (1) more than 99% of food consumed by humans comes from terrestrial ecosystems (Food and Agriculture Organization of the United Nations, 1996); (2) more than 90% of this food is produced by using only 15 plant and 8 animal species (Wilson, 1988), while estimates of the existing number of species on the Earth are in the millions; (3) worldwide, arable land is already less than 0.27 ha per head, and this ﬁgure is expected to continue to shrink because of population growth (Pimentel et al., 1995). In addition, arable land is being lost. During the past 40 years nearly one-third of the world’s cropland (1.5×109 ha) has been abandoned because of soil erosion and other types of degradation (World Resources Institute, 1992; Kendall and Pimentel, 1994; Pimentel et al., 1995). Most of the land (about 60%) added to replace this loss has come from marginal land made available mainly by deforestation (Pimentel et al., 1992). High productivity per hectare on marginal lands requires large amounts of inputs based on fossil energy. This occurs at the very same time that the

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Table 32.1 Correlation between bio-economic pressure (BEP) 1 and some major indicators of development 2 Indicators of development

r3

Economic indicators Log (Gross National Product)

+0.89

% of GNP from agriculture

−0.85

Added value per hour of paid labor ($US)

+0.90

% of work force in agriculture

−0.93

% of work force in services

+0.88

Log (energy consumption per head)

+0.98

% of income spent on food

−0.89

Physiological indicators Life expectancy (yr)

+0.88

Energy intake (in the diet:kJ per head per day)

+0.83

Fat intake (in the diet: g per head per day)

+0.80

Protein intake (in the diet: g per head per day)

+0.79

Malnutrition of children (% of children under 5 under −0.83 weight) Infant mortality (per thousand live births)

−0.86

Low birth weight (% of live births)

−0.62

Social indicators Log (television sets per inhabitant)

+0.90

Log (cars per inhabitant)

+0.90

Log (newspapers per inhabitant)

+0.89

Log (telephones per inhabitant)

+0.89

Log (population per physician)

−0.87

Log (population per hospital bed)

−0.76

Pupils per teacher

−0.74

Illiteracy rate (% of population aged above 15)

−0.67

Primary school enrollment (% of school-age population)

+0.58

Access to safe water (% of population)

+0.81

1

Bio-economic pressure: total energy consumed by a society in a year/total amount of working time used in the productive sectors of its economy in the same year. 2 After Pastore et al. (1996). 3 The y variable in each case is log (BEP).

economic growth of many developing countries is dramatically increasing the demand for alternative uses (e.g., infrastructure construction and household consumption) on the shrinking stocks of oil. The degree of disturbance of ecosystems for food security is not only linked to increases in demographic pressure, but also linked to the level of economic de-

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Mario GIAMPIETRO

Fig. 32.3. Examples of correlation between Bio-Economic Pressure (BEP) and indicators of socioeconomic development (after Giampietro et al., 1997). The sample includes 107 countries comprising more than 90% of world population (all countries with a population >5 million, excluding those of the former-Soviet Union). Data refer to 1991. GNP, Gross National Product expressed in $US; energy consumption in kg coal equivalents per head yr−1 .

velopment of the society within which food production is taking place. In fact, at a given level of demographic pressure the need to produce food at a high level of productivity of labor implies more disturbance to the managed ecosystems. That is, to get levels of labor productivity of the order of hundreds of kilograms of grain per hour, farmers must heavily rely on technical inputs, switching from cyclic to linear nutrient ﬂows, from multicropping to monocultures, from traditional varieties to selected seeds (Giampietro, 1997b,c). Put in another way, the pattern of production and consumption of food in human societies is heavily affected by both the characteristics of the managed ecosystems and the characteristics of society (Giampietro et al., 1994; Giampietro, 1997b,c). The two main factors affecting techniques of food production discussed below are: (1) population density or demographic pressure; high demographic pressure implies that only a small amount of arable land is available per head.

This in turn translates into the tendency to adopt techniques of production with a high productivity per unit of land area; (2) level of technological development of society or socioeconomic pressure; a high socioeconomic pressure implies that only a small fraction of work supply is available for food production. This in turn translates into the tendency to adopt techniques of production with a high productivity of labor. Demographic pressure can be quantiﬁed in terms of arable land available per head, whereas socioeconomic pressure can be conveniently quantiﬁed by either of the following indicators: (1) average income per head; (2) average consumption of commercial energy per head (Giampietro, 1997b). An increase in socioeconomic pressure results in a dramatic reduction of the work force allocated to agriculture. A cross-section analysis over a signiﬁcant sample of countries (Fig. 32.4) shows that the fraction

ECONOMIC GROWTH, HUMAN DISTURBANCE TO ECOLOGICAL SYSTEMS, AND SUSTAINABILITY

Fig. 32.4. Relation between proportion of working force in agriculture and proportion of Gross Domestic Product (GDP) from agriculture on the one hand, and GDP ($US per head yr−1 ) on the other. Data of 1991. The 60 countries included are: Algeria, Argentina, Australia, Bangladesh, Brazil, Burkina Faso, Burundi, Cambodia, Cameroon, Canada, Central African Republic, Chad, China, Colombia, Congo, Costa Rica, Ecuador, Egypt, El Salvador, Ethiopia, EU (average for the 12 countries of European Union), Finland, Gambia, Guatemala, Honduras, India, Indonesia, Iran, Jamaica, Japan, Jordan, Kenya, South Korea, Madagascar, Malawi, Mali, Mauritania, Mexico, Morocco, Mozambique, New Zealand, Nicaragua, Niger, Nigeria, Pakistan, Paraguay, Philippines, Senegal, Sri Lanka, Sweden, Switzerland, Tanzania, Thailand, Tunisia, Turkey, U.S.A., Uganda, Uruguay, Venezuela and Zimbabwe. From Giampietro (1997b).

of the labor force engaged in agriculture is inversely correlated with the Gross Domestic Product (GDP) per head. Countries with a GDP per head of more than $US 10 000 yr−1 have less than 7% of the working force in agriculture, and the richest countries have as few as 2%. Two types of constraints affect technological choices of societies undergoing rapid economic development: (1) the acceleration of throughputs in the productive sectors of the economy (increase in socioeconomic pressure) demands higher labor productivity; and (2) fast population growth (increase in demographic pressure) is likely also to demand higher productivity of land (Giampietro, 1997b; Giampietro et al., 1997). The combined effect of these two pressures in turn affects the level of disturbance as a result of agricultural production (e.g., reduction of biodiversity, soil erosion, pollution, depletion of underground water reservoirs, higher biophysical cost of food in terms of

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energy input). Higher energy inputs mean an increasing dependence of food security on disappearing stocks of fossil energy (Giampietro, 1997c; Conforti and Giampietro, 1997). To discuss the effect of demographic and socioeconomic pressure on the choice of agricultural techniques it is useful to compare: (1) levels of land productivity required for selfsufﬁciency in relation to current availability of arable land (APDP ) versus actual levels of land productivity (APha ) as shown in Fig. 32.5a; and (2) levels of labor productivity required for selfsufﬁciency in relation to current work supply in agriculture (APSEP ) versus current levels of labor productivity (APhour ) as shown in Fig. 32.5b. From this comparison it is possible to see the existence of a direct link between (1) characteristics of a socioeconomic system – determining the values of APDP and APSEP ; and (2) characteristics of techniques of agricultural production adopted in a deﬁned society – determining the values of APha and APhour (Giampietro, 1997b). According to these ﬁndings one can expect that general economic development of the planet will determine a trend of a continuous increase in both productivity of labor and land in world agriculture. What are the implications of this trend in terms of disturbance of ecological systems? Agriculture can be deﬁned as a human activity that exploits natural processes and natural resources in order to obtain food and other products considered useful by society (e.g., ﬁbers and stimulants). The verb “exploits” suggests that one is dealing with an alteration of natural patterns, which is disturbance. Indeed, within a deﬁned area, humans alter the natural distribution of both animal and plant populations in order selectively to increase (or reduce) the density of certain ﬂows of biomass that they consider more (or less) useful for the socioeconomic system. Following the schema of ecosystem structure proposed by Odum (1983), a natural ecosystem can be seen as a network of matter and energy ﬂows in which nutrients are mainly recycled within the system and solar energy is used to sustain this cycling (Fig. 32.6a). The amount of solar energy used for self-organization by the ecosystem is proportional to the speed of its cycles of material (in terrestrial ecosystems this also has to do with availability and circulation of water). Agriculture then involves a boosting of only those ﬂows of matter and energy in the network that humans consider beneﬁcial, and eliminating or reducing the ﬂows

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Fig. 32.5. The effect of demographic and socioeconomic pressure on agricultural productivity. Data of 1991. The countries included are the same as for Fig. 32.4. (A) Demographic pressure versus actual land productivity; (B) socioeconomic pressure versus actual labor productivity. From Giampietro (1997b).

that they consider detrimental to their purposes. In doing so, humans amplify some of the existing genetic information (that contained in populations representing “useful” biomass) and reduce or eliminate some other information (that contained in populations representing “unwanted” biomass). Therefore, agriculture has to do with the direct manipulation of biodiversity. This implies, as discussed before, that a massive spread of agricultural production into the remaining natural ecosystems of the world, will induce a further dramatic reduction of biodiversity at the global level. Depending on the amount of harvested biomass such

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Fig. 32.6. Ecosystem representation in terms of matter and energy ﬂows following Odum (1983). The arrows linking different compartments represent matter and energy ﬂows. Different shapes of compartments indicate different roles: Q1 are autotrophs (e.g., plants ﬁxing solar energy); Q2, Q3, Q4 and Q5 are heterotrophs (such as herbivores and carnivores depending on primary productivity of Q1; Q6 are detritus feeders, recycling nutrients within the ecosystem. (A) Natural ecosystem as a balanced network of ﬂows in which nutrients are mainly recycled by using solar energy; (B) agroecosystem subject to intense exploitation for crop production; nutrients ﬂows are mainly linear with severe leaks outward. From Giampietro (1997c).

a process of alteration can have serious consequences for an ecosystem’s structure (Fig. 32.6b). For example, in order to increase signiﬁcantly the yield of biomass (amount taken away from the agroecosystem per unit of area) it is necessary to ﬁght competing species (pests) and to supplement physiological rates of cycling of nutrients and water in the ecosystem (by fertilization and irrigation) when these become limiting. This calls for (1) selection of crop species and varieties that perform well under human-managed conditions (erosion of crop diversity within cultivated species); and (2) adoption of monocultures in order to synchronize the operations on the ﬁeld (substitution of

ECONOMIC GROWTH, HUMAN DISTURBANCE TO ECOLOGICAL SYSTEMS, AND SUSTAINABILITY

machine power for human power). When agricultural production (per hectare and per hour) becomes very intensive, production techniques are forced to ignore the functional mechanism of natural ecosystems – that is, the cycling of nutrients powered by solar energy (Giampietro, 1997c). Traditional low-input agriculture, in which humans cooperate with the natural system of control of matter and energy ﬂows, is based on cycles of nutrients and interactions among several species in the agroecosystems. This solution has the advantage of a low requirement of human inputs but the disadvantage of a low density of throughputs. Whenever demographic and/or socioeconomic pressure force very high levels of production (e.g., several thousands of kilograms of grain per hectare and/or hundreds of kilograms of grain per hour of labor), such a form of agriculture becomes no longer viable. The reverse is true for highinput agriculture, which, at the moment, is the only available option to reach very high levels of productivity both per hour and per hectare. A high-density agricultural throughput means a drastic reduction of biodiversity in the agroecosystems (related to the use of monocultures), high environmental loadings (related to the heavy use of technical inputs), and an increasing dependence of food security on depletion of stocks (the mining of fossil energy).

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of arable land and more than $US 10 000 yr−1 of GNP per head. (3) Socioeconomic systems, such as China and Egypt, with high demographic and low socioeconomic pressure – that is, less than 0.2 ha of arable land and less than $US 1000 yr−1 of GNP per head. (4) Socioeconomic systems, such as several West European countries and Japan, with high demographic and high socioeconomic pressure – that is, less than 0.2 ha of arable land and more than $US 10 000 yr−1 of GNP per head. According to existing trends in population growth and economic development for these four different types of socioeconomic systems, one may expect the following movements in the plane (see Fig. 32.7):

Future scenarios for agricultural development General trends in the evolution of techniques of food production for different types of socioeconomic systems can be represented on the two-dimensional plane: productivity of land (kg ha−1 ) and productivity of labor (kg hr−1 ) as illustrated in Fig. 32.7. Four main types of socioeconomic systems, having different combinations of demographic and socioeconomic pressure, are represented there: (1) Socioeconomic systems with low demographic and low socioeconomic pressure. This situation is characterized by more than 0.5 ha of arable land per head (depending on population size) and less than $US 1000 yr−1 of GNP per head (depending on economic performance). This type of socioeconomic system occurs in several African countries, such as Burundi. (2) Socioeconomic systems, such as the United States and Canada, with low demographic and high socioeconomic pressure – that is, more than 0.5 ha

Fig. 32.7. Current position and expected trends for labor and land productivity in the agricultural sector of different groups of countries. For the sake of simplicity the axes represent productivity of land and labor in terms of kg of grain. EU, average value for the 12 countries of the European Union. From Giampietro (1997c).

(1) Societies with low demographic and socioeconomic pressure (e.g., some African countries): the population is growing faster than the GNP per head, which means that APDP will grow faster than APSEP . Hence, they will move toward a situation typical of China. (2) Societies with low demographic and high socioeconomic pressure (e.g., Canada, U.S.A.): economic development is expected to be maintained (GNP

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per head will remain high) and population growth will be relatively slow but steady (low internal fertility and high immigration rate). On the plane, this means a slow movement toward higher values of APDP . (3) Societies with high demographic and low socioeconomic pressure (e.g., China) tend to have good economic growth (increasing GNP per head) and maintain if not expand their already huge population size. At a national level, an increasing GNP per head will result in an accelerated absorption of the labor force currently engaged in agriculture (60% in China at present) by other sectors (production and service sectors) of the economy. This will inevitably require a dramatic increase in agricultural labor productivity (APhour ) to maintain food security. Hence, a movement toward the West European conditions of agricultural production is to be expected. (4) Societies with high demographic and socioeconomic pressure (e.g. The Netherlands, Japan): these societies have no alternative but to try to maintain a high material standard of living and keep population growth to a minimum. This means a more or less stable and high level of APSEP and a very slowly increasing value of APDP (mainly due to a strong immigration pressure). For these societies, trying to reduce the environmental impact of their food production becomes of major importance. One may note that food imports from the international market, a necessity for countries where APDP > APha and/or APSEP > APhour , is based on the existence of surpluses produced by countries where the relation between these parameters is inverse. At present, the United States, Canada, Australia, and Argentina combined produce over 80% of the net export of cereals on the world market (World Resources Institute, 1992). But, at their present rate of population growth (including immigration) and because of an increasing concern for the environment (policies for set-aside and development of low-input agriculture), this surplus may be eroded in the near future. For instance, the United States is expected to double its population in 60 years (United States Bureau of the Census, 1994). A general concern for the environment leading to less intensive agriculture all over the world (a slow-down, at the farming level, of the rate of increase in APha ) could also work against the production of food surpluses in those countries where it is possible. In

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spite of these trends, at the world level demographic and socioeconomic pressures are certainly expected to increase, forcing the countries most affected by these two pressures to rely on imports for their food security. It is often overlooked that, at the world level, there is no option to import food from elsewhere. When increases in demographic and socioeconomic pressure are not matched by an adequate increase in productivity of land and labor, food imports of the rich will be based on starvation of the poor. How do these trends ﬁt with overall biophysical constraints to food production? Boundary conditions and physical laws impose biophysical constraints on what is feasible for human societies in spite of their wishes and aspirations. Therefore, it is important to have a sense of what are the most likely biophysical constraints faced by humankind in the future in terms of food security. In addition to shortages of arable land, soil erosion, and a lack of alternatives to high input agriculture to guarantee elevated productivity both per hour and per hectare, there are other crucial constraints. Water shortages: Presently, 40% of the world’s people live in regions where there is competition for limited water supplies. Related to these growing shortages is the decline in availability of fresh water for food production and other purposes in the arid regions of the world (Postel et al., 1996). There is very little that technology can do when human development clashes against limits in the water cycle. As noted earlier, the amount of energy required to sustain the water cycle in the biosphere is almost 4000 times the entire amount of energy controlled by humankind worldwide. The idea that human technology will substitute for the services provided today by nature in this ﬁeld is simply based on ignorance of biophysical realities. Biodiversity for the long-term stability of the biosphere: The dramatic reduction of species caused by the conversion of natural ecosystems into agroecosystems has been discussed earlier. Since biodiversity is needed to stabilize the structure and functions of the biosphere, one cannot transform all terrestrial ecosystems into agricultural ﬁelds and/or cement for housing and infrastructures. A large diversity of species is vital to agriculture and forestry, and plays an essential role in maintaining a quality environment and in recycling the vital elements such as water, carbon, nitrogen, and phosphorus. Other well-known global problems can be added

ECONOMIC GROWTH, HUMAN DISTURBANCE TO ECOLOGICAL SYSTEMS, AND SUSTAINABILITY

to this list: changes in the composition of the atmosphere (greenhouse effect, ozone layer depletion), the cumulative effects of pollution, the intensiﬁcation of the occurrence of contagious diseases (e.g., AIDS). I mention them here not in order to enter the argument between “cornucopians” and “neo-malthusians” but only to make the point that, in spite of current disagreements about the evaluation of the seriousness of these problems, one has to acknowledge the obvious fact that there are biophysical limits to the expansion of human activity. These limits can be avoided only by adequately reacting to feedback signals coming from disturbed ecosystems. Whenever humans are not able to obtain reliable indications about the room left for expansion (or when they are not able to understand those signals) they should be reluctant to expand their disturbance further.

CONFRONTING THE INNATE DRIVE TOWARD UNSUSTAINABILITY OF HUMAN SOCIETIES

Jevons’s paradox and the myth of steady-state efﬁciency “Jevons’s paradox” (F. Jevons, 1990) was ﬁrst enunciated by William Stanley Jevons in his book The Coal Question (1865). Brieﬂy, it states that an increase in efﬁciency in using a resource leads, in the medium to long term, to an increased use of that resource rather than to a reduction in its use. At that time, Jevons was discussing the trend of future consumption of coal. At the same time, others were predicting a future reduction in consumption of fossil energy due to more efﬁcient engines generated by technological progress. Jevons’s paradox proved to be true not only with regard to demand for coal and other fossil-energy resources but also with regard to demand for food resources. Doubling the efﬁciency of food production per hectare over the last 50 years by a dramatic increase in “efﬁciency” in producing food (the Green Revolution) did not solve the problem of hunger, it actually made it worse, since it increased the number of people requiring food (Giampietro, 1994b). In the same way, doubling the area of roads did not solve the problem of trafﬁc – it made it worse, since it encouraged the use of personal vehicles (Newman, 1991). As more energy-efﬁcient automobiles were developed as a consequence of rising oil prices, American car owners increased their leisure driving (Cherfas, 1991).

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Not only the number of miles increased, but also the expected performance of cars grew; United States residents are increasingly driving mini-vans, pick-up trucks, and four-wheel-drive vehicles. More efﬁcient refrigerators have become bigger (Khazzoom, 1987). In economic terms one can describe these processes as increases in supply boosting the demand. Jevons’s paradox has also been called the “rebound effect” in energy literature and the “paradox of prevention” in relation to public health. In the latter case, the paradox consists of the fact that the amount of money “saved” by prevention of a few targeted diseases generates in the long term an increase in the overall bill of the health sector. Owing to the fact that humans sooner or later must die (which is a fact that seems to be ignored by steady-state efﬁciency analysts), any increase in life span of a population directly results in an increase in health-care expenses. It is well known that hospitalization of the elderly is much more expensive than hospitalization of adults in the working years. This last example leads to the heart of the paradox. Technological improvements in “efﬁciency” of a process (e.g., increases in distance traveled per liter of fuel) represent improvements in intensive variables. That is, they can be deﬁned as “improvement” under the ceteris paribus hypothesis that everything remains the same. Increases in efﬁciency translate into savings only when the system does not evolve in time (when a steadystate representation of the system is satisfactory for the purposes of decision-making). Unfortunately, complex systems, especially human systems, tend to adapt quite fast to changes. As soon as such “technological improvements” are introduced into a society, room is generated for a further expansion of its level of activity (e.g., more people make more use of their cars). Latter expansion represents a change in extensive variables – that is, in the dimension of the process. For these reasons, sound decision making should be based on a holistic assessment of the problem considered (including both a steady-state view and an evolutionary view of the process analyzed) and on an explicit deﬁnition of what is optimized. Going back to the paradox of prevention, policies aimed at reducing health costs (e.g., smoking restrictions to prevent lung cancer) have the effect, in the long term, of increasing the cost of health care. However, paradox in the paradox, this is a good result for society. In fact, when assessing health-care costs in an evolutionary

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perspective one can easily recognize that a larger bill for health care is an indicator of development. I will now examine other examples of the systematic error generated by the description of evolving systems as if they were in a steady state. This logical mistake has generated a series of myths that are relevant for the issue of sustainable development. The myth of dematerialization of developed economies A naive economic deﬁnition of energy efﬁciency (e.g., $ of GNP per MJ of energy consumed by a society) can generate the false impression that technological progress has decreased the dependence of modern economies on energy. In fact, looking at the gross ratio GNP/MJ (dollars of GNP generated per unit of exosomatic energy consumed) one could get the idea that technological progress has implied a decrease in the demand for energy in modern economies. More accurate analyses, based on opportune corrections to the two factors of such a ratio, can be used to show that this is not the case (Costanza, 1980; Cleveland et al., 1984; Hall et al., 1986; Kaufmann, 1992). In fact, “$ of GNP” has to be corrected in terms of parity purchasing power; and “MJ of energy consumed” has to be corrected for energy quality factors and the effect of different mixes of end uses. Without getting into sophisticated analyses, the distinction between intensive and extensive variables is helpful in discussing such a myth also in terms of gross values. For example, in 1991 the United States operated at a much lower ratio of energy consumption per unit of GNP than, for instance, China (12 MJ per $ versus 70 MJ per $, respectively), but because of that the United States managed to have a GNP per head much higher than China ($US 22 400 yr−1 versus $US 360 yr−1 , respectively) (World Resources Institute, 1994). When the gross ratio of energy consumption per unit of GNP is multiplied by the GNP per head, one ﬁnds that, in spite of a signiﬁcantly higher so-called “economic energy efﬁciency”, the energy consumed per United States citizen in 1991 was 11 times higher than that consumed by a Chinese citizen in the same year. An overview of energy consumption per unit of GNP, GNP per head, population size, total energy consumption, and energy consumption per head is given in Figs. 32.8a,b for the United States for the period 1950–1990. The inverse relation between energy

Mario GIAMPIETRO

consumption per unit of GNP and GNP per head is clearly indicated by Fig. 32.8a. The degree of “dematerialization” which according to such a myth has been induced by technological progress in the United States economy can be checked by analyzing data of aggregate energy consumption published by the United States Bureau of the Census (1991): a reduction in the energy consumption per unit of GNP from 113 MJ per dollar in 1950 to 25 MJ per dollar in 1990 had the effect of increasing the aggregate consumption of exosomatic energy in the United States economy from 34.5 TJ (= terajoules = 1012 joules) in 1950 to 77.0 TJ in 1990. As indicated by Fig. 32.8b, aggregate energy consumption increases not only because of an increase in consumption per head but also because of an increase in population size. Population growth is generated also by immigration, driven by the attractive economy. Strong gradients in standard of living (bio-economic pressure) among countries – generated by gradients in “efﬁciency” – tend to drive labor from poorer to richer countries (Giampietro, 1998; see pp. 738–739 below). For example, the dramatic improvement in energy efﬁciency that the state of California (U.S.A.) has achieved in the past decade [in terms of the ratio (useful energy)/(energy input)] will not necessarily curb total energy consumption in that state. In fact, present and future technological improvements are likely to be nulliﬁed by the dramatic increase in immigration, both from outside and inside the United States, which make the Californian population among the fastest growing in the world. The myth of “win/win” solutions (technology will ﬁx it) Many decision makers and scientists are used to thinking in terms of “optimum level of exploitation”, “optimization of resources use”, “technical improvements”. Unfortunately, all these concepts belong to a steadystate description of a process. When these concepts are confronted with an evolutionary view of the same process (enlarging the scale at which the problem is described), one discovers that technological changes are not likely to generate “absolute improvements”. In contrast, one can only expect to obtain trade-offs, when assessing the effects of the same change on different scales. “Recent enthusiasm regarding win–win scenarios in many cases is buoyed by scaling error. Explicit recognition of the implications of necessary trade-offs,

ECONOMIC GROWTH, HUMAN DISTURBANCE TO ECOLOGICAL SYSTEMS, AND SUSTAINABILITY

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Fig. 32.8. Intensive and extensive variables describing “changes in efﬁciency” and “dematerialization of US economy” in the period 1950– 1990. “Intensive variables” deﬁne values per unit of something, such as MJ of energy consumed per $US generated, Gross National Product (GNP) and Energy Consumption are expressed per unit of population. “Extensive variables” reﬂect the size of the system, such as Total Energy consumption and Gross National Product (obtained by multiplying an intensive variable by the size of the system, in this case population). (A) Changes in the Ratio “MJ/US$ of GNP” and “GNP per head” in the USA during the period 1950–1990; (B) Changes in Total Energy Consumption, Total Population and Energy Consumption per head in the USA during the period 1950–1990. From Giampietro et al. (1997).

both positive and negative, promotes the development of mechanisms to support losers. Failure to confront the fact that losers are consistently produced exaggerates the negative impact they have on system performance” (Wolf and Allen, 1995). New technology tends to change the terms of the problem, thereby changing the terms of the trade-offs. Unavoidable legitimate but contrasting perspectives about induced changes should be expected in human systems. What is good for a household (not paying taxes) can be bad for the community to which the household belongs; what is good for the economy

(shut down an obsolete military base) can be bad for local communities; what is good for developing nations (fast economic development fueled by export of natural resources) can be bad for natural ecosystems; what is good for the present generation (ﬁghting poverty by using better technology, and using more available stocks of resources) can be bad for future generations. As will be discussed in the ﬁnal part of this chapter, without a mediation between contrasting perspectives (a holistic evaluation of trade-offs), a quick technological ﬁx can very well become a problem rather than a solution.

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The myth of “perpetual growth” (economic growth will ﬁx it) The industrialization and development of the United States and West European countries in the 19th and early 20th centuries was based on the availability of a large amount of natural resources per head within their borders, such as fertile land and stocks of minerals and fossil energy, and/or on the importation of large quantities of natural resources through colonial exploitation and international trade. In other words, the industrial revolution of the last century occurred in a world that was still “empty” (Goodland and Daly, 1992) in the sense that, at a world level, natural resources were abundant. Being able to rely on depletion of supplies of local resources and on import, developed countries managed to remove systematically any local ecological and biophysical constraints to the expansion of their population and its consumption of resources per head. This strategy forms the basis of neo-classical economic theory that assumes (1) that supplies of natural capital, such as land, energy, and water, impose no limit to economic development; and (2) that production factors – that is, technological capital, labor, and natural capital (e.g., land) – are substitutes for one another rather than complements. The latter assumption implies that technology can overcome shortages in both natural capital and labor. However, these two assumptions represent another myth. For instance, in agriculture one supposedly can obtain the same output with more capital and less land or with more capital and less labor. However, the more the world economy approaches the biophysical limits of the biosphere, the more it becomes evident that natural and technological capital are complements rather than substitutes. Therefore, if it is true that a shortage in one of the production factors can be compensated to a certain extent by substituting one of the others, the feasibility of such a substitution drastically diminishes as natural resources become scarce. Moreover, the substitution of natural resources has a price in terms of a lower biophysical efﬁciency (a higher demand for input to get the same output) of the economic process. When the shortage of natural capital becomes severe, substitution becomes practically impossible, and biophysical constraints then limit further economic development. For instance, increasing the size and efﬁciency of ﬁshing vessels has enabled them to overﬁsh the oceans but not to increase the quantity

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of ﬁsh produced per head – which, on the contrary, is steadily declining (Ehrlich et al., 1993); technology cannot double the world arable land, double the ﬂow of the Colorado River, or replace bees for pollination. As Daly (1992) explains, “the limiting factor determining the ﬁsh catch is the reproductive capacity of ﬁsh populations, not the number of ﬁshing boats; for gasoline the limiting factor is petroleum deposits, not reﬁnery capacity; and for many types of wood it is the remaining forests, not saw mill capacity.” In spite of the obviousness of these commonsense observations the myth of perpetual economic growth remains a major pillar of the model of development followed by most developing countries. The reasons for this locking-in of cultural-mode will be discussed below (pp. 740–741). The myth of the demographic transition as a problem-solving event The hypothesis of demographic transition (that strong economic development of poor countries will stop their population growth, as has already happened in rich countries) has far-reaching political consequences, since it is used to justify the lack of concern for biophysical limits imposed by the biosphere on human expansion. Technological progress is supposed to induce, sooner or later, the completion of the demographic transition by stabilizing fertility and mortality rates at a low level, and hence to be able to avert the threat of the “demographic bomb”. I happen to believe that this optimism has no ﬁrm grounds and that a thorough analysis of the possible links between economic development and demographic behavior provides opposite indications [for a detailed analysis see Giampietro (1998) and Giampietro et al. (1997)]. Several factors which play against such a benign solution are: (1) The increase in bioeconomic pressure that led to the completion of the demographic transition in the industrialized world was, and still is, based on the unique characteristics of fossil energy use: huge levels of power per hour of labor, and the possibility to purchase fossil energy by creating monetary debts. Similar performances cannot be achieved with known alternative energy sources. It is doubtful that such a solution will be feasible in the next decades on this planet for a population of 8–12×109 people. There is simply not enough fossil energy, nor enough technical and ﬁnancial

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capital, to build the required amount of exosomatic devices per head in order to make the whole of humankind rich in a few decades. (2) Gradients in bioeconomic pressure, both within and among countries, generate gradients in fertility rates and migration ﬂows that can neutralize the mechanism stabilizing the zero-growth equilibrium. That is, in the real world, additional units of human mass are produced where they are cheaper in biophysical terms (in countries operating at a low level of bio-economic pressure), and are then transferred to places where they provide a higher biophysical return (in countries operating at a high level of bioeconomic pressure). This is the biophysical mechanism “pumping” immigrants from developing countries into developed countries. (3) A halt in population growth obtained by reduced fertility signiﬁcantly decreases the labor supply through aging of society, and signiﬁcantly increases the demand for labor in the service sector. Therefore, the stability of this equilibrium would require a slow but continuous increase in labor productivity in productive sectors of the economy. That implies a further acceleration of throughputs in agriculture, energy sectors, mining, and ﬁsheries. Ironically, this happens when the already excessive environmental loadings linked to the elevated level of economic growth require a reduction in the intensity of throughputs in these sectors. The myth of the collapse due to clash against external biophysical limits To complete the list of misperceptions about the issue of sustainability, it should be observed that the scientiﬁc debate on possible reasons for the future collapse of modern civilization is mainly focused on external biophysical constraints. That is, serious problems are expected to occur when biophysical constraints generate shortages in the supply of required inputs. However, the increased instability of different patterns of socio-economic organization all over the world is pointing at another possible cause of collapse of models of development, through internal causes. Civil wars in Somalia, Burundi, former Yugoslavia, Iraq, Algeria, and Afghanistan, coupled with the widespread unrest in many of the republics of the former Soviet Union are examples of a diffuse destabilization of social order resulting from a wide discrepancy between

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internal and external pressures. When the gap between aspirations of individuals and performance of socioeconomic systems is large, and becoming wider with the passage of time, there is a serious risk that the social and economic fabric making up society will break down. This phenomenon can be related to a lack of synchronization between (1) the speed at which internal bio-economic pressure is building up (the example given by developed countries is raising the minimum acceptable material standard of living in almost every corner of the world). This rapid increase in expectations especially affects young people, and is probably accelerated by the spread of television to many developing countries; (2) the speed at which changes in the economic process are increasing the material standard of living in these socio-economic systems. To make things worse, this worldwide surge in internal pressure is occurring at the very same moment when the external pressure on human societies, in the form of shortage of natural resources per head, is also increasing at the world level. As a result of this sudden “crunch”, social organization of rural communities is crumbling and cities are under heavy stress by the uncontrolled explosion of their population. These trends spell little good for future world development. In fact, without a robust socio-economic fabric, the simple addition of technology, economic capital, and natural resources cannot sustain a decent standard of living in any society. Kaplan (1994), in his article “The coming anarchy”, argues: “It is time to understand ‘the environment’ for what it is: the national security issue of the early twenty-ﬁrst century. The denser the human population becomes, the more countries are forced to closely interact and compete for the shrinking endowment of natural resources. Intensiﬁcation of such interaction may result in an emphasis of differences in cultural, religious, and political identities, and standard of living, and may precipitate international conﬂicts.” Why is the “steady-state view” attractive? The concept of sustainable development is generally viewed as aiming at an optimum level of exploitation of natural processes coupled with a continuous improvement in technological efﬁciency. The idea, however, that it is possible to arrive at a stabilized situation through a series of smooth, linear adjustments should also be considered a myth. “Optimum levels

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of exploitation must be determined by trials and error . . . Initial over-exploitation is not detectable until it is severe and often irreversible” (Ludwig et al., 1993). Additionally, “the tendency of using equilibrium models makes it seem as if the perpetuation of the status quo were up to us; and if we were just careful enough to behave in the proper way we could be sure that we had our lives in our own hands” (Dyke, 1988). Indeed, the widespread denial by many “cornucopians” of risks associated with technological development seems to conﬁrm that humans are not willing to accept the idea of a world regulated by natural selection, nor to face the fact that they are part of structures stabilized by dynamic processes far from equilibrium and therefore continuously at risk of collapsing. The economic paradigm currently used all over the world to allocate resources in the process of technological development (decision making) has proven to be unable to deal with the problem of longterm sustainability of human development. Still, such a paradigm is preferred by decision makers, since it enables them to deﬁne cost/beneﬁt analyses to be used as surrogates for a value based assessment (beneﬁt = good, cost = bad). However, an economic description of cost/beneﬁt must be based on a clear deﬁnition of a speciﬁc problem at a particular point in time and space (and the assumption that it is widely representative – the ceteris paribus hypothesis). Unfortunately, globalization and concern for future generations now imply that everything depends on everything else. The very beneﬁt deﬁned on one space– time scale (e.g., 10 yr), and/or seen from the viewpoint of one social group, becomes a cost when viewed on a different space–time scale (e.g., 100 yr) and/or by a different social group. The challenge of sustainable development is to achieve solutions that are compatible at the same time both with biophysical constraints (they are ecologically compatible) and with human aspirations (they are socially acceptable) when viewed from the different perspectives that can be found in the world. This would require the selection of strategies of development accepted unanimously by all the different groups into which humankind can be divided on the basis of cultural identity, age, and geographical location. On the other hand, these shared goals regarding the issue of sustainability are simply not there. What is badly needed is for humankind to learn how to regulate themselves within the external constraints provided by the limited endowment of natural

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resources. Things are changing so fast that the need to urgently establish a new set of rules determining an internal control on human expansion could imply the re-evaluation of our cultural identity. For example, moral obligations and ethical values have to be put in perspective with the larger spatiotemporal scale (e.g., respect for the resilience of the biosphere today means respecting the quality of life of future generations). This should include for the different human groups the respect of alternative cultural identities with contrasting but still legitimate “value dependent” deﬁnition of priorities. Such an exercise can very easily encounter controversial applications – see, for example, the “lifeboat ethics” of Hardin (1968). In fact, cultural changes are slow, painful, and difﬁcult to make, especially since it is difﬁcult to reach a consensus on the description of what the problems to be solved are in the ﬁrst place (what one is willing to give away in order to get something new). However, an excessive conservatism against changes in cultural identity can also be dangerous, especially if one is concerned with the protection of fragile social groups. Weaker social groups (e.g., the poor, women, ethnic minorities, the disabled, the elderly, or children) will be the most impacted by a sudden, dramatic readjustment of social organization toward an economy of scarcity, if such changes are allowed to occur by default. Implications for action: the need to ﬁght cultural resistance to change Step one: facing the reality, stopping the denial When looking at the current predicament of humankind by adopting a steady-state picture (accepting the current technology and resources as given) one is forced to conclude that: the current path of economic growth of the world economy is already based on activities that, when assessed at the global level, are (1) depleting the very same resources on which they depend; and (2) disturbing the natural processes which provide the environmental services needed for the survival of the human species. According to present technology, and present endowment of natural capital, too many people are looking for too high a material standard of living. When one looks at future scenarios for technology and resources, the situation does not appear brighter. The direction of technological development is currently trapped into a “ratchet effect” (Ludwig et al., 1993), meaning that science and technology

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have no alternatives but to “improve” the standard of living of a world population that is dramatically expanding. Consequently, technological development is becoming a desperate and continuous search for “quick ﬁxes”. When prospects on the “demand” side (current size of population and trend of population growth, widespread poverty in many areas of the world) and the “supply” side (fossil energy depletion and the shortage of natural resources and ﬁnancial capital for a worldwide economic development) are considered, it is easily recognized that the feasibility of a future “sustainable development” at the world level will have to be based on technological breakthroughs that have yet to be made (e.g., solar technology, genetic engineering to increase food production). Many scientists are conﬁdent that these goals are well within the reach of human ingenuity, but many others are not. Thus, the current strategy driving human development is no more than a sort of gambling of the future of humankind, which has been accepted by the world community without much discussion at cultural or political levels. But even if it be accepted that gambling cannot be avoided (after all, that is what life is about), one should at least be able to deﬁne the terms of the bet (what can be gained, what could be lost) and the rules of the game (who is calling the bet and who will pay for or gain by it); unfortunately, at present, these terms are not at all clear. The hope for technological breakthroughs is becoming increasingly similar to the expected arrival of the “7th cavalry” in the ﬁnal scene of old-fashioned western movies. What will happen if they do not arrive in time? In this situation, technology is at risk of becoming a problem rather than a solution. The sustainability battle is being fought with inadequate weapons The scientiﬁc reductionistic paradigm used in developed societies today often has no predictive power when dealing with the evolution of the real world. The “old Newtonian view of scientiﬁc knowledge” cannot be used to predict and control the evolution of complex systems such as ecosystems and socioeconomic systems. The complex nature of the real world was ignored until now by reductionistic science, since traditional scientiﬁc disciplines could afford to describe events on a delimited space–time scale, and therefore process only a small amount of information at a time. In the past, individual scientiﬁc disciplines used to ignore the hierarchical and the evolutionary

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nature of studied processes, dividing them into smaller and smaller parts, or looking for linear relations among selected variables within simpliﬁed descriptions of events (Funtowicz and Ravetz, 1994b). Concepts such as globalization (enlargement of space scale) and sustainability (enlargement of time scale) are impossible to deal with by using a reductionistic scientiﬁc paradigm. When too much information has to be processed within the same analysis and when systems have a non-linear behavior, reductionistic science is bound to fail. Nowadays, scientists are forced to admit their inability to ban uncertainty from their description of reality (e.g., predict the weather of New York in 60 days). The “business as usual” way of decision-making must be challenged The unsustainable pattern of “perpetual growth” proposed by developed countries is clashing with traditional cultures all over the world. The unchecked expansion of the model of production and consumption of developed countries is wiping out the diversity of experiences and the cultural values accumulated by centuries of human history. Tragically, in this way, such a model of development is destroying valuable information that could be used to correct present dangerous trends. Often in the past socio-economic systems experienced a dangerous cultural-resistance to change that pushed them above the carrying capacity of their supporting ecosystems and therefore to a collapse (Tainter, 1989). However, at the current degree of globalization, humankind is risking a repeat of such a pattern on the world-wide scale. What is wrong with the basic default assumptions is that the free market economy and technological development per se have no value except in relation to the choices and goals of the societies in which they operate. Unfortunately, nobody is currently deﬁning goals or making choices at the global level regarding sustainability. Today, the world free market and world technological development are going at full speed, as never before in the history of human civilization, while nobody is steering this runaway train (no control from the cultural level) to avoid an excessive disturbance to the biosphere. Looking for causes of the current predicament Over the whole world, existing cultural identity is under stress and it is no longer possible to deﬁne

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“choices and goals” on that scale. This is due to the fact that human cultural identity (complex of values, rules, institutions, and laws) is accustomed to changes that occur very slowly. In a time of fast changes of boundary conditions, individuals do not have the time to reinforce our cultural identity by confronting the correctness of proposed policies in terms of ability to deal with the reality. At the present time, it is not even possible to deﬁne the boundaries of the society (community, country, humankind, biosphere) for which choices and goals should be set. Cultural evolution has prepared humankind to ﬁlter changes taking place on a time scale of decades, which until now were considered just as random ﬂuctuations. Therefore, cultural identity is not able to detect and react to them appropriately. Culture refers to human experience accumulated over a large time scale (e.g., hundreds of years); even for individuals, there is the problem of continuously updating the perception of “values” when the context is changing too fast. One’s judgments, in fact, are based on the picture one has of the world. Today, however, even a period of a few years is sufﬁcient to change such a picture dramatically. For example, 10 years ago communism was a legitimate economic option endorsed by half of the world as an alternative to capitalism. Before the industrial revolution, all human civilizations were forced to develop a “cultural” control that, in one way or another, limited the rate of their expansion – such as cultural control of fertility, and a hierarchical control to regulate access to resources. With the industrial revolution, stocks of fossil energy and technology freed humans (at least temporarily) from strict external ecological control. This made it possible to achieve material standards of living much higher than in the past. At the same time, this made obsolete the old set of cultural controls developed to avoid ecological constraints (such as famines and plagues). The speed of change driven by industrial revolution was so fast, that cultural striving for modernization focused mainly on ﬁghting the old cultural paradigms aimed at curtailing the expansion of human activity – e.g., positivistic science ﬁghting against other ways of thought, especially in the form of religion and superstition. The need to develop new forms of cultural control (values) to slow down human expansion was simply ignored. The neo-classic economic paradigm provided the justiﬁcation for doing so, by introducing the myth of an inexhaustible supply of resources (the unlimited ability of technology to

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substitute for limiting resources). The basic dogma underlying modern civilization was – and, unfortunately, still is – that nature must be adapted to human wants (through science and technology) and not vice versa. As noted earlier, this dogma is generated by a mixture of ignorance and arrogance. The human dream of replacing the work of self-organization of the biosphere by technology is simply ludicrous, but yet has tragic consequences. Looking for a cure: ﬁghting dangerous cultural resistance to change Global trade is not necessarily good for sustainability The neoclassic economic paradigm endorses the idea that unlimited trade is a useful tool to enhance world economic growth (Kaldor, 1980). The hidden assumption here is that the global level is the optimal scale at which to analyze and manage economic development. There are very good reasons for questioning this basic assumption, especially when the issue of sustainability is considered. Increasing the scale of national economic systems to arrive at a unique world economic system (e.g., expanding world trade as much as possible through international agreements) has both positive and negative consequences. A larger scale of the economic process translates into economies of scale, comparative advantages, and a reduction in smaller-scale perturbations (improvements in the steady-state picture). On the other hand, the very large scale of the world economy represents a ﬁlter on the feed-back signals from local ecosystems. When one operates within a world market for rice, one is no longer aware of the different ecosystems used for rice production. World markets of commodities cannot read feed-back messages coming from individual agroecosystems. Changes in the supply of a commodity due to negative ecological feed-backs will be detected at the world level only when they have reached a scale (an aggregate effect) implying major irreversible effects at the local level. Therefore, the goal of harmonizing the circulation of energy and material in the economic process with characteristics of local ecosystems (increasing its ecological compatibility) can be more easily achieved if the space–time scale of the activities of production and consumption in the socioeconomic system is similar to the space–time scale of the cycles of matter occurring in local ecosystems. This would make it easier for the economic process to respond quickly to changes and signals coming from its

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surrounding ecosystems, in spite of their intricate and varying nature (Norgaard, 1981). There are no “good” or “bad” models of development Often the conventional approach to development (“top-down” drive for change) is criticized in favor of an alternative approach to development (“bottom-up” drive for change). However, one cannot expect that the direction of development to be taken in a particular situation can always be determined a priori. By the “conventional, top-down approach” I mean here the approach adopted by many United Nations development agencies: after deﬁning the problems in their headquarters, they send experts to a particular place to “ﬁx” them. By “alternative, bottom-up approach” I mean the approach adopted by many NGOs (NonGovernmental Organizations) in making an effort to develop “grass-root” movements. Their ﬁnal goal is to preserve as much as possible of the original identity of the community and ecosystem targeted by the project of development. Neither of these two approaches should be considered the right one or the wrong one by default. The opportunity to focus more on “top-down” or “bottomup” approaches depends on the particular situation, and should be deﬁned in terms of the need to increase (or decrease) the connectedness of the particular socioeconomic system which is supposed to change with its socioeconomic context. A short discussion follows. The reasons for “bottom-up”: When a very poor socio-economic system interacts with a very rich one, “overconnectedness” is detectable from the fact that such an interaction generates excessive friction. The poor system is forced to change its internal organization too quickly. When the pace of change is not compatible with the lag time required to adjust internal social rules, the cultural identity of the system is at risk of disappearing. Poor socio-economic systems tend to adapt as fast as possible to any new set of risks and opportunities with which they are confronted. Therefore, social groups higher in the hierarchy of power (urban population, local elites) are the ﬁrst to be involved in the process of adaptation. This will result, at the national level, in further friction within the poor society: local elites will force changes on weaker social groups. In conclusion, fast changes always occur in one direction. When such a one-way process of adaptation (“topdown”) is spread on a large scale (e.g., all over the

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developing world), then there is a problem. The risk is that expansion of the model of development adopted by industrialized countries can wipe out the diversity of cultural experiences, values, knowledge, and alternative economic paradigms that exist in developing countries. That is, developed countries are propagating and amplifying throughout the world their own system of cultural control, without receiving any feed-back about possible adjustments of it over different sets of boundary conditions (in different areas of the planet). For this reason it is extremely important to preserve different cultural identities found in the world. That is, when facing an excessive “top-down” drive for change, it is important to promote an opposite force from the bottom, able to slow down the pace of change to a certain extent. This would buy the necessary time to negotiate trade-offs at the cultural level related to the tragedy of change (one must lose something in order to get something new). On the other hand, an excessive period of negotiation means prolonging the unsatisfactory situations needing to be changed, which in general are disadvantaging weak social groups. The reasons for “top-down”: In a deﬁned society, socio-economic characteristics (related to the cultural identity of a community) and technical coefﬁcients (determined by the availability of natural and humanmade capital) affect each-other. On account of this mutual interaction, it is not possible to change unsatisfactory situations in a reasonable time, without a ﬁrm action pushing the system away from the status-quo. In this case, a quick injection of external inputs affecting the factors determining the unsatisfactory situation is the only way out. For example, in order to achieve certain goals within a given deadline, it could become crucial to (1) change the social status of women (which is determined by existing cultural identity); and/or (2) import those natural resources, technology, and ﬁnancial capital that are required (e.g., open a closed society to trade when demographic pressure has already reduced the availability of critical natural resources per head). When the urgency of change is high and there is no time to work out an internal solution, a strict respect for speciﬁc local issues (e.g., cultural diversity, preservation of a particular endangered species) can become an obstacle to the process of change. Clearly, balancing the trade-offs implied by the adoption of a “top-down” model of development (imposing fast changes and killing diversity) and/or a “bottom-up” one (preserving diversity but slowing

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down changes) is not simple. Within the two interacting societies (developed and developing) there are always different social groups with different perceptions about how important it is (1) to change as soon as possible; and (2) to preserve cultural and biological diversity. Such a different perception in turn will depend on (1) the gap between the aspirations of the people living in the society and current performance of the socioeconomic process; (2) who is making the decisions (whether they are within or outside the society to be changed; whether they are among the rulers or the ruled within the society to be changed). In conclusion, in discussing models of development there are no solutions that represent absolute “improvements” (in evolving complex systems they do not exist) but only solutions that imply trade-offs, which, in general, are difﬁcult and controversial to compare. Developed economies are not based on “rational behavior of economic actors” As soon as one assesses the evolutionary effect (long-term perspective) of the aggregate actions of economic actors in developed countries, one should challenge the assumption of their rationality. The neoclassical model of development based on the rational behavior of Homo economicus leads to the destruction, as fast as possible, of the resources on which society depends in order to get the maximum return from one’s activities. Put in another way, the rational behavior of market economies totally ignores the heritage of the ancestors of the present generation and the fate of their great-grandchildren. This is due to the fact that in economic terms one cannot discount with adequate accuracy future values against present values. Can this be termed “rational behavior”? Can the behavior of members of pre-industrial civilizations be dismissed as “not rational” simply because they were keeping a strong historical perspective of their societal activity, that is, considering ancestors and descendants in the process of every-day decision making? Reasons for optimism: the Robinson Crusoe paradox The evident malfunctioning of the current system of control on world development is also a reason for optimism. After all, when dealing with human systems, cultural mode-locking (resistance to changes in the way of looking at and deﬁning problems) is the most persistent and important cause of lack of adaptability

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in response to changing boundary conditions. As soon as one is forced to deal with the fact that the terms of reference to be used for regulating behavior are dramatically changed, human ability to adapt to disturbance is simply amazing. This ability to make sudden changes is what I am proposing as the “Robinson Crusoe paradox”. Each person has experienced in her or his daily life an example of such a dramatic set of changes. That is, during a period of steady-state one feels her or his daily life as totally constrained by existing boundary conditions (by a heavy mix of work, social and family commitments). Then, one is suddenly hit by a perturbation large enough to generate a collapse in the set of steady-state constraints. Such a perturbation can be falling in love, being hospitalized after a serious car accident, being dismissed from a job, or – as in the case of the character giving the name to the paradox – becoming shipwrecked on a desert island. At that particular moment, when faced with the need to change dramatically the way one’s daily life has been managed, one has the opportunity to realize the remarkable ability of the human system to adapt. After a transitional period needed to tune internal characteristics to external boundary conditions, the human system (be it an individual person, a small group, or a country at war) will ﬁnd new steady-state solutions, new routines, for its every-day life. These new equilibria would have been totally unthinkable before the perturbation took place. This almost magical ability of adaptation to novelties (dealing with disturbance) is probably what is missing in the analysis provided by neo-Malthusian pessimists. It is probably true that current civilization, as one knows it, will collapse in the next century (after all, all civilizations have to pass, exactly as humans must die). However, this expression means only that human civilization will become something else – something that cannot be imagined at present. This should not necessarily be equated with a major and negative cataclysm for humankind; on the contrary, a dramatic change can also be perceived as an opportunity to make desired changes. To make things even more intriguing one should consider another special property typical of human systems. Humans are characterized by what Funtowicz and Ravetz (pers. comm.) suggest may be called “passional entailment”. Such a name emphasizes the difference from “rational entailment”. That is, the present behavior of human systems is affected not

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only by experience about the past (rational entailment which is conditioning the possible choices), but also by the “virtual future” that humans are generating for themselves (based on their vision of future states of the system to which they belong). This virtual future (be it a project or a dream to make true), when stabilized on a scale large enough in time and space by their common and continuous effort, will be able consistently to affect everyday choices. In this way, it will progressively pull present states of the system toward itself. Human systems are subject to self-fulﬁlling prophecies, and because of that they are well known for their ability to achieve seemingly impossible results. However, the basic requisite for doing that is that people believe in and are ready to ﬁght for them. This is the reason why such an ability is enhanced when people are facing major challenges (e.g., wars, situations of emergency, major transitional periods). It is my opinion that the current predicament of humankind in confronting the issue of sustainability represents one of these critical moments. It is also my opinion that the potential of humankind for dealing with any type of challenge is virtually unlimited. On the other hand, such a tremendous force for adaptability must be freed as soon as possible. This would require a general questioning of all the default assumptions that often are hampering one’s ability to cope with changes. One will need to do things in a different way very soon. One should stop being scared by that, but rather think about what one would like to do differently. Humans have to deﬁne a totally new set of terms of reference for their development. In this new deﬁnition, scientiﬁc, political, ethical, and socio-economic analyses should be mixed in a way that has never been done before. On the other hand, this is the century in which humankind has done many things for the ﬁrst time.

REFERENCES Alexandratos, N. (Editor), 1995. World Agriculture: towards 2010, an FAO Study. Wiley, Chichester, 488 pp. Blanchet, D., 1991. On interpreting observed relationships between population growth and economic growth: a graphical exposition. Popul. Dev. Rev., 17: 105–114. Brown, L., 1988–1996. The State of the World, Series. Worldwatch Institute, Washington, D.C. Cherfas, J., 1991. Skeptics and visionaries examine energy savings. Science, 251: 154–156. Chesnais, J.C., 1992. The Demographic Transition. Clarendon Press, New York, 625 pp.

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Cleveland, C.J., Costanza, R., Hall, C.A.S. and Kaufmann, R., 1984. Energy and the US economy: a biophysical perspective. Science, 225: 880–889. Conforti, P. and Giampietro, M., 1997. Fossil energy use in agriculture: a cross-country comparison. Agric. Ecosyst. Environ., 65: 231–243. Costanza, R., 1980. Embodied energy and economic valuation. Science, 210: 1219–1224. Council for Agricultural Science and Technology, 1994. How Much Land Can Ten Billion People Spare for Nature. CAST, Ames, Iowa (prepared by Paul E. Waggoner Task Force Report, No. 121, February 1994). Daly, H.E., 1992. From empty-world economics to full-world economics: recognizing an historical turning point in economic development. In: R. Goodland, H.E. Daly, and S. El Serafy (Editors), Population, Technology, and Lifestyle. Island Press Washington, DC., pp. 23–37. Dyke, C., 1988. Cities as dissipative structures. In: B.H. Weber, D.J. Depew and J.D. Smith (Editors), Entropy, Information, and Evolution. MIT Press, Cambridge, Massachusetts, pp. 355–367. Ehrlich, P.R., Ehrlich, A.H. and Daily, G.C., 1993. Food security, population, and environment. Popul. Dev. Rev., 19: 1–32. Food and Agriculture Organization of the United Nations, 1996. Food Balance Sheets. FAO, Rome. Funtowicz, S.O. and Ravetz, J.R., 1994a. The worth of a song bird: ecological economics as a post-normal science. Ecol. Econ., 10: 197–207. Funtowicz, S.O. and Ravetz, J.R., 1994b. Emergent complex systems. Futures, 26: 568–582. Giampietro, M., 1994a. Using hierarchy theory to explore the concept of sustainable development. Futures, 26: 616–625. Giampietro, M., 1994b. Sustainability and technological development in agriculture: a critical appraisal of genetic engineering. Bioscience, 44: 677–689. Giampietro, M., 1997a. Linking technology, natural resources and socioeconomic structure of human society: a theoretical model. In: L. Freese (Editor), Advances in Human Ecology, Vol. 6. JAI Press, Greenwich, Connecticut, pp. 73–128. Giampietro, M., 1997b. Socioeconomic pressure, demographic pressure, ecological loading and technological changes in agriculture. Agric. Ecosyst. Environ., 65: 201–229. Giampietro, M., 1997c. Socioeconomic constraints to farming with biodiversity. Agric. Ecosyst. Environ., 62: 145–167. Giampietro, M., 1998. Energy budget and demographic changes in socioeconomic systems. In: V. Gansl¨osser and M. O’Connor (Editors), Life Science Dimensions of Ecological Economics and Sustainable Use. Filander Verlag, F¨urth, pp. 327–354. Giampietro, M., Bukkens, S.G.F. and Pimentel, D., 1994. Models of energy analysis to assess the performance of food systems. Agric. Syst., 45: 19–41. Giampietro, M., Bukkens, S.G.F. and Pimentel, D., 1997. Linking technology, natural resources and socioeconomic structure of human society: examples and applications. In: L. Freese (Editor), Advances in Human Ecology, Vol. 6. JAI Press, Greenwich, Connecticut, pp. 129–199. Goodland, R.J.A. and Daly, H.E., 1992. Ten reasons why northern income growth is not the solution to southern poverty. In: R. Goodland, H.E. Daly and S. El Serafy (Editors), Population,

746 Technology and Lifestyle. Island Press, Washington, DC, pp. 128– 145. Hall, C.A.S., Cleveland, C.J. and Kaufmann, R., 1986. Energy and Resource Quality. Wiley, New York, 577 pp. Hardin, G., 1968. The tragedy of the commons. Science, 162: 1243–1248. Jevons, F., 1990. Greenhouse – a paradox. Search, 21: 171–172. Jevons, W.S., 1865. The Coal Question. Augustus M. Kelley, New York, NY. (Reprint of the Third Edition 1906). Kaldor, N., 1980. The foundations of free trade theory and their implications for current world recession. In: E. Malinvaud and J.P. Ftoussi (Editors), Unemployment in Western Countries. MacMillan Press, London, pp. 85–100. Kaplan, R.D., 1994. The coming anarchy. Atl. Monthly, 273: 44–76. Kaufmann, R., 1992. A biophysical analysis of the energy/real GDP ratio: implications for substitution and technical change. Ecol. Econ., 6: 35–56. Kendall, H. and Pimentel, D., 1994. Constraints to the world food supply. Ambio, 23: 198–205. Khazzoom, J.D., 1987. Energy saving resulting from the adoption of more efﬁcient appliances. Energ. J., 8: 85–89. Leopold, A., 1949. A Sand County Almanac, and Sketches Here and There, 1987 edition. Oxford University Press, New York, 228 pp. Ludwig, D., Hilborn, R. and Walters, C., 1993. Uncertainty, resource exploitation, and conservation: lessons from history. Science, 260: 17. Malthus, T., 1803. An Essay on Population, 1999 edition. Prometheus Books, Amherst, New York, 400 pp. Meadows, D.H., Meadows, D.L., Randers, J. and Behrens III, W.W., 1972. The Limits to Growth. Universe Books, New York, 205 pp. Newman, P., 1991. Greenhouse, oil and cities. Futures, May: 335–348. Norgaard, R.B., 1981. Sociosystem and ecosystem coevolution in the Amazon. J. Environ. Econ. Manage., 8: 238–254. Odum, H.T., 1983. Systems Ecology. Wiley, New York, 644 pp. O’Neill, R.V., DeAngelis, D.L., Waide, J.B. and Allen, T.F.H., 1986. A Hierarchical Concept of Ecosystems. Princeton University Press, Princeton, 256 pp. Pastore, G., Giampietro, M. and Mayumi, K., 1996. Bioeconomic pressure as indicator of material standard of living. In: Fourth

Mario GIAMPIETRO Biennial Meeting of the International Society for Ecological Economics, August 4–7, 1996. Boston University, Boston, Massachusetts. Pimentel, D., Stachow, U., Takacs, D.A., Brubaker, H.W., Dumas, A.R., Meaney, J.J., O’Neil, J.A.S., Onsi, D.E. and Corzilius, D.B., 1992. Conserving biological diversity in agricultural/forestry systems. BioScience, 42: 354–362. Pimentel, D., Harman, R., Pacenza, M., Pecarsky, J. and Pimentel, M., 1994. Natural resources and an optimum human population. Popul. Environ., 15: 347–369. Pimentel, D., Harvey, C., Resosudarmo, P., Sinclair, K., Kurz, D., McNair, M., Crist, S., Spritz, L., Fitton, L., Saffouri, R. and Blair, R., 1995. Environmental and economic costs of soil erosion and conservation beneﬁts. Science, 267: 1117–1123. Postel, S.L., Daily, G.C. and Ehrlich, P.R., 1996. Human appropriation of renewable fresh water. Science, 271: 785–788. Simon, J.L., 1998. The Ultimate Resource 2. Princeton University Press, Princeton, 656 pp. Tainter, J.A., 1989. The Collapse of Complex Societies. Cambridge University Press, Cambridge, United Kingdom, 250 pp. United States Bureau of the Census, 1991. Statistical Abstract of the United States 1991. U.S. Department of Commerce, Washington, DC. United States Bureau of the Census, 1994. Statistical Abstract of the United States 1994. U.S. Department of Commerce, Washington, DC. Vitousek, P.M., Ehrlich, P.R., Ehrlich, A.H. and Matson, P.A., 1986. Human appropriation of the products of photosynthesis. Bioscience, 36: 368–373. Wilson, E.O. (Editor), 1988. Biodiversity. National Academy Press, Washington, DC, 521 pp. Wolf, S.A. and Allen, T.F.H., 1995. Recasting alternative agriculture as a management model: the value of adapt scaling. Ecol. Econ., 12: 5–12. World Development Forum, 1989. World Development Forum, 7: 1. World Resources Institute, 1992. World Resources 1992–93. Oxford University Press, New York, 385 pp. World Resources Institute, 1994. World Resources 1994–95. Oxford University Press, New York, 400 pp.

Chapter 33

DISTURBANCE IN TERRESTRIAL ECOSYSTEMS: SALIENT THEMES, SYNTHESIS, AND FUTURE DIRECTIONS Michael R. WILLIG and Lawrence R. WALKER

INTRODUCTION

Disturbance has become a leitmotif of contemporary ecology. It affects all terrestrial biomes and may play a critical role in determining key structural and functional aspects of many ecosystems. Theories of disturbance have changed the way that ecological systems are viewed; rather than static entities, landscapes comprise patches affected by various disturbances and undergoing temporal changes as a consequence of succession (i.e., they are shifting mosaics). Simple equilibrial approaches may be inadequate; models that incorporate meta-stability, chaotic behavior, or complex-system perspectives may be more useful. Recognition of the non-equilibrial or dynamic nature of ecosystems is especially important because humans, as potentially intense agents of disturbance on a global scale, are effecting changes in structural and functional aspects of ecosystems, as well as modifying disturbance regimes to an extent to which few species are adapted. Understanding the ecology of disturbed ground, and how to manage heterogeneous and increasingly fragmented landscapes effectively, while balancing concomitant needs of human societies, may become the greatest challenge in the next century. Success at meeting this challenge will require the dedicated efforts and cooperation of the scientiﬁc establishment, governmental and private agencies, and local populaces throughout the world. In the following section, we summarize fourteen recurrent themes that appeared throughout the chapters of this volume, and are pertinent to understanding disturbance ecology and management. We do not extensively survey other disturbance literature, but instead refer the reader to the extensive literature citations in the preceding chapters. In the second part

of this chapter, we present two models that crystallize our current understanding of the relationships between disturbance and ecosystem response as mediated by patch dynamics and successional processes. We also suggest future directions for research on disturbance ecology and management.

SALIENT THEMES

Of the recurrent themes that emerged from the diverse contributions to this volume (Fig. 33.1), perhaps the most important is the interaction between disturbances. The various elements of a disturbance regime are not independent or random in occurrence; rather, particular agents related to earth, air, water, ﬁre, or the biota form an interacting network of correlated disturbances (Walker and Willig, Chapter 1). Directly or indirectly, such disturbances result in spatial heterogeneity and temporal variation in ecological characteristics of sites. In view of this, we focus on interactions early in our exposition, followed by a consideration of spatial heterogeneity. Aspects of the physical environment which cause heterogeneity (e.g., slope, aspect, elevation, geological characteristics) combine with the disturbance regime to produce a geographic mosaic of patches which differ to various degrees in abiotic and biotic features. Equally important, postdisturbance changes in the ecological characteristics of sites occur as a result of dynamic interactions among the surviving biota, immigrating taxa, and the disturbance regime, continually changing the abiotic and physical characteristics of sites. Taken together, these changes represent ecological succession. After considering the importance of disturbance regime,

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Fig. 33.1. Major themes of the chapters in this volume include interactions, spatial heterogeneity, succession, competition, nutrient cycling, productivity, stability and resilience, predictability, thresholds, biodiversity, functional redundancy, invasive species, restoration and management, modeling. The proportion of chapters [excluding the Introduction (Chapter 1) and the current chapter] in each of the four sections of the volume (natural disturbances, n = 12; anthropogenic disturbances, n = 7; natural processes, n = 7; and human responses to disturbance, n = 5) that consider each theme in a substantive manner is indicated by the height of bars in the corresponding diagram. Table 1.2 (p. 14) identiﬁes which of these themes are discussed in a particular chapter.

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spatial heterogeneity, and ecological succession, we continue our discussion by focusing on disturbancerelated characteristics of systems: nutrient cycling and productivity, stability and resilience, predictability, and threshold effects. Thereafter, we address two interrelated themes of disturbance ecology (biodiversity and invasive species) at the level of populations and communities, which are particularly important from the perspective of conservation biology, and we follow with a consideration of restoration and management. We address modeling as an integral part of synthesis, and conclude with a suite of recommendations for areas of research that will advance the conceptual development of the ecology of disturbance. Interactions Disturbance regimes include a number of interacting elements and are illustrated for the tabonuco (Dacryodes excelsa) forest in Puerto Rico (Fig. 33.2). Some disturbance elements are much more likely to trigger others (e.g., in Puerto Rico, hurricanes may induce drought, ﬂooding, tree-falls, and landslides, and even affect forestry practices). Conversely, some disturbance elements are predisposed to occur as a consequence of previous disturbances (e.g., in Puerto Rico, herbivory may be stimulated as a consequence of stress on plants induced by hurricanes, droughts, tree-falls, forestry practices, road construction, or landslides). Clearly, the occurrence of a particular disturbance modiﬁes the likelihood that at least a subset of the elements in the disturbance regime subsequently will impinge on the same area. From a biogeographic perspective, interactions are pervasive and characterize the disturbance regimes of all biomes. Consequently, interactions are considered in many of the chapters in this book (Fig. 33.1). Indeed, the interactive nature of these elements contributes to the difﬁculty of associating particular abiotic and biotic effects unequivocally with their causal agents. For example, anthropogenic activities in Los Angeles enhance the dry deposition of nitrogen in portions of the eastern Mojave Desert (Rundel, Chapter 10). Enhanced levels of nitrogen then facilitate expansion of cover by annual grasses, the dead stalks of which subsequently act as kindling for ﬁre. Fire is then fueled by annuals between desert shrubs, which were previously outside the range of this disturbance. Finally, declines in the abundance of shrubs may have concomitantly facilitated the invasion of the exotic

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herb Erodium cicutarium in regions of the western Sonoran Desert, especially during wet years. Clearly, disturbances can interact in complex fashions, and their effects are often synergistic rather than additive. From an applied perspective, chapters dealing with restoration (Hobbs, Chapter 29) and environmental policy (Barrow, Chapter 28) clearly recognize the interactive nature of disturbance. In fact, they consider wise management policies and activities to be predicated on understanding how human intervention (or absence of it) interacts with natural disturbance regimes to attain desired goals. Humans, as agents of disturbance, are unique in having the ability consciously to modify the frequency, extent, and intensity of their own activities in order to minimize or maximize the severity of their effects on populations, communities, or ecosystems. Humans may be equally unique as biotic agents of disturbance which have the potential to jeopardize their continued existence. Especially as a result of recent technological advances, humans have become an invasive species with the capability to cause severe disturbances threatening the continued existence of many species, the structure of communities, and functioning of ecosystems.

Fig. 33.2. The disturbance regime of an area comprises a number of interacting agents. For example, in the tabonuco forest of Puerto Rico, hurricanes, herbivory, landslides, roads, forestry, treefalls, ﬂoods, and droughts affect the structure and function of the forest. The occurrence of one agent of disturbance (e.g., a hurricane) may enhance the likelihood of subsequent disturbances to various degrees, with stronger enhancements indicated by solid arrows (e.g., herbivory, landslide, tree-fall, and drought) and weaker enhancements indicated by dashed arrows (e.g., forestry). Some agents of disturbance have reciprocal effects, indicated by doubleheaded arrows (e.g., landslides enhance the likelihood of ﬂooding, and ﬂooding enhances the likelihood of landslides).

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Spatial heterogeneity The disturbance regime of an area comprises different agents (e.g., herbivore outbreaks, tree-falls, hurricanes, droughts, termite mounds, anthropogenic activities). As a consequence, the effect of a particular disturbance (e.g., hurricane) on the structure and function of a biological system, at the population, community, or ecosystem level, depends on the distribution, intensity, frequency, and extent of each agent; their degree of spatial and temporal correlation; their synergisms; and the current status of the biological system upon which they impinge (see Figs. 33.3 and 33.4, and the discussion on pp. 760–765). Given the variety of agents of disturbance that can affect any system, and the complexity of their interactions, it is often difﬁcult to predict with reasonable conﬁdence the severity of effects. Disturbance regimes, by their very nature, inject heterogeneity to the landscape at a variety of spatial and temporal scales. Ecological succession superimposes additional heterogeneity on the system, as disturbed patches change in biotic and abiotic characteristics over time. Early views of disturbance (White and Pickett, 1985) emphasized its discrete nature in time and space, as well as the production of patches with discrete boundaries (e.g., a landslide). That view has evolved and gained considerable sophistication during the past decade (Pickett et al., 1994; see also Pickett et al., Chapter 31), and now recognizes that disturbances may be characterized along a gradient from highly discrete (e.g., a tornado) to diffuse (e.g., a drought). The characterization of a disturbance along the gradient from discrete to diffuse is predicated on the particular scale at which heterogeneity is manifested in the landscape. To the extent that a disturbance alters heterogeneity at a particular scale, without creating discrete gaps at that scale, it is diffuse in nature. Moreover, a disturbance must be deﬁned within the context of the system of interest (Pickett et al., Chapter 31). Explicit delineation of a preliminary model of that system (structural components and their interactions) aids in distinguishing disturbances from other ecological “forcing functions”. Nonetheless, clear and unambiguous deﬁnition of a system model is not associated frequently with ecological studies of disturbance, and may at times contribute to controversy (e.g., compare Schowalter and Lowman, Chapter 9, with Willig and McGinley, Chapter 27). Although many of the chapters in this book address

Michael R. WILLIG and Lawrence R. WALKER

issues of patch-formation and spatial heterogeneity (Fig. 33.1), few provide a comprehensive description of patch dynamics for any particular system or an explicit conceptual model. Indeed, such coverage is beyond the scope of each chapter and likely is unavailable for many if not most ecological systems. Broad landscapelevel heterogeneity in abiotic and biotic conditions is a consequence of some disturbances, as illustrated by the following examples. Volcanoes create a variety of sterile patches (lava, pyroclastic ﬂows, and most lahars) on which primary successions proceed, as well as patches containing some biotic elements (thin tephra and some lahars) on which secondary successions proceed (del Moral and Grishin, Chapter 5). Largescale differences in the frequency and intensity of ﬁres exist among boreal regions, which in part affect differences in species composition and life history characteristics (Engelmark, Chapter 6). For example, forests with intense ﬁres often have patches comprising even-aged stands of the dominant trees, whereas in areas with less intense ﬁres the stands of dominant trees are more mixed in age structure. Winds create gaps of various sizes both in temperate forests (Webb, Chapter 7; Binkley, Chapter 18) and in tropical forests (Whigham et al., Chapter 8; Hartshorn and Whitmore, Chapter 19), as well as in polar regions (Kom´arkov´a and Wielgolaski, Chapter 3), and in wetlands (McKee and Baldwin, Chapter 13), creating a mosaic-like structure across the landscape. In general, the size of the patch and the grain of the organismal unit (sensu Kolasa and Rollo, 1991) affect the severity of response, whereas the size and the dispersion of the patches affect predisposition to subsequent disturbance. Reiterating the major premises of Schowalter and Lowman (Chapter 9) and Willig and McGinley (Chapter 27), many of the chapters focusing on a particular biome identify animals as important agents of disturbance through patch-generating activities, sometimes accelerating and at other times retarding rates of succession. In boreal ecosystems (Engelmark, Chapter 6), beavers (Castor spp.) are keystone species, creating discrete patches with regard to canopy structure, the distribution of woody debris on the forest ﬂoor, and the distribution and extent of wetlands. Reindeer (Rangifer) and moose (Alces) have more diffuse effects through their selective foraging behavior. Insect outbreaks are also important in boreal forests. Relatively regular outbreaks of spruce budworm (Choristoneura fumiferana (spruce budworm)) in balsam ﬁr (Abies balsamea) forests have a dramatic effect

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on the structure of the canopy; mild infestations of budworms favor ﬁr, whereas intense infestations favor spruce (Picea) (Baskerville, 1975; Holling, 1992; Morin, 1994). In contrast, outbreaks of spruce beetle (Dendroctonus ruﬁpennis) are episodic and cause large canopy gaps in forests of the Rocky Mountains of the western United States (Veblen et al., 1989). In temperate forests, overgrazing by deer (Odocoileus virginianus) causes considerable loss of understory plant species and alteration of vertical structure, which in turn, modify the response of the system to other disturbances (Webb, Chapter 7). Insects may also indirectly initiate diffuse disturbances by acting as vectors of plant pathogens such as “blue stain” fungus (Leptographium engelmannii, carried by the bark beetle D. ruﬁpennis) and exotic Dutch elm disease (caused by Ceratocystis ulmi, carried by the exotic bark beetles Scolytus multistriatus) in temperate regions of North America (Binkley, Chapter 18). Large patches denuded of vegetation (“eat-outs”) are created by snow geese (Anser caerulescens) and nutria (Myocaster coypus) in North American marshes, whereas a variety of animals, including invertebrates as well as alligators (Alligator), muskrats (Ondatra), beavers (Castor), seals (species of Pinnipedia), bison (Bison), and elephants (Elephas and Loxodonta), create patches by grubbing, burrowing, nest building, and trampling (McKee and Baldwin, Chapter 13). Succession and competition Succession and disturbance are intertwined concepts; a study of one must inevitably consider the other. Disturbance initiates succession, inﬂuences its subsequent trajectory, and can determine its rate, endpoint, and duration through subsequent intervention. Although succession is not inevitable, it is likely to occur following most disturbances, and disturbance effects often are measured by their inﬂuence on succession. The manipulation of succession (i.e., management) can reduce the severity of disturbance and contribute to restoration. Most chapters in this volume address succession (Fig. 33.1), suggesting its importance in understanding disturbance; this is particularly true for restoration following human activities. Moreover, ten chapters address competition as it is inﬂuenced by disturbance. The interactions of competition with disturbance and succession will be summarized brieﬂy, more extensive

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coverage of competition being found in the contributions by McKee and Baldwin (Chapter 13) and Wilson (Chapter 26). Each disturbance type puts a unique imprint on the landscape, which initiates a particular successional sequence. Attempts at generalization about both disturbance effects and subsequent successional trajectories are difﬁcult but useful exercises. Generalizations are useful because patterns emerge at broad spatial and temporal scales (Walker, Chapter 25). Succession generally is considered to be reset by disturbance (McKee, Chapter 13). However, disturbance also can accelerate succession by enhancing the dispersal of propagules, reducing competitive intensity (Wilson, Chapter 26), creating safe sites, or providing protection from grazers (Matthews, Chapter 2). On the other hand, disturbance can keep succession at a particular stage (Matthews, Chapter 2). The intensity of a disturbance, and the shape of the patch that it creates, determine which species survive the initial impact, what propagules remain viable in the soil or on plants to colonize the disturbed site, and how far exogenous propagules must travel to reach the site of disturbance (Whigham et al., Chapter 8). Substrate characteristics affect the initial supply of nutrients, water, seeds, and spores. Microbes are not only central to the success of plant colonization through mycorrhizae and nutrient mobilization, but also undergo successional sequences of their own (Allen et al., Chapter 22). The pre-disturbance biota often inﬂuences the pattern of the disturbance (cf. Whigham et al., Chapter 8). Patchiness (Pickett et al., Chapter 31), and spatial and temporal heterogeneity, have a pivotal impact on all interactions between disturbance and succession (e.g., dispersal, colonization, nutrient supply, microbial activity, and susceptibility to subsequent disturbance). Animals interact with plants in the post-disturbance environment in numerous ways and often undergo successional changes as well (Walker, Chapter 25; Willig and McGinley, Chapter 27). Animals affect plant pollination, fecundity, dispersal, productivity, health, competitive balances, and, ultimately, plant species composition (Crawley, 1997). Animals (including insect herbivores) can retard succession or accelerate it (Matthews, Chapter 2; Engelmark, Chapter 6; Schowalter and Lowman, Chapter 9) through their preferences for mid- or early-successional plant species, respectively. Plant–plant interactions (e.g., Cooke, Chapter 14;

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Wilson, Chapter 26) may be categorized as negative (competition) or positive (facilitation), even though they actually represent a continuum of possible interactions along a competition–facilitation gradient (Callaway and Walker, 1997). Competition is capable of altering the rate of succession, whereas facilitation can alter its rate or trajectory. Disturbance subsequent to an initial disturbance (termed an “inﬂux variable” by Matthews, Chapter 2) can determine the relative importance of competitive and facultative interactions through its potential determination of both proportions and densities of plant species (e.g., tall trees may be preferentially damaged during a windstorm). The ultimate convergence of vegetational composition following spatially separated or ﬂoristically unique seres may be guided by non-disruptive, but ubiquitous, disturbances (Matthews, Chapter 2). Alternatively, disruptive disturbances, stochastic dispersal processes, or differential species colonization in early succession may promote divergence (Matthews, Chapter 2; del Moral and Grishin, Chapter 5; Webb, Chapter 7). Recovery times for various pre-disturbance conditions differ widely. Soil nitrogen can take more than 1000 years to recover following some primary successions (Walker, Chapter 25). In contrast, population- or community-level characteristics of soil faunas (Allen et al., Chapter 22), some soil nutrient pools (Johnson and Schultz, Chapter 23), and plant canopies (Whigham et al., Chapter 8) under favorable conditions of secondary succession may take less than ﬁve years to recover. Forest species composition (Binkley, Chapter 18), ﬁne root biomass (Silver et al., 1996), and woody litter-fall typically recover within 50–100 years (Zimmerman et al., 1996). Only multidisciplinary and long-term studies of disturbance can provide the necessary understanding of the mode and tempo of responses by biotic and abiotic characteristics during secondary succession (see ﬁg. 1 in Zimmerman et al., 1996). The study of succession has been more formalized than the study of disturbance during the past 100 years, despite the initial linkage of the two concepts by Clements (1916) and other early ecologists. Disturbance theory is developing rapidly, particularly as the roles of humans (landscape ecology) and spatial heterogeneity (patch dynamics: Pickett et al., Chapter 31) are examined. Generalizations about disturbance must incorporate successional concepts and vice versa. As demonstrated by the geoecological model of Matthews

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(1992), disturbance mediates ecological processes, initially as well as throughout the successional process. Nutrient cycling and primary productivity Disturbance is a major factor controlling nutrient cycling and primary productivity (Lodge et al., 1991, 1994; see also D’Antonio et al., Chapter 17). These ecosystem attributes, discussed in several chapters of this volume (Fig. 33.1), not only are affected strongly by disturbance, but have considerable inﬂuence on disturbance regimes themselves. Speciﬁc effects of and responses to disturbance depend on the timing and nature of the disturbance, as well as the particular aspect of nutrient cycling or primary productivity of interest. Despite this complexity, some generalizations are possible. To the extent that disturbances remove plant biomass, they generally initiate a typical pattern of responses: an immediate increase in available nutrients, decomposition, and light, as well as a decrease in intra- and inter-speciﬁc competition. These increases can be offset by subsequent decreases in soil organic matter, leachable and volatile nutrients, and density or diversity of soil organisms. Recovery times to pre-disturbance levels of nutrient supply and primary productivity differ widely, from years to centuries, depending on the climate, the nature of the disturbance, and the target biota. Effects of disturbance on secondary productivity (i.e., rate of production of heterotrophic biomass) have not been well studied (Majer, 1989; Schowalter and Lowman, Chapter 9; Willig and McGinley, Chapter 27), although limited descriptive data are available for some systems (see Garrison and Willig, 1996; McMahon, 1996; Pfeiffer, 1996; Reagan, 1996; Stewart and Woolbright, 1996; Waide, 1996; Willig and Gannon, 1996). Disturbance impacts and ecosystem responses It is important to distinguish between the intensity (e.g., wind speed) and severity (damage caused) of a disturbance, and to delineate carefully the immediate, short-, mid-, and long-term responses of the abiotic and biotic aspects of the system of interest (Willig and McGinley, Chapter 27). Disturbance can continue to affect ecosystem processes long after the initial event (e.g., Matthews, Chapter 2). Similarly, the postdisturbance ecosystem interacts with and may itself alter aspects of the system. Characterization of responses to disturbance should include explicit spatial and temporal scales (Pickett et al., Chapter 31). Hereafter,

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we summarize the initial impacts of disturbance on nutrient cycling and primary productivity, as well as the immediate, short-term, and mid-term responses to disturbance. Subsequently, we discuss long-term responses and ecosystem recovery. Impacts, and short- to mid-term responses: Terrestrial disturbances generally involve some type of soil degradation (Barrow, 1991; see also Sojka, Chapter 21) including compaction, burning, loss of organic matter, acidiﬁcation or alkalization, ﬂooding, salinization, and addition of toxins, herbicides, or agrochemicals. In the most severe disturbances, soils are entirely lost through burial (e.g., under lava, ice, pavement), removal (blasting, mining), or burning. Erosion is a pervasive source of reduced productivity through decreased water storage (increased runoff) and associated losses of nutrients and organic matter. Pimentel and Harvey (Chapter 4) note that, during the last 40 years, 30% of the world’s arable land has become unproductive because of erosion. Wind erosion, with or without human inﬂuence, also can substantially reduce the capacity of soils to support plant growth. However, nutrient deposition from windborne particles also is a signiﬁcant input to many ecosystems (Goodall and Perry, 1981; Burrows, 1990; Walker, 1993; Perry, 1994; see also Sukopp and Starﬁnger, Chapter 16). Direct wind damage alters primary productivity and nutrient cycling in many ecosystems, but particularly in temperate, tropical, and high-elevation forests. The immediate consequences include loss of plant biomass, increases in light penetration to the forest ﬂoor, and increases in leaf litter and woody debris. Often, leaves have a higher content of nutrients than after natural leaf-fall, because no retranslocation has occurred (Lodge et al., 1991; see also McKee and Baldwin, Chapter 13). Such litter inputs generally increase the availability of nutrients (particularly nitrogen) as a result of enhanced leaching and decomposition (Whigham et al., Chapter 8). Microbial processes (including denitriﬁcation and soil respiration) are altered as well, with above- and belowground spatial heterogeneity increased through the formation of canopy gaps and the uprooting of trees (Webb, Chapter 7). The likelihood of wind damage is difﬁcult to predict, and the distribution of particular ecosystem responses is difﬁcult to generalize – they depend on the species composition of the forest (Webb, Chapter 7). In many ecosystems, humans have

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altered species composition, but even forest harvesting has the same general impact on nutrient cycling and productivity: nutrient availability increases from reduced plant uptake and increased decomposition of soil organic matter (Binkley, Chapter 18). The general pattern of increased resource availability in forests following wind disturbance is not universal (e.g., litter may not decompose) and repeated disturbances often lead to overall decreases in nutrient availability and primary productivity. Fire can be a natural or anthropogenic disturbance. It affects most vegetation types, especially grasslands, shrublands, and forests. Initial impacts of a ﬁre include destruction of plant biomass, volatilization of nutrients, and deposition of nutrient-rich, but nitrogen-poor, ash on the surface of the ground. Abiotic responses that follow include erosion, leaching, and cation adsorption. Biotic responses include decomposition of soil organic matter and initial increases in primary productivity associated with increased nutrient uptake or ﬁxation by plants (Bradbury, Chapter 24). Loss of soil organic matter from the ecosystem can inhibit or stimulate productivity, relative to pre-disturbance conditions (Engelmark, Chapter 6). Grazing is another disturbance of both natural and anthropogenic origin, which can decrease nutrient availability (through changes in soil pH, increased leaching, changes in litter quantity and quality), and primary productivity (through foliage removal, alteration of life forms, mortality; Schowalter and Lowman, Chapter 9). However, nutrient availability also can increase following grazing (e.g., less litter to inhibit nitrogen-ﬁxing organisms, or immobilize nutrients; McKee and Baldwin, Chapter 13), and compensatory growth can lead to net increases in primary production in fertile environments (Schowalter, Chapter 9; Oesterheld et al., Chapter 11). Insect outbreaks are a special case of herbivory which can dramatically reduce plant biomass, and alter nutrient cycling and microclimatic conditions over large areas. Grazing and other agricultural practices can make trees more susceptible to herbivore outbreaks (Schowalter and Lowman, Chapter 9). The interactions of grazing, ﬁre, and precipitation are modeled by Oesterheld et al. (Chapter 11). They suggest that primary productivity, essentially driven by precipitation, is most affected by ﬁre at high levels of precipitation (more current-year standing litter that is of low quality for herbivores). Further, the effects of grazing on primary productivity presumably remain constant (positive or negative)

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for all levels of precipitation, because as primary productivity increases with precipitation, the proportion of plant biomass eaten by grazers increases also. The net effects of grazing on nutrient cycling and productivity are therefore dependent on numerous interacting variables, including the intensity of the grazing itself. Flooding can have the same general effects as wind, ﬁre, and grazing (increased nutrient availability; increased decomposition and detritivore populations; decreased plant biomass and plant uptake of nutrients, followed by re-establishment of plants; and decreased nutrient availability). However, much depends on the duration and timing of inundation (McKee and Baldwin, Chapter 13). Disturbances typical of wetlands include wave action, ice formation, saltwater intrusion, ﬂooding, burial by wrack, dredging, ﬁshing, boating, and water-borne pollutants. Animal activities (e.g., trampling, mound-building, defecation) constitute a suite of disturbances that tend to be widespread but not severe (Willig and McGinley, Chapter 27). Losses in productivity, and increases in nutrient cycling and decomposition, may be small but cumulative. Animal activities generally remove plant biomass, thereby opening space and redistributing resources for use by colonizing organisms. Earthworm invasions can decrease or increase soil turnover and nutrient availability, depending on the type of earthworm and pre-invasion conditions (D’Antonio et al., Chapter 17). This again indicates an essential requirement in assessing any disturbance effect or response: a careful deﬁnition of what speciﬁc preand post-disturbance ecosystems are being compared. Animals are especially important in increasing local spatial heterogeneity of nutrient supply and associated primary productivity. Urbanization represents an extreme type of disturbance, with a unique impact on nutrient cycles and primary productivity (Sukopp and Starﬁnger, Chapter 16). Nutrient availability, microbial activity, and primary productivity can be reduced through leaching in acidic soils, soil compaction, and toxins or heavy metals in land-ﬁll soils. Nutrient levels and productivity can increase as a consequence of aeration from the addition of rubble soils or the addition of nutrients in dust or fertilizers. Unlike most disturbances, where increases in detritivore populations result in increased decomposition, the severity of urban disturbances (similar to severe ﬁres or primary succession) may severely reduce microbial biomass or

Michael R. WILLIG and Lawrence R. WALKER

diversity, resulting in reduced decomposition and litter accumulation. Long-term responses and restoration: Clearly, each disturbance elicits a characteristic suite of ecological responses within a particular community type, depending on the nature of the disturbance (intensity, frequency, extent) and the community (land-use history, as well as physiological and life-history attributes of species; Bradbury, Chapter 24). Zimmerman et al. (1996) evaluated several responses of the biota of a tropical rainforest in Puerto Rico to hurricane damage, and found at least six different response curves by various components of the biota over a ﬁve-year period. Some responses involved posthurricane increases (stream concentrations of nitrate, forest-ﬂoor biomass, primary productivity) followed by decreases back to pre-hurricane levels over a period of one to ﬁve years. The other responses (including aboveground potassium pools, tree biomass, litter-fall, and root biomass) involved initial decreases followed by various rates of return to pre-hurricane levels. This illustrates the difﬁculty in generalizing about responses to disturbance, and long-term responses in particular. Obviously, long-term responses for the microbiota may not represent the extended time periods that are applicable to long-lived trees; temporal scales must be relative to the taxon of interest (Willig and McGinley, Chapter 27; Pickett et al., Chapter 31). Nonetheless, generalizations about long-term responses include a gradual decrease in nutrient availability coupled with increased decomposition and plant growth. Primary productivity peaks and then declines as longer-lived plants sequester nutrients. Changes in species composition affect the disturbance regime, and plants affect soil development through stabilization, break-up of the substrate, transfer of nutrients from the subsoil, rhizosphere exudates, litter addition, and as hosts for mycorrhizae and nitrogen-ﬁxing organisms (Walker, Chapter 25). Spatial heterogeneity in nutrient availability and primary productivity generally increases with disturbance (e.g., formation of canopy gaps, unburned patches), but some disturbances (e.g., plowing) can decrease spatial heterogeneity. Stability and resilience Ecosystem stability can be deﬁned as the constancy of a parameter that characterizes community- or ecosystemlevel attributes through time (e.g., species composition,

SALIENT THEMES, SYNTHESIS, AND FUTURE DIRECTIONS

nitrogen mineralization). As such, it includes resistance to disturbance, as well as the ability to recover from a disturbance. Resilience is speciﬁcally the likelihood that parameters of a community or ecosystem will recover following disturbance. Ecosystems can be easy to disturb (unstable) or not (resistant), and may easily recover (be resilient) or not (Barrow, Chapter 28). Stability is thus the antithesis of disturbance, and resiliency characterizes the process of recovery from disturbance. Both concepts evolved from a steady-state view of ecosystems (Odum, 1971; see also Giampietro, Chapter 32), wherein disturbances were aberrations from which ecosystems recovered, to resemble eventually a pre-disturbance condition. Despite the fact that disturbance is an integral part of ecosystem dynamics (Cooper, 1926; Sousa, 1984b), both terms are still useful – stability as a contrast to disturbance, and resiliency as an indication of the potential for recovery that can guide restoration (Hobbs, Chapter 29). A further evolution of the concept of stability is embodied in meta-stability, the condition in which patches of a landscape may change, but all types of patches (or species, or processes) are still represented in the larger spatial and temporal context of the landscape. This shifting mosaic suggests that an ecosystem does not either resist or become totally altered by a disturbance, but is partially modiﬁed by it at various scales (Pickett et al., Chapter 31). Other chapters of this book also address the concept of stability or resilience (Table 1.2, p. 14). Disturbance does not necessarily prevent stability. Disturbance can maintain populations of disturbanceadapted or colonizing species. Disturbance also can maintain later successional stages, as in old-growth forests in southeastern Alaska (U.S.A.) by recycling nutrients through wind-throws (Matthews, Chapter 2) or through a wave-like regeneration pattern of ﬁr forests in New England (U.S.A.) or Japan (Engelmark, Chapter 6). Later successional stages often are considered to be more stable than are early ones, in part, perhaps, because stress-tolerant species are more common in later stages. Ghersa and Le´on (Chapter 20) note the difﬁculty of eradicating stress-tolerant weeds in agricultural systems. In this case, annual disturbance by grazers or cropping promotes stability of species composition. Resilience also can be inﬂuenced by past or current land-use and disturbance history (Kom´arkov´a and Wielgolaski, Chapter 3; Hobbs, Chapter 29). Eckert and Carroll (Chapter 30) deﬁne resilience as the range of variation that ecosystem processes

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and patterns can undergo before the system changes into a new system. Thus, patterns of ﬂuctuations and extreme conditions in a disturbance regime become critical. Because most management practices involve manipulation of the disturbance status of an ecosystem (Pickett et al., Chapter 31), land-use policies should prevent human exploitation from exceeding the limits of resiliency of an ecosystem (Eckert and Carroll, Chapter 30). Similarly, Barrow (Chapter 28) suggests that local and regional management of ecosystems gives the best chance of recovery and sustainability of ecosystem functions, because there is maximal overlap between patches of the regional mosaic to ensure survival of some if others are destroyed. Yet, increasingly, disturbances are global in nature, and international cooperation is essential for arriving at effective scientiﬁc and societal solutions (Barrow, Chapter 28). Predictability Disturbance at most temporal and spatial scales is unpredictable (consider, for example, human attempts at weather prediction). This large stochastic component may please theoretical ecologists, but makes management of disturbance effects difﬁcult. Humans approach this dilemma in two ways: through evaluation of historical disturbances to ﬁnd patterns or explanations, and through extrapolations of models or short-term patterns into the future. Some of these approaches have been addressed elsewhere in this volume. Historical reconstruction of disturbances ranges in scale from looking for meteorite impacts to explain Permian extinctions, through paleontological examination of pollen records, to evaluation of windstorm impacts during the last century. Catastrophic disturbances often leave a legacy, such as carbon residues from major ﬁres. Windstorm impacts can be analyzed by examining historical land surveys of blow-downs, residual effects of mounds on soil proﬁles, tree rings, presence of shadetolerant or shade-intolerant tree species, meteorological models, and aerial photographs (Webb, Chapter 7). Historical records of agricultural practices can indicate changes in plant species composition (Ghersa and Le´on, Chapter 20). Predicting the future disturbance regime is especially difﬁcult (Lugo and Scatena, 1996; see also Whigham et al., Chapter 8), even if one has determined historical frequencies and return intervals of a disturbance, and has distinguished that disturbance type (e.g., hurricane)

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from the background disturbance regime including all other causes of mortality (e.g., tree-falls from senescence or disease). Successional pathways often are predicted in advance, and some general patterns (e.g., increasing biomass, decreasing allogenic inﬂuences) appear to be robust. However, closer examination of speciﬁc seres often indicates little congruence between reality and prediction. The rate of change often is predicted more easily than is the speciﬁc trajectory of species change. This could be due to variability in the potential suite of colonizers at a given site and the random success of those colonizers. Also, similar habitats can support different species (del Moral and Grishin, Chapter 5). Disturbance may have a non-directional or directional inﬂuence on succession. Matthews (Chapter 2) suggests that disturbance alters the direction of succession under several distinct conditions. First, he proposes that disturbance early in succession can cause sites of similar age to diverge (e.g., in species composition) as a consequence of differential response to a variable disturbance regime. These sites may later converge to a similar species composition as a consequence of biotic factors. Alternatively, Matthews (Chapter 2) suggests that a low-level, homogeneous, and widespread disturbance that does not alter species composition can cause convergence of sites of similar age. The ultimate test of ability to predict the consequences of disturbance involves estimating human impacts on the biosphere (Giampietro, Chapter 32). Human economies, lifestyles, indeed survival, depend on proper forecasting, balancing trade-offs, and effective national and international leadership. Thresholds Biological thresholds can be deﬁned as the minimum level of stimulus needed to elicit a response. In this sense, disturbance needs to reach a given intensity or frequency to alter some ecosystem parameters. The alteration can come abruptly or gradually (Pickett et al., Chapter 31). Given the relative absence of simple linear responses in ecology and the prevalence of complex interacting factors and indirect responses (Barrow, Chapter 28; Hobbs, Chapter 29; Eckert and Carroll, Chapter 30), ecological thresholds are more likely to be gradual than abrupt. The importance of thresholds in the processes of disturbance and response is reﬂected in the chapters of this book addressing this concept (Fig. 33.1).

Michael R. WILLIG and Lawrence R. WALKER

Thresholds along a precipitation gradient determine the relative importance of three variables for aboveground net primary production in Argentina. Climate predominates where annual precipitation ranges between 200 and 450 mm; grazing and ﬁre balance each other where annual precipitation ranges between 450 and 700 mm; and ﬁre predominates where annual precipitation exceeds 700 mm (Oesterheld et al., Chapter 11). A threshold of irreversibility is a condition from which an ecosystem is no longer capable of recovery without intervention and restoration. This threshold is illustrated as the amplitude or range of a system parameter. Hartshorn and Whitmore (Chapter 19) suggest that isolated forest remnants reach a threshold for loss of biodiversity or ecosystem function beyond which they are both susceptible to invasion and difﬁcult to restore. Finally, Schowalter and Lowman (Chapter 9), use the concept of threshold differently. They suggest that, at some undeﬁned level of intensity or frequency of disturbance, insect outbreaks go from being background trophic interactions to become a disturbance (see also Willig and McGinley, Chapter 27). Biodiversity and functional redundancy Biological communities exhibit emergent properties, distinct from those at the level of populations or individuals. Among such emergent attributes is biodiversity. Biodiversity comprises three interrelated components: genetic diversity, taxonomic diversity, and functional diversity (Solbrig, 1991), and considerable controversy surrounds the contrast between functionally redundant species and those taxa which perform keystone services in maintaining ecosystem integrity (Jones and Lawton, 1995). Nonetheless, most disturbance studies addressing issues of biodiversity have been restricted to considerations of species diversity (for an exception, see Willig et al., 1996, who consider changes in functional diversity as a consequence of natural and anthropogenic disturbance). Although species diversity is an index which integrates information concerning the number of species (richness component), as well as their proportional abundances (evenness component), discussion of community-level issues in disturbance ecology most frequently is restricted to species richness. In part, this occurs because species richness is a conveniently calculated index of community organization which may be affected by various attributes of a disturbance regime. Indeed, both spatial variation in the compartmentalization of

SALIENT THEMES, SYNTHESIS, AND FUTURE DIRECTIONS

diversity (landscape heterogeneity and patch dynamics) and temporal variation in diversity (ecological succession) are consequences of disturbance. On the other hand, biodiversity is thought by some to enhance ecosystem stability and resilience, and thus moderate the effects of disturbance on various properties of ecological systems. A consideration of differences in species composition may also reﬂect functional aspects of biodiversity. Themes related to biodiversity ﬁgure prominently in a number of the chapters in this compendium, whereas consideration of functional redundancy is infrequent and cursory (Fig. 33.1). The effects of disturbance on species richness are variable and driven by complex and context-dependent mechanisms. Some disturbances (e.g., erosion and volcanism) consistently reduce diversity, whereas other types of disturbance may enhance diversity, depending on spatial scale and intensity (Willig and McGinley, Chapter 27). Soil erosion consistently reduces biodiversity in both terrestrial and aquatic systems (Pimentel and Harvey, Chapter 4). This is especially true of agricultural systems, where the loss of soil organic matter and reduction in soil quality directly reduce productivity, and indirectly reduce diversity of soil animals and microbes. Because agriculture dominates nearly half of the earth’s terrestrial systems (United States Department of Agriculture, 1993), the effects of erosion on diversity are quite severe. Nonetheless, a variety of amendments (additions of organic matter or straw mulch) may restore (at least temporarily) an appreciable proportion of the soil diversity previously lost to agricultural development. The effects of erosion are not limited to the terrestrial system in which the disturbance occurs. The soil particles transported by wind and water as a result of erosional processes are frequently deposited in aquatic systems, leading to eutrophication and reductions in species diversity. Applications of biocides (herbicides, fungicides, rodenticides, insecticides) to agricultural crops as a means of enhancing yield have effects on non-target taxa as well. When such non-target taxa provide essential ecosystem functions, such as nitrogen-ﬁxation, pollination, or seed dispersal (Allen et al., Chapter 22; Johnson and Schultz, Chapter 23; Willig and McGinley, Chapter 27), the consequences can be far-reaching. For example, bee populations may be declining or extirpated in some agricultural areas, reducing the reproductive success of plant species dependent upon them for pollination and altering their genetic structure.

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Similarly, volcanoes consistently reduce biodiversity by decimating plant, animal, and microbial populations through production of lava, lahars, pyroclastic ﬂows, and debris ﬂows (del Moral and Grishin, Chapter 5). Plant communities that develop subsequently on volcanoes are non-equilibrial and support fewer species than do areas unaffected by volcanic eruptions. In addition, the plant communities associated with volcanoes are in disharmony, supporting eclectic combinations of taxa as a response to aleatory features of colonization dynamics. Wind-generated disturbances (e.g., hurricanes, tornadoes) can increase or decrease diversity, or have no effect at all (Webb, Chapter 7). A similar variety of disturbance effects was found in a survey of ecological research in North American wetlands by McKee and Baldwin (Chapter 13). At the level of the patch, the speciﬁc impact of a disturbance depends on the size and boundary characteristics of the patches that it creates, the degree to which it reconﬁgures the distribution of essential resources, and the pattern of mortality that it causes (e.g., differential mortality of shade-tolerant versus shade-intolerant species). At the level of the landscape, the creation of patches of different sizes and qualities, and the attendant secondary succession that follows disturbance, may enhance diversity regardless of within-patch effects. Many systems attain highest diversity at intermediate levels of disturbance (Connell, 1978). Shrublands and woodlands under Mediterranean climates attain highest taxonomic and structural diversity if subjected to moderate grazing such that open habitats are maintained with regular disturbance (Rundel, Chapter 10). This is true of both plant and animal communities. Similarly, vegetation attains highest species richness in coastal marshes with intermediate intensities of disturbance by grazing by mammals, deposits of wrack, or scouring (McKee and Baldwin, Chapter 13). Demarais et al. (Chapter 15) intimate that intermediate levels of disturbance from military training activities enhance plant species richness by re-setting successional stages to pre-climax seres. Among the most important ways whereby a disturbance regime affects biodiversity is through evolutionary mechanisms. That is, particular disturbance regimes act as selective agents with respect to lifehistory characteristics of microbes, plants, and animals. Indeed, disturbance may have a signiﬁcant impact on the ecological attributes of species that persist as part of regional species pools. Differences in species

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composition among areas differing in disturbance history are frequently a consequence of the adaptive nature of life-history characteristics of potential colonists. This is clear in areas subject to volcanic disturbances (del Moral and Grishin, Chapter 5), where hemicryptophytes and geophytes are favored in areas subjected to blast effects; tephra is particularly harmful to mosses, lichens, low herbs, and taxa incapable of rhizomatous growth. In tropical forests, the successful germination and persistence of many plants (e.g., Cecropia, Heliconia, Macaranga, Ochroma, and Trema) depend on the continual production of gaps where microclimatic conditions (light regime, temperature, and moisture) differ from those of the otherwise undisturbed forest matrix (Hartshorn and Whitmore, Chapter 19). In some cases, gaps are the prime regeneration sites for up to half of the species of native trees (Hartshorn, 1978). The pervasive distribution of naturally produced gaps in time and space may result in some tropical species of trees being “pre-adapted” to survive anthropogenic disturbances such as those produced by modest selective logging. For example, 3– 8 years after the cessation of logging in a tropical forest of Borneo, population densities of most bird species and avian guild structure returned to pre-disturbance conditions (Lambert, 1992). Organisms that could ﬂee from the initial impact of logging practices, or that could easily recolonize once logging ceased, showed few long-term changes as a consequence of disturbance. Similarly, life-history characteristics of many species of plants, animals, and microbes in the boreal forests of the New and Old Worlds are ﬁre adaptations (Engelmark, Chapter 6). Anthropogenic disturbance is perhaps the greatest threat to world-wide biodiversity. Lands directly modiﬁed for agricultural development (croplands, grazing lands, managed forests) and urbanization clearly experience severe in situ reductions in species richness (Pimentel and Harvey, Chapter 4; Cooke, Chapter 14; Sukopp and Starﬁnger, Chapter 16; Hartshorn and Whitmore, Chapter 19). Moreover, anthropogenic disturbance has a cosmopolitan distribution (see the back endpaper of this volume), and, because of the rate of increase of human populations, will likely continue to expand well into the next century (Barrow, Chapter 28; Giampietro, Chapter 32). Ironically, some areas dedicated to human uses may act as refuges for important components of the biota. Military lands, although subjected to training exercises, may be considerably less developed or modiﬁed than adjacent areas used

Michael R. WILLIG and Lawrence R. WALKER

for agriculture, industry, or cities (Demarais et al., Chapter 15). As a consequence, such lands may harbor relatively unmodiﬁed natural communities as well as endangered species of plants and animals. Similarly, land-ﬁlls located in areas of high urbanization (Sukopp and Starﬁnger, Chapter 16) can be rich in species, and include populations of taxa that are otherwise endangered or threatened by other aspects of human encroachment. Nonetheless, improved protection of biological diversity will require concerted efforts to preserve ecological systems rather than the protection of single species or endangered taxa (Eckert and Carroll, Chapter 30). Invasive species Because disturbance removes biomass and opens space, colonization by neighboring or newly introduced species is inevitable, except in the harshest conditions. The process of colonization is dependent on many variables including climate, microsite availability, dispersal distances and capabilities, substrate conditions, residual soil or organisms, secondary disturbances, and stochastic events (del Moral and Grishin, Chapter 5). In this section, we focus on the interaction of a particular subset of colonizers, the invasive (nonindigenous, alien, or exotic) species, and summarize conclusions from chapters of this book addressing the interaction between invasive species and disturbance (Fig. 33.1). Recent books that provide an overview of the topic include those of Cronk and Fuller (1995), Pyˇsek et al. (1995), Luken and Thieret (1997) and Shigesada and Kawasaki (1997). Although invasive species of plants are found in areas where natural disturbances have reduced plant cover, they are particularly associated with anthropogenic disturbances. In both cases, their introduction is usually from human activities. For example, alien grasses, herbs, shrubs, and trees dominate volcanic disturbances in New Zealand, Hawaii, and Japan, and these invasive species are often the product of purposeful human introductions (e.g., to control erosion; del Moral and Grishin, Chapter 5). Invasive species in a disturbed area are often from similar but distant environments. Most grasses that are prevalent as invaders of habitats in Mediterranean climates around the world originated in Europe (Rundel, Chapter 10). Such habitats may be particularly susceptible to invasion because of low vegetative cover during summer droughts, and a long history of species adaptation to human

SALIENT THEMES, SYNTHESIS, AND FUTURE DIRECTIONS

intervention in Europe (ﬁre, grazing, deforestation: Rundel, Chapter 10). Interestingly, native grasses still dominate on low-nutrient serpentine soils in California (Rundel, Chapter 10), and are reinvading old, nutrientpoor agricultural systems in Argentina (Ghersa and Le´on, Chapter 20), suggesting that habitats of high productivity favor invasion. Urban environments also are dominated by weeds, in part because cities are centers of transportation and provide warmer microsites for weeds from more southerly climates (Sukopp and Starﬁnger, Chapter 16). Disturbance generally promotes invasion of nonnative species, particularly when the disturbance is severe or persistent (D’Antonio et al., Chapter 17). Good invaders tend to be ruderals with high fecundity, well-dispersed seeds, rapid growth, low root/shoot ratios, large seed banks, and ﬂexible life histories (Rundel, Chapter 10). However, invasions also occur without disturbance, and invasive species often colonize undisturbed areas surrounding a disturbance (D’Antonio et al., Chapter 17). The impacts of invaders are complex. They can outcompete native vegetation, alter successional pathways, inﬂuence ecosystem parameters, and change disturbance regimes. For example, Argentine ants (Solenopsis and related genera) hinder dispersal of native seeds by indigenous ants in South Africa, exotic fungal pathogens kill plants in the family Proteaceae in Australia, and invasive grasses promote ﬁre in the arid southwestern United States (Rundel, Chapter 10). In many cases, humans intentionally introduce nonnative plants to restore ecosystem function to a badly damaged site rapidly, because the non-natives grow more quickly. However, the risk is that introduced species may inhibit establishment of native vegetation (Hartshorn and Whitmore, Chapter 19; Hobbs, Chapter 29). Widespread human disturbances have promoted a global mixing of species and sometimes a reduction in species diversity. Understanding the dynamics of invasions will aid humans to incorporate exotics properly into management and restoration plans. Restoration and management Restoration and management are human attempts to alleviate negative impacts of disturbance. As such, they represent the optimistic view that human intervention can restore damaged ecosystems faster than the ecosystems are being destroyed. Earlier in human history, relocation was an option (Barrow, Chapter 28), but the

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effects of human impact are too severe and extensive to avoid any longer. Many chapters address the issues of restoration, management, and the related topic of risk assessment (Fig. 33.1). We use the term “restoration” in the broad sense for all efforts to repair ecosystems, not in the narrow sense of return to the original system (Cooke, Chapter 14; Hobbs, Chapter 29). The best management option is to avoid degradation or to prevent it from crossing irreversible thresholds, after which restoration efforts are needed (Hobbs, Chapter 29). For example, Demarais et al. (Chapter 15) document ways in which military training activities may be modiﬁed to reduce environmental damage, and Hartshorn and Whitmore (Chapter 19) note that stripcutting is better than clear-cutting for regeneration of tropical forests. With respect of endangered species, stopping habitat destruction is essential before the species is lost (Bowles and Whelan, 1994). Risk assessment attempts to predict where early intervention can reduce site degradation. Predicting which trees are at risk from wind-throw can help foresters to design planting densities and thinning schedules so as to minimize the loss of mature trees (Webb, Chapter 7). Certain types of hazard, both natural (frosts and avalanches) and anthropogenic (military training), are readily predictable, others (effects of global climate change) are not (Barrow, Chapter 28). Humans can attempt to minimize damage from predictable events (e.g., make ﬁre breaks or build according to earthquake codes), but too often governments require proof that a disturbance is real before action is taken (Barrow, Chapter 28). Such delays frequently make mitigation efforts more costly. Similar problems exist on a personal level if perception and probability of risk are not strongly coupled. People are much more worried about the (rare) earthquake or nuclear accident than the (very common) car accident. Environmental impact assessments are now commonly used to predict the effects of human activities (e.g., on wetlands, or the fate of endangered species). Evaluations of the success of such predictions are critical for both individuals and governments to prevent the negative impacts of disturbance, or to ameliorate conditions in postdisturbance scenarios. Restoration is deemed necessary for a wide variety of reasons, including the management of catchments or air quality, the preservation of species or ecosystem functions, acceleration of natural successional processes, growth of crops, removal of pollutants, and even the re-establishment of realistic conditions for military

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maneuvers (Demarais et al., Chapter 15). When the explicit goals of a restoration effort are stated clearly, the success of the project can be evaluated in an unambiguous fashion. In many cases, a return to a pre-disturbance condition or to a nearby, relatively undisturbed condition is desired, but this approach is challenging because so-called “natural” habitats are spatially heterogeneous and may also vary through time (Hobbs, Chapter 29). Increasingly, the re-establishment of certain vital ecosystem attributes (Aronson and Le Floc’h, 1996) is a criterion for success. Restoration techniques are of necessity site-speciﬁc, and vary from complete reconstruction of an ecosystem to minor manipulations of ecosystem attributes. Techniques include identiﬁcation and removal of stressors, and replacement of key ecosystem components lost in the disturbance (Hobbs, Chapter 29). Recovery of severely disturbed habitats, such as mine wastes, is impeded by soil acidity and presence of toxins (Cooke, Chapter 14). In the United States, rushing the process, for instance to meet certain regulations, ironically has resulted in delayed succession. For example, laws mandating a return to 90% of original site productivity within ﬁve years resulted in the planting of grasses and legumes or pine trees (Pinus strobus) which inhibited subsequent forest succession. This result is generalizable: without a good understanding of natural regenerative processes, human interference can slow rather than accelerate recovery. A less comprehensive restoration technique might simply be the scariﬁcation of the surface of mine wastes to increase soil moisture (Allen et al., Chapter 22). Disturbance itself is used to adjust successional trajectories toward a desired endpoint, as when competition from early-successional species is reduced so that later-successional trees can establish. Successful restoration depends on clear identiﬁcation of ecological endpoints with regard to system structure and function, as well as an understanding of natural succession and the successional dynamics at the disturbed site. Restoration also depends on the ability of management procedures to mimic, replace, or compensate for the multiple effects of a disturbance regime (Pickett et al., Chapter 31). Such understanding is rarely complete, so that trial and error will likely characterize restoration efforts. The strongly site-speciﬁc nature of restoration does not preclude some generalizations about the best procedures to utilize. The reassembly of a functioning ecosystem, whether it mimics the pre-disturbed system or not,

Michael R. WILLIG and Lawrence R. WALKER

is an acid test for ecologists (Jordan et al., 1987), and reliable guidelines are emerging (see Hobbs, Chapter 29, for references). However, the link between biological realities and environmental policy is still too tenuous (Wali, 1992). Past environmental policies have had both favorable consequences (establishment of national parks) and unfavorable ones (laws allowing uncontrolled mining or ranching or forest removal). One may hope for a future in which management of ecosystems emphasizes the necessity for land managers to recognize the value of natural components of ecosystems, as well as the importance of participation by those who will be affected by management decisions (Eckert and Carroll, Chapter 30).

SYNTHESIS AND FUTURE DIRECTIONS

As a ﬁrst step toward synthesis, we develop a conceptual model that explicitly links disturbance regime, successional dynamics, and spatial heterogeneity into a common framework. Any particular point on the earth’s surface occupies a portion of geographic space represented by its latitude, longitude, and elevation. In addition, that point occupies a position in ecological space (E) deﬁned by its location on any of a number of gradients representing environmental characteristics. Environmental characteristics may be biotic (e.g., the number of individuals of each of a set of species, species richness, functional diversity) or abiotic (e.g., light level, soil moisture, pH, temperature). Succession Disturbance (e.g., a hurricane) affects a region of geographic space by altering abiotic or biotic portions of ecological space (Fig. 33.3). For example, high winds generated by a hurricane can uproot and kill trees, thereby creating gaps in the canopy, with an attendant increase in light and temperature at the ground. The direct effects of the hurricane (D) cause a shift (E) in the ecological conditions of the point at which the tree fell. Such direct effects initiate a cascade of subsequent responses by the biotic (B) and abiotic (A) environments, which constitute secondary succession. For example, previously dormant seeds of light-tolerant species in the soil may germinate in the newly formed gap, changing the species composition of the site. As seedlings grow, they shade the litter (decrease soil temperature) and attract a suite of insect

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damaged another tree on the perimeter of the gap (Fig. 33.2). This may sufﬁciently weaken the tree at the perimeter so that it responds to the added wind turbulence at the gap’s edge during a subsequent tropical storm by the trunk snapping, producing a treefall and increasing the size of the gap. In summary, at any point in time, trajectories of succession depend upon (1) direct effects associated with current disturbances (D); (2) factors associated with previous disturbances, both abiotic (A) and biotic (B); and (3) the environment in which a site occurs (e.g., the pool of colonist species1 ). The change in ecological characteristics of a site can be quantiﬁed by the relationship: Fig. 33.3. Generalized conceptual model representing the manner in which a disturbance (D) affects a geographically explicit portion of a landscape (E), causing a change (DE) in its ecological characteristics (see text for details). A and B represent feedbacks from the current environment based on abiotic and biotic characteristics, respectively, that affect successional trajectories. Both A and B embody the legacy of previous disturbances. The current state of the ecosystem has feedbacks to the disturbance regime (M ) through alteration of the frequency, extent, or intensity of each disturbance element in the disturbance regime (Fig. 33.2). Arrows represent inputs or ﬂows; squares represent state spaces. Dashed arrows represent changes in ecological space wrought by previous disturbances (Di−1 , Ai−1 , Bi−1 ), whereas dashed squares represent past (left, E i−1 ) or future (right, E i+1 ) ecological space, which has characterized or will characterize the site as a result of succession.

herbivores, changing the faunal composition of the site. Hence, the site will follow a path in ecological space initiated by the disturbance and modulated by the surrounding environment (e.g., seed bank, species pool of insect herbivores). Of course, the occurrence of a disturbance at a site in the past does not necessarily mean that the site is immune to future disturbances. For example, two weeks after the fall of a tree, a six-week drought may occur which dries the litter and causes enhanced mortality of the snail fauna occupying the site. This second disturbance directly alters successional trajectories, and brings with it a suite of attendant indirect effects as well. Together, the legacy of changes initiated by the hurricane combine with those of the drought to produce a new trajectory of change in the biotic and abiotic characteristics of the site (Fig. 33.4). Finally, the occurrence of a disturbance can affect the likelihood and characteristics of future disturbances. The tree that fell during the hurricane may have 1

E = F(D, A, B | E), where E incorporates the effects of D, A, and B that are conditional on the current ecological state of the site (E). Any particular site is subject to a regime representing

Fig. 33.4. Ecological succession may be considered to be a consequence of the trajectory of changes (DE i ) that a site experiences over time as a result of the cumulative impact of the disturbance regime (see Fig. 33.2). Elements of the disturbance regime derived from earth, air, water, ﬁre, and the biota (Walker and Willig, Chapter 1) interact with each other, often in a complex fashion, and synergistically constitute a forcing function at any point in time. The likelihood of any of these elements of the disturbance regime impinging on a site, as well as their deﬁning characteristics, may have been affected by previous disturbances. Hence, trajectories of response at a site may be complex and difﬁcult to predict because of the multifaceted way in which past and current disturbances interact with each other in the context of the current environment to elicit changes in the biotic characteristics of a site.

This aspect is often referred to in the literature as a “mass effect” or “rescue effect”.

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all of the possible disturbances to which it could be subjected (Fig. 33.4). Hence, E = F(D1 , D2 , . . . Dn , A, B | E), where the subscripts of D represent each of the possible disturbance elements that constitute the disturbance regime of the area. Each of the possible disturbances (Di ), whether they are associated with earth (tectonic), air, water, ﬁre, or the biota, has associated with it a likelihood of occurrence (P, proportional to the frequency of occurrence and extent) and a probable intensity (I , based on a distribution of possible intensities). Hence,

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Di = P i I i . Clearly, quantitative prediction of trajectories of response to disturbance is difﬁcult because of the inherent complexity of the system. This sentiment echoes similar statements that characterize most of the preceding chapters of this book. The complexity likely is a consequence of the structure of the system (Figs. 33.3, 33.4), as well as the non-deterministic and non-linear dynamics characterizing many of the cause and effect relationships (Kolasa and Pickett, 1991; Jones and Lawton, 1995; Haefner, 1996). Nonetheless, modeling approaches based on Markovian and semiMarkovian perspectives hold promise, especially when they incorporate higher-order dynamics and history in predicting ecological trajectories (Henderson and Wilkins, 1975; Horn, 1975; Cohen, 1976; Usher, 1981, 1987; Caswell, 1989; Tanner et al., 1994, 1996) Patch dynamics Any site or patch occupies a geographically deﬁned position in a heterogeneous landscape, with various patches interacting with each other (Forman and Gordon, 1981; Forman, 1997). Just as a particular site can be represented by its position in ecological space, a suite of sites can be visualized in the same ecological space (Fig. 33.5). The proximity of such sites in ecological space represents their similarity in biotic or abiotic characteristics, but because of disturbance their relationships may differ over time, some moving in parallel, others converging, and still others diverging. General patterns of succession may be visualized in such a scenario (Fig. 33.5A) and provide insight to questions concerning the spatial organization of the ecological variability of a landscape subject to a disturbance regime.

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Fig. 33.5. Disturbance-initiated succession involves a sequence of changes in abiotic attributes (e.g., temperature, soil moisture, solar irradiation) and biotic attributes (e.g., species composition, species richness, species evenness, N mineralization rates) Any point in geographical space can be represented by a suite of ecological characteristics. Nonetheless, visualization in multidimensional space is difﬁcult. Consequently, data reduction techniques are used to produce a few important axes of variation (e.g., axes I, II and III), which are orthogonal combinations of the suite of ecological characteritics. (A) A disturbed point (open circle) undergoes succession whereby its ecological characteristics (position on axes I, II, and III) change over time, as represented by the color changes in the circles from open, through shades of gray, to black. Sequences of states over time representing trajectories of response to disturbance may be depicted by circles connected by arrows. (B) The tabonuco forest of Puerto Rico prior to Hurricane Hugo was visualized as an extensive forest matrix (solid circles) interrupted by a number of sites subjected to tree-falls (dark gray circles) and a few sites subjected to landslides (light gray circles). Each of the disturbed sites was believed to undergo secondary successional changes which eventually return sites to the general condition of the matrix (i.e., recovery). (C) The direct effects of Hurricane Hugo were to cause considerable damage to the forest, killing and damaging many trees and opening extensive portions of the canopy (gray circles). Only a few stands occupying protected sites based on topography and slope were unaffected by the hurricane (solid circles), reconﬁguring the forest so that it was a heterogeneous m´elange of disturbed sites interspersed with a few undisturbed sites.

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For example, the tabonuco forest of Puerto Rico was considered to be a mosaic of patches prior to the impact of Hurricane Hugo, with differences among sites related to the effects of disturbances such as tree-falls and landslides (Fig. 33.5B). The general perception was that most sites were relatively “undisturbed” and considered part of the forest matrix. Nonetheless, variation among sites in the matrix may be a consequence of a variety of attributes including slope, aspect, elevation, and stochastic events. In addition, some patches were recently subjected to tree-falls and considered to be moderately disturbed, whereas a few sites were recently subjected to landslides and considered to be severely disturbed. As a result, the “Swiss cheese” image of a forest based on geographic considerations (i.e., an extensive forest matrix interrupted by occasional small and larger gaps, tree-falls and landslides) can be visualized in ecological space as a core of numerous sites (matrix) surrounded by near (tree-fall) and distant (landslide) outliers (Fig. 33.5B). This same approach can be used to visualize the ecological variability of the tabonuco forest after the impact of a severe disturbance such as Hurricane Hugo (Walker et al., 1991, 1996). The direction and intensity of the hurricane, in conjunction with topographic features of the Luquillo Mountains, resulted in severe damage to most areas of the forest, although some sites were relatively undisturbed (Fig. 33.5C). As a result, most sites in the forest were highly, though variably, dispersed in ecological space, and only a few sites retained characteristics of the former matrix. In geographic space, the Swiss cheese was mostly holes. Agents of control during succession Temporal trajectories in ecological space followed by disturbed sites as a consequence of secondary succession (Fig. 33.5) can be examined within the framework of a conceptual model such as that outlined above (Fig. 33.6). For heuristic purposes, we consider the ecological trajectories associated with three of the possible control agents: stochastic events (Fig. 33.6A), current ecological characteristics (Fig. 33.6B), and physical characteristics (Fig. 33.6C). If physical characteristics of the site primarily guide succession, then sites with similar slope, aspect, or elevation should converge, regardless of current ecological characteristics (Fig. 33.6C). This same pattern of response would characterize controls associated with previous landuse history. In contrast, if ecological characteristics

763 A. Stochastic Events Control Trajectories

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Fig. 33.6. Sites within a landscape can be represented by their position in ecological space, as deﬁned by biotic or abiotic gradients. Trajectories of secondary succession may be controlled by a number of factors, including current ecological characteristics (represented by shape of points; squares versus circles), previous land use (shading of points; black versus gray), or stochastic processes. (A) If stochastic processes affect successional pathways, then neither previous landuse (e.g., shading of points) nor current ecological characteristics (shape of points) of sites should produce patterns in the trajectories of response to a disturbance. (B) If current ecological conditions (shape of points) primarily determine the path of succession, then sites sharing current ecological space should follow parallel trajectories of recovery. (C) If physical characteristics or previous land-use (shading of points) has the dominant role in affecting succession, then sites should converge in ecological space based on conditions related to previous land-use history.

of the immediate post-disturbance environment play the dominant role in channeling succession, then disturbed sites occupying similar positions in ecological space should follow parallel trajectories (Fig. 33.6B). Finally, if stochastic events or factors unknown to the investigator direct secondary succession, then trajectories should not follow either pattern (Fig. 33.6A). Of course, ecological systems are notoriously complex, and the challenge for the future is to determine the biological circumstances which may favor one sort

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of control agent over another (sensu the evolution or ontogeny of theory: Pickett et al., 1994). At least four tasks are inherent in that challenge. First, to determine if some types of disturbance are more associated with one type of control mechanism than another. Second, to evaluate if the relative importance of control mechanisms varies over time (e.g., stochastic control early in secondary succession, followed by physical controls later in secondary succession). Third, to discover if some organismal groups (e.g., herbs versus hardwood trees, bats versus rodents, microorganisms versus macro-organisms) or processes (e.g., N mineralization, P cycling, rates of herbivory) are more often regulated by one control mechanism than another. And fourth, to evaluate if considerations of ecological scale affect the detection of types of control mechanisms. Accomplishments Disturbance as a theory unites concepts in succession and landscape ecology. Like all theories, it has exhibited an ontogeny since its popular formalization (Bormann and Likens, 1979; Sousa 1984a,b; Pickett and White, 1985). Nonetheless, the theory is not yet mature and might best be characterized at a stage between the consolidating and empirical-interactive phases of development described by Pickett et al. (1994). In many ways, the chapters in this book substantially contribute to the maturation of disturbance theory. Almost all of them provide facts (conformable records of phenomena), deﬁnitions (conventions and prescriptions necessary to communicate clearly), or concepts (regularities in phenomena). Some (e.g., Schowalter and Lowman, Chapter 9; Willig and McGinley, Chapter 27; Pickett et al., Chapter 31) have clariﬁed the domain of the theory (the scope of the phenomena in space and time). Others have contributed new models (conceptual constructs that represent or simplify nature) or tested extant hypotheses (statements representing components of theory) related to disturbance (e.g., Rundel, Chapter 10; Oesterheld et al., Chapter 11; MacMahon, Chapter 12; Ghersa and Le´on, Chapter 19; Allen et al., Chapter 22; Walker, Chapter 25; Wilson, Chapter 26; Willig and McGinley, Chapter 27; Hobbs, Chapter 29; Eckert and Carroll, Chapter 30). The earlier part of this chapter has presented the many conﬁrmed generalizations (condensations and abstractions from

Michael R. WILLIG and Lawrence R. WALKER

a body of facts that have been tested) that appear throughout the volume. Future directions Regardless of location, the study of disturbance requires a long-term perspective (Magnuson, 1995), and is best conducted by multidisciplinary teams of scientists who simultaneously examine patterns and processes at a variety of spatial and temporal scales (Levin, 1995). Many of the chapters in this book conclude with a suite of recommendations for future research. Rather than repeat these recommendations here, we focus on overarching issues that are not biome-, taxon-, or process-speciﬁc. (1) The broad generalizations that currently characterize disturbance theory need to engender falsiﬁable hypotheses. The theory of disturbance needs to become more predictive, and the predictions need to be more quantitative. This is particularly critical in restoration and management scenarios. (2) Models such as those that appear in this chapter (Figs. 33.3, 33.4), as well as those that appear elsewhere in this volume, need to be incorporated into synthetic landscape models. More speciﬁcally, the interacting cells of the landscape model should represent the geographic space of an ecosystem, with explicit ﬂows occurring between adjacent cells in the landscape. (3) For any particular disturbance type, research should be designed to distinguish the effects of frequency, extent, and intensity on a suite of biotic and abiotic attributes. The recent work on the effects of the extent and pattern of ﬁres on secondary succession (Turner et al., 1997) serves as a worthy model. (4) The study of disturbed ground needs to become more comparative. This is particularly critical, given the importance of historical legacies and contingencies (Berlow, 1997). Site-speciﬁc response to disturbance must be distinguished from more pervasive or general results; replication is necessary to achieve this goal. Indeed, many of the insights presented in this volume derive from observations at multiple sites. Moreover, identiﬁcation of important environmental gradients along which to stratify ﬁeld observations or experiments may provide the best basis for comparative study. (5) Future research should assess the degree to which disturbance-mediated changes in above-ground structure and function correspond to those which occur

SALIENT THEMES, SYNTHESIS, AND FUTURE DIRECTIONS

below ground. The degree to which above-ground trajectories of recovery reﬂect or fail to reﬂect below-ground trajectories has not been sufﬁciently addressed in the contemporary literature. (6) The disciplines of land and resource management, range and wildlife management, risk assessment, and urban planning need to become more familiar with the theory of disturbance. Effective utilization, conservation, and protection of the earth’s resources require the development of intervention strategies that are harmonized with the natural disturbance regimes of target ecosystems and landscapes. (7) Theoreticians and ﬁeld biologists studying the ecology of disturbed ground need to consider more fully the application of their work to practical questions of management, conservation, and remediation. Indeed, all scientists have a responsibility to consider the needs of society and to contribute actively to public discourse on matters about which a scientiﬁc perspective is critical. The study of disturbed terrestrial ecosystems has provided a wealth of knowledge concerning ecological relationships. Understanding the dynamics of such ecosystems contributes to wise stewardship and may be critical to the future of the biosphere. Indeed, if scientists and managers are to provide effective guidance to decision-makers and politicians, then theories must become more robust and the empirical evidence on which they are constructed must be more broadly based. For both practical and theoretical reasons, the scientiﬁc community should marshal considerable effort in the future to advance the study of disturbance in terrestrial ecosystems.

EPILOGUE

Studying disturbance is difﬁcult. Understanding its consequences is even more challenging. To the scientist and environmental manager, historical legacies, scale-dependence, multiple causation, inter-correlation, temporal variation, and spatial heterogeneity conspire to make patterns complex and detection of underlying mechanisms elusive. We are reminded of an observation by Plato in The Republic, Book VII: “Picture men in an underground cave, with a long entrance reaching up towards the light along the whole width of the cave . . . such men would see nothing of themselves or of each other except the shadows thrown by the ﬁre on

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the wall of the cave . . . The only truth that such men would conceive would be the shadows.” Although we are limited by the data that we have collected and by the theories that we have developed, we are encouraged that, even at this early stage of understanding disturbance, the shadows reveal the substance of the phenomena to which this book is dedicated. Indeed, we are inspired by the content of this book, and hope that others will be similarly motivated to redouble efforts to study the ecology of disturbed ground and seek the light at the end of the tunnel.

ACKNOWLEDGMENTS

The ideas presented in this chapter were inspired by the various contributions to this book. In addition, our graduate students and faculty at our home institutions, as well as our colleagues in the Luquillo Longterm Ecological Research program, have contributed in numerous ways to the development of the chapter. Charles V. Cogbill, Mark M. McGinley, Daryl L. Moorhead, and Jess K. Zimmerman provided helpful comments on earlier versions of the manuscript. In part, this research was performed under grant BSR-8811902 from the National Science Foundation to the Institute for Tropical Ecosystem Studies, University of Puerto Rico and the International Institute for Tropical Forestry, as part of the long-term ecological research program in the Luquillo Experimental Forest. Additional support was provided by the Forest Service (U.S. Department of Agriculture), the University of Puerto Rico, Texas Tech University, and the University of Nevada.

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Wali, M.K. (Editor), 1992. Ecosystem Rehabilitation, Vol. 1, Policy Issues. SPB Academic Publishing, The Hague, The Netherlands, 230 pp. Walker, L.R., 1993. Nitrogen ﬁxers and species replacements in primary succession. In: J. Miles and D.W.H. Walton (Editors), Primary Succession on Land. Blackwell Scientiﬁc Publications, Oxford, England, pp. 249–272. Walker, L.R., Brokaw, N.V.L., Lodge, D.J. and Waide, R.B. (Editors), 1991. Ecosystem, plant, and animal responses to Hurricanes in the Caribbean. Biotropica, 23: 313–521. Walker, L.R., Silver, W.L., Willig, M.R. and Zimmerman, J.K. (Editors), 1996. Long term responses of Caribbean ecosystems to disturbance. Biotropica, 28: 414–614. White, P.S. and Pickett, S.T.A., 1985. Natural disturbance and patch dynamics: An introduction. In: S.T.A. Pickett and P.S. White (Editors), The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, San Diego, California, pp. 3– 13. Willig, M.R. and Gannon, M.R., 1996. Mammals. In: D.P. Reagan and R.B. Waide (Editors), The Food Web of a Tropical Forest. The University of Chicago Press, Chicago, Illinois, pp. 399–432. Willig, M.R., Moorhead, D.L., Cox, S.B. and Zak, J.C., 1996. Functional diversity of soil bacterial communities in the tabonuco forest: interaction of anthropogenic and natural disturbance. Biotropica, 28: 471–483. Zimmerman, J.K., Willig, M.R., Walker, L.R. and Silver, W.L., 1996. Introduction: disturbance and Caribbean ecosystems. Biotropica, 28: 414–423.

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GLOSSARY abiotic. Pertaining to non-biological factors (e.g., abiotic controls over plant establishment in succession might include wind, temperature, light, erosion) aeolian (= eolian). Wind-borne (e.g., aeolian processes build sand dunes, disperse plants) allogenic. See disturbance anthropogenic. Of human origin (e.g., mining is an anthropogenic disturbance; antonym: natural) arrested succession. Succession where species change is delayed due to dominance by one species autogenic. See disturbance biodiversity. The variety of living species and their genetic variation biological invasion. The movement of organisms into new territories; when the invasive species are not native to that area (alien, exotic), the invasion may have undesirable results; cf. feral, ruderal, weed biological legacy. Organisms that survive a disturbance and inﬂuence subsequent succession biome. A biogeographical region characterized by a particular life form biomass. The mass of all living organisms at a site, or of a particular group blue-green algae. See cyanobacteria bog. A wet, acidic environment with low decomposition rates boreal. The zone between the temperate and arctic zones dominated by the boreal forest (taiga), a vegetation type (biome) that is associated with long, cold winters, short summers during which temperatures exceed 10ºC, and low levels of precipitation carnivore. Meat-eater carrying capacity. The maximum number of organisms a particular environment can sustain (e.g., the number of humans the earth can support) chronosequence. A developmental sequence of soils and/or plant communities clear-cutting. The removal of all logs from a site (antonym: selective logging) compensatory growth. A positive growth response from grazed plants (where the rate of loss of plant mass due to grazing is less than the rate of consumption by the grazer)

competition. An interaction between two organisms, generally of the same trophic level, where one organism experiences a negative effect convergence. The increasing similarity of two seres over time (antonym: divergence) Cornucopians. Those who believe technology and human ingenuity will prevail over environmental limitations (antonym: neo-Malthusians) corridor. An elongated area of similar land use or disturbance (e.g., roads can be considered disturbance corridors across the landscape) cryoturbation. Soil disturbance through freeze/thaw processes cryptobiotic crust. A surface layer (often on arid land soils) composed of spore-producing plants including algae and mosses cyanobacteria. A diverse group of bacteria often found in extreme environments; some are capable of ﬁxing atmospheric nitrogen (synonym: blue-green algae) decomposer. Any organism that feeds by degrading organic matter demographic transition. The hypothesis that as nations develop economically their rate of population growth declines; this has happened historically but may no longer be true desert. An area where evaporation exceeds annual precipitation (generally <250 mm but up to 600 mm) and where plants and animals are sparse and adapted to the lack of free water desertiﬁcation. The formation or expansion of deserts due to natural climatic factors and/or human-induced disturbances such as overgrazing detritivore. Animal that feeds on detritus (dead plant or animal matter); contrast with decomposer, herbivore, and carnivore dieback. Large-scale death or defoliation of plants (typically forest trees) disturbance. A relatively discrete event in time and space that alters the structure of populations, communities, or ecosystems; a loss of biomass that exceeds accumulation of biomass; disturbance can be natural or anthropogenic, discrete or diffuse, from within (via generally biotic means: autogenic)

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or outside (via generally abiotic means: allogenic) the system of interest; disturbance has such characteristics as frequency, extent, magnitude, intensity, and severity. divergence. See convergence ecosystem. A unit of study at any deﬁned scale that encompasses the interaction of both biotic and associated abiotic components ecosystem management. An approach to maintaining or restoring the structure and function of an ecosystem for long-term sustainability ˜ Southern Oscillation (ENSO). A warmEl Nino water current that ﬂows eastward toward Peru (El Ni˜no) generally associated with a ﬂuctuation of atmospheric circulation (Southern Oscillation); changes in ocean temperatures from the norm can affect weather patterns around the world endotherm. Animal that maintains a stable body temperature (antonym: exotherm) entrainment. The dispersal of propagules in largescale wind patterns epiphyte. A plant growing on the surface of another plant establishment. The stage following immigration when a propagule becomes a permanent member of the local community eutrophication. Nutrient enrichment of aquatic communities that can stimulate algal growth and lead to low oxygen levels and death of ﬁsh exotherm. An organism whose body temperature varies with the surroundings (synonym: poikilotherm; antonym: endotherm) facilitation. The positive inﬂuence of one species on another, as early colonizers improve the environment for later arrivals in plant succession feral. Formerly domesticated animals (or plants such as crops) that become wild (e.g., feral cats or dogs) ﬂoodplain. The land at the margins of a river that is periodically ﬂooded folivore. Leaf-eater fragmentation. The creation of patches (increase in spatial heterogeneity), as from logging functional redundancy. The performance of the same ecological role by two or more species fynbos. A vegetation type in South Africa dominated by evergreen shrubs with small leaves glacial foreland. The area of newly-exposed landscape at the front of a retreating glacier

GLOSSARY

glacial moraine. A ridge of rock debris deposited by a glacier groundwater. Water occurring below the earth’s surface that is free to move by gravity herbivore. Plant-eater hurricane. A strong, circular windstorm with maximum wind speeds exceeding 117 km/h humus. Thoroughly decomposed plant matter hydric. Wet hygrophilous. Growing in a moist habitat, as a hygrophilous forest hypoxia. The condition of oxygen deﬁciency in body tissues such as can occur in humans at high elevations immobilization. The conversion of a chemical compound into an organic form as the result of biological activity; the loss of that chemical from the available pool of nutrients in the soil impact assessment. The policy of addressing future environmental effects of a given activity island of fertility. The zone of increased biological activity and associated higher soil nutrient and water availability in the immediate vicinity of a plant especially shrubs in aridlands Jevon’s paradox. The idea that an increased efﬁciency in the use of a resource will lead to an increased use of that resource rather than to a reduction in its use (more roads means more, not less, road trafﬁc) karst. Any region dominated by eroded limestone landforms krummholz. Trees stunted by strong winds, typically at tree line lahar. A catastrophic mud ﬂow on the slopes of a volcano lava. Molten rock expelled by a volcano leaching. The removal of soil nutrients in solution life history characteristics. Traits of organisms that are linked to the developmental stages of the organism (e.g., for plants: dispersal, germination, growth rate, fecundity, longevity) light gap. An opening (typically in a forest) where light levels are higher than in the surrounding forest lignin. An organic substance that binds cellulose ﬁbers (in wood, for example) litterfall. The accumulated plant remains on the soil surface longevity. The life span of an organism

GLOSSARY

mallee. A shrubland in Australia with low, evergreen plants, mostly in the genus Eucalyptus marginalization. The movement of people to highrisk, low-productivity environments marsh. A frequently inundated mineral (not peaty) soil, often at the edge of a lake or river Mediterranean climate. Mild, wet winters and warm, dry summers; this climate occurs in ﬁve regions of the world (southern California, central Chile, southwestern Australia, South Africa, and the Mediterranean Basin mesic. Moist (neither wet [hydric]), nor dry [xeric]) metallophyte. A plant associated with or restricted to metal-rich soils, as on mine wastes mineralization. The conversion of organic matter to an inorganic state through decomposition by soil microorganisms monophagy. A diet composed of only one or a few related plant species (e.g., as demonstrated by insect larvae) mound. The upraised root ball exposed when a tree falls over (cf. pit) mycorrhiza. An association between certain fungi and plant roots that is mutually beneﬁcial neo-Malthusians. Those who believe in Malthus’ prediction that human population growth will overrun human ingenuity at repairing environmental problems (antonym: Cornucopians) nitriﬁcation. The oxidation of ammonia to nitrite or nitrite to nitrate by bacteria (antonym: denitriﬁcation) nitrogen ﬁxation. The reduction of gaseous nitrogen to form organic nitrogenous compounds; performed by certain free-living or symbiotic bacteria old-growth forest. Forest stands that have not recently been cut by humans pampas. A temperate South American grassland patch dynamics. The study of the shifting mosaic of relatively homogeneous areas or patches across a landscape where patches are created and destroyed by disturbance perennial. A plant that lives for more than two years permafrost. Permanently frozen ground phreatophyte. A plant with its roots in ground water pit. The hollow left when a tree is uprooted (cf. mound) plasticity. The capacity of an organism to vary as a result of environmental change playa. The lowest part of an intermontane basin

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that is frequently ﬂooded by run-off from nearby mountains polyphagy. A diet composed of many plant species primary succession. Succession on substrates with no prior soil development productivity. The mass of plant (primary productivity) or animal (secondary productivity) matter produced in an ecosystem during a certain time period propagule. Any dispersing plant or animal part capable of reproduction (e.g., seed, bulb, vegetative sprout, spore, egg, etc.) pyroclastic ﬂow. An eruption from a volcano of a cloud of gas and solids, driven by gravity and hugging the ground regeneration. Renewal of a population through vegetative or sexual reproduction; regrowth of a damaged appendage regeneration wave. Patterns of mortality and subsequent reproduction in forests that move, with time, along lines resembling wave movement toward a shore resilience. Ability of an ecosystem to recover following a disturbance restoration. The recovery of an ecosystem to its predisturbance state; more generally, any reconstruction or rehabilitation of former ecological processes retrogression. Succession that proceeds from a diverse or structurally complex community to one that is less diverse or complex rhizosphere. The zone of soil around a root inﬂuenced by root activity riparian. Pertaining to a river bank Robinson Crusoe paradox. The suggestion that humans are capable of creative problem-solving when placed in difﬁcult situations ruderal. A plant that inhabits a disturbed site, often associated with human dwellings savanna. A tropical vegetation characterized by extensive grasslands interspersed with sparse, emergent trees; typically with seasonal rain and drought sclerophyll vegetation. A plant community dominated by woody plants with thick, evergreen leaves; typical of Mediterranean climates secondary forest. Forest that develops following disturbance secondary succession. Succession on substrates where some pre-disturbance soil remains selective logging. Removal of some top-quality logs

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while the rest are left standing (antonym: clearcutting) sere. The entire sequence of developmental stages during succession serotiny. Fire-stimulated seed release (e.g., from certain pine cones) serpentine soils. Soils high in magnesium that restrict plant growth silviculture. Applied forest ecology soil texture. The proportion of various particle sizes in a soil proﬁle (e.g., gravel, sand, silt, clay) that has a strong inﬂuence on water retention, nutrient availability, and plant growth soliﬂuction. The downhill movement of saturated soils spatial heterogeneity. Variation in ecosystem characteristics across a landscape at any scale stability. Resistance by an ecosystem to change due to disturbance stochastic. Unpredictable strip mine. An open-pit mine that removes surface layers to obtain ore stress. Any existing environmental condition that limits growth and productivity subtropics. The latitudinal zone between 23.5º and 34.0º in either hemisphere succession. A sequential change in species composition or other ecosystem characteristic; cf.: primary, secondary, arrested, convergence, sere, retrogression, chronosequence

GLOSSARY

sustainable (development or management). A policy that attempts to balance resource harvests with long-term maintenance of ecosystem function and structure swamp. A wetland intermittently inundated and dominated by trees, but with no peat development taiga. See boreal forest temperate zone. The latitudinal zone between the tropics (and subtropics) and the boreal zone in the northern hemisphere or the tropics and the polar zone in the southern hemisphere tephra. All air-borne particles ejected from a volcano thermokarst. A landscape where differential melting of permafrost leaves pockets of enclosed depressions, some holding water, and resembling a karst landscape threshold. A limit above or beyond which a process changes dramatically, perhaps no longer able to return to the original state (e.g., severe erosion can change forest into grassland with no chance of forest regeneration) tropical zone. The zone between 23.5ºN and 23.5ºS latitude vivipary. Producing live offspring from within the body of the parent weed. A plant growing where it is unwanted wetland. A low-lying area periodically inundated or submerged by fresh or saline water wrack. Marine plant life cast up on shore xeric. Dry

SYSTEMATIC LIST OF GENERA All taxa mentioned in the text are here included in their taxonomic position. Division of genera is to the Family level in higher plants, to the Order level in arthropods and vertebrates, to the Phylum, Subphylum or Class elsewhere. Names of intermediate taxa mentioned are inserted immediately below the higher taxon to which they are ascribed, unless they are themselves divided. Within Class, Order or Family, lower taxa are listed in alphabetical order.

MONERA (PROKARYOTA) ARCHAEBACTERIA EUBACTERIA Rhizobium

CYANOBACTERIA NOSTACALES

Collemataceae Collema Parmeliaceae Parmeliopsis Ramalinaceae Ramalina Stereocaulaceae Stereocaulon

PTERIDOPHYTA EQUISETINES Equisetaceae Equisetum

FILICOPSIDA Aspleniaceae Thelypteris Cyatheaceae Alsophila Cyathea Dennstaedtiaceae Pteridium Dicksoniaceae Dicksonia Dryopteridaceae Dryopteris Gymnocarpium Matteuccia Polypodiaceae Microsorum Thelypteridaceae Dicksonia

OOMYCOTA Phytophthora

Nostoc

ZYGOMYCOTA

FUNGI ACTINOMYCOTA Frankia Streptomyces

Entomophaga Glomus

PLANTS CHLOROPHYTA

ASCOMYCOTA Alternaria Ceratocystis Cochliobolus Cryphoonectria Drechslera Endothia Leptographium

Chlorophyceae Prasiolaceae Prasiola

Armillaria Cyathus Cronartium Daedaleopsis Geaster Gloiodon Haploporus Lycoperdon Pisolithus

DEUTEROMYCOTA Humicola Penicillium

LICHENES Caloplacaceae Caloplaca Cladoniaceae Cladina Cladonia

Lycopodiaceae Lycopodium

CHRYSOPHYTA Chrysophyceae Phaeocystis

EUGLENOPHYTA BASIDIOMYCOTA

LYCOPSIDA

PINOPHYTA (GYMNOSPERMAE)

Chlamydomonas

Cupressaceae Chamaecyparis Juniperus Thuja Cycadaceae Macrozamia Pinaceae Abies Larix Picea Pinus Pseudotsuga Tsuga Podocarpaceae Dacrydium Prumnopitys Taxaceae Taxus Taxodiaceae Cryptomeria Sequoia Taxodium

BRYOPHYTA MUSCI Bryaceae Bryum Pohlia Dictanaceae Dicranum Funariaceae Funaria Grimmiaceae Rhacomitrium Hypnaceae Hylocomium Pleurozium Ptilium Rhytidiadelphus Polytrichaceae Polytrichum Sphagnaceae Sphagnum

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MAGNOLIOPHYTA (ANGIOSPERMAE) LILIOPSIDA (MONOCOTYLEDONES) Alismataceae Sagittaria Bromeliaceae Pitcairnia Cyperaceae Bulbostylis Carex Cladium Elyna Kobresia Scirpus Heliconiaceae Heliconia Hydrocharitaceae Egeria Hydrocharis Juncaceae Luzula Prionium Liliaceae Clintonia Erythronium Maianthemum Uvularia Orchidaceae Aceras Eria Goodyera Herminium Listera Ophrys Poaceae Agropyron Agrostis Ammophila Andropogon Anthoxanthum Arctagrostis Arctophila Aristida Arrhenatherum Avena Bothriochloa Bouteloua Brachiaria Briza Bromus Calamagrostis Cenchrus

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SYSTEMATIC LIST OF GENERA Chionochloa Chusquea Cinna Cynodon Dactylis Cortaderia Danthonia Deschampsia Digitaria Distichlis Dupontia Echinochloa Ehrharta Elymus Elytrigia Eragrostis Eriophorum Festuca Holcus Hordeum Hyparrhenia Lerchenfeldia Leymus Lolium Loudetia Melica Melinis Microstegium Miscanthus Nardus Oryza Oryzopsis Panicum Paspalum Pennisetum Phalaris Phragmites Piptochaetium Poa Puccinellia Rendlia Sasa Schismus Schizachyrium Sesleria Setaria Sinarundinaria Sorghum Spartina Sporobolus Stipa Taeniatherum Tristachya Triticum Zea Pontederiaceae Eichhornia Pontederia

MAGNOLIOPSIDA (DICOTYLEDONES) Acanthaceae Isoglossa Aceraceae Acanthopanax

Acer Aextoxicaceae Aextoxicon Aizoaceae Carpobrotus Amaranthaceae Amaranthus Anacardiaceae Rhus Schinus Apiaceae Aegopodium Daucus Heracleum Araliaceae Acanthopanax Aralia Asteraceae (Compositae) Ajania Ambrosia Ammophila Anaphalis Artemisia Aster Baccharis Berroa Bidens Brachylaena Cecropia Centaurea Chaptalia Chevreulia Chrysanthemum Chrysothamnus Cirsium Conyza Crepis Encelia Erigeron Eupatorium Facelis Flourensia Gnaphalium Gutierrezia Helianthus Hieracium Hymenoclea Hypochaeris Jaumea Lactuca Leersia Micropsis Nassauvia Parthenium Petasites Picris Rhanterium Senecio Solidago Sonchus Tagetes Taraxacum Tussilago Ursinia Vernonia Xanthium

Balsaminaceae Impatiens Betulaceae Alnus Betula Corylus Ostrya Bombacaceae Ochroma Boraginaceae Cordia Brassicaceae Berteroa Brassica Cakile Cochlearia Coronopus Sisymbrium Thlaspi Bursuraceae Dacryodes Cactaceae Carnegiea Ferocactus Opuntia Caesalpinaceae Gleditsia Caprifoliaceae Diervilla Leycesteria Linnea Lonicera Sambucus Caryophyllaceae Cerastium Colobanthus Minuartia Sagina Silene Stellaria Casuarinaceae Casuarina Celastraceae Maytenus Ceratophyllaceae Ceratophyllum Chenopodiaceae Atriplex Beta Chenopodium Corispermum Halogeton Kochia Salicornia Salsola Commeliniaceae Murdannia Convolvulaceae Dichondra Coriariaceae Coriaria Cornaceae Cornus Nyssa Crassulaceae Crassula

Cucurbitaceae Cucumis Cucurbita Cunoniaceae Cunonia Weinmannia Cyrillaceae Cliftonia Cyrilla Dipsacaceae Dipsacus Dipterocarpaceae Shorea Ebenaceae Diospyros Eleagnaceae Eleagnus Ericaceae Arctostaphylos Calluna Chimaphila Empetrum Ledum Rhododendron Vaccinium Euphorbiaceae Euphorbia Jatropha Macaranga Manihot Ricinus Sapium Tragia Fabaceae Acacia Adesma Astragalus Brachystegia Cercidium Crotalaria Dacryodes Cytisus Desmanthus Glycine Leucaena Lotus Lupinus Medicago Melilotus Mimosa Olneya Ononis Parkinsonia Phaseolus Pithecellobium Prosopis Robinia Soja Trifolium Vicia Fagaceae Castanea Castanopsis Fagus Nothofagus Quercus

SYSTEMATIC LIST OF GENERA Frankeniaceae Frankenia Gentianaceae Gentiana Geraniaceae Erodium Megalomastoma Geranium Haloragaceae Myriophyllum Juglondaceae Carya Juglans Juncaceae Juncus Lamiaceae Becium Haumaniastrum Hedeoma Lauraceae Cryptocarya Sassafras Linaceae Linum Lythraceae Heimia Lythrum Magnoliaceae Drimys Liriodendron Magnolia Malvaceae Anoda Sphaeralcea Melastomaceae Dreissenia Miconia Meliaceae Melia Swietenia Trichilia Monimiaceae Doryphora Peumus Moraceae Broussonetia Cecropia Ficus Morus Myricaceae Myrica Myristicaceae Myristica Myrsimaceae Ardisia Myrtaceae Angophora Eucalyptus Leptospermum Melaleuca Metrosideros Psidium Syzygium Tristaniopsis Nyctaginaceae Abronia

775 Nymphaeaceae Nuphar Ochnaceae Ochna Oleaceae Fraxinus Ligustrum Olea Oxalidaceae Oxalis Pandanaceae Freycinetia Papaveraceae Chelidonium Passiﬂoraceae Passiﬂora Pittosporaceae Pittosporum Plantaginaceae Plantago Plantanaceae Platanus Plumbaginaceae Armeria Polemoniaceae Ipomopsis Polygalaceae Polygala Polygonaceae Fallopia Muehlenbeckia Polygonum Reynoutria Portulacaceae Portulaca Primulaceae Trientalis Pyrolaceae Pyrola Randiaceae Randia Ranunculaceae Anemone Rhamnaceae Ceanothus Discaria Maesopsis Rhizophoraceae Rhizophora Rosaceae Acomastylis Adenostoma Coleogyne Crataegus Dasiphora Dryas Filipendula Margyricarpus Prunus Quillaja Rosa Rubus Sorbus Spiraea Rubiaceae Coffea

Galium Rutaceae Teclea Vepris Salicaceae Populus Salix Sapindaceae Allophylus Sapotaceae Mimusops Sarraceniaceae Sarracenia Saxifragaceae Mitella Ribes Scrophulariaceae Melampyrum Paulownia Verbascum Veronica Simaroubaceae Ailanthus Solanaceae Datura Nicotiana Physalis Solanum Tamaricaceae Tamarix Tiliaceae Tilia Ulmaceae Celtis Trema Ulmus Urticaceae Dendrocnide Urtica Valerianaceae Valeriana Verbenaceae Aloysia Lantana Phyla Tectona Verbena Zygophyllaceae Larrea

ANIMALS PROTOZOA NEMATODA ANNELIDA CHAETOPODA Lumbricidae

OLIGOCHAETA Aporrectodea Megascolecidae Millsonia

POLYCHAETA

ARTHROPODA CRUSTACEA COPEPODA Euphausia MALACOSTRACA Corophium Cambaridae Carcinus Hemilepistus Sphaeroma Uca Ucides OSTRACODA

MYRIAPODA DIPLOPODA Diplopoda

INSECTA COLLEMBOLA COLEOPTERA Anoplognathes Dendroctonus Elaphidion Hoplostines Hylurgopinus Ips Nebria Onomacris Platypodidae Scolytidae Scolytus Zygogramma DIPTERA Drosophilidae HEMIPTERA HOMOPTERA Adelges Heteropsylla HYMENOPTERA Antistrophus Apis Coccinoidea Crematogaster Iridomyrmex Perga Pheidole Pogonomyrmex Sirex LEPIDOPTERA Bucculatrix Choristoneura Coloradia Dendrolimus Epirrita Euphydryas Lymantria Operophtera Orgyia Plutella ORTHOPTERA Acrididae Agamemnon Arphia Blattidae Brachystola Dactylotum

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SYSTEMATIC LIST OF GENERA

Gryllidae Lamponius Locustana Mantidae Melanoplus Schistocerca THYSANOPTERA

ARACHNIDA ACARI Cryptostigmata Glycyphagus Mesostigmata Oribatida ARANEAE Leucauge Modisimus Pholcidae TARDIGRADA

MOLLUSCA BIVALVIA Corbicula

GASTROPODA Alcadia Austroselenites Caracolus Cepolis Euchordrus Gaeotis Littorina Marisa Megalomastoma Melampus Nenia Oleacina Platysuccinea Polydontes Subulina Vaginulus

CHORDATA VERTEBRATA PISCES Cyprinodontiformes Cyprinodon AMPHIBIA Anura Bufo Eleutherodactylus Rana REPTILIA Crocodylia Alligator Crocodylus Squamata Anolis Testudinata Gopherus

AVES Anseriformes Anas Anser Branta Chen Cygnus Apodiformes Anthracothorax Chlorostilbon Threnetes Ciconiiformes Catharacta Chionis Diomedea Eudyptes Gypaetus Macronectes Procellariidae Pygoscelis Sarcorampus Columbiformes Columba Geotrygon Coraciaformes Todus Galliformes Colinus Coturnix Gallus Lagopus Phasianus Lariformes Larus Passeriformes Alauda Amphispiza Carduelis Catharus Coereba Corvus Cyanocompsa Cyphorhinus Dendroica Dumetella Dysithamnus Empidonax Euphonia Gymnopithys Hylophylax Loxigilla Margarops Neospingus Passer Pipra Platyrinchus Regulus Seiurus Setophaga Spizella

Sturnella Toxostoma Turdus Vireo Wilsonia Zonotrichia Piciformes Melanerpes Picoides Sphenisciformes Aptenodytes Strigiformes Speotyto Strix MAMMALIA Artiodactyla Alces Antilocapra Bison Bos Bubalus Camelus Capra Cervus Lama Odocoileus Oreamnos Ovibos Ovis Rangifer Rupicapra Sus Vicugna Carnivora Callorhinus Canis Felis Lobodon Martes Mirounga Odobenus Panthera Pinnipedia Arctocephalus Pterepodidae Taxidea Ursus Vulpes Cetacea Balaena Balaenoptera Ceratodon Eschrictius Chiroptera Artibeus Chaerophon Chalinolobus Emballonura Eptesicus Hipposideros

Monophyllus Mormopterus Nyctophilus Pteropus Rhinolopus Scotorepens Stenoderma Diprotodontia Trichosurus Hyracoidea Procavia Insectivora Erinaceus Lagomorpha Lepus Oryctolagus Sylvilagus Perissodactyla Equus Tapirus Primates Cerceopithecus Proboscidea Elephas Loxodonta Rodentia Castor Clethrionomys Cynomys Dicrostonyx Dipodomys Eutamias Geomys Hystrix Lemmus Marmota Microtus Mus Myocaster Myospalax Ochotona Ondatra Onychomys Otomys Perognathus Peromyscus Proechimys Rattus Reithrodontomys Sciurus Spermophilus Tamias Thomomys Zapus Sirenia Hydrodamalis Trichechus

AUTHOR INDEX 1 Aarrestad, P.A., 21, 35 Aarssen, L.W., 615, 625–627, 627, 630, 631 Abbas, I., 471, 475, 484 Abbott, L.K., 381, 383, 527, 541 Abbott, M.L., 555, 565 Abe, S., 188, 199, 208, 215, 219 Abele, G., 62, 76, 78, 94, 116 Aber, J.D., 198, 217, 462, 464, 546, 549, 550, 565, 615, 621, 627, 674, 686, 760, 766 Abernathy, C.L., 509, 517 Ables, E.D., 390, 395 Abrahams, A.D., 313, 328 Abrams, M.D., 169, 182, 189, 190, 203, 205, 215, 220, 238, 246, 641, 652 Abrams, R.H., 697, 705 Abyzov, S.S., 27, 32, 42, 94 Acea, M.J., 533, 538 Ackerman, J.D., 241, 242, 246, 469, 485 Ackley, S.F., 50, 105 Acosta, A., 418–420, 447 Acra, M.A., 578, 582 Adachi, N., 599, 602 Adam, K.M., 76, 94 Adam, P., 337, 356 Adams, A.B., 17, 27, 29, 35, 591, 594, 598, 603 Adams, B.W., 556, 566 Adams, G., 433, 434, 449 Adams, J.A., 188, 209, 210, 215, 388, 394 Adams, J.B., 225, 228, 250, 471, 485, 713, 721 Adams, M.S., 317, 325 Adams, P.W., 590, 603 Adams, R.H., 663, 670 Adams, S., 317, 325 Adams, V.D., 591, 603 Adams, W.P., 43, 94, 131, 134 Adamson, D.A., 31, 36, 44, 52, 66, 91, 94, 116 Adamson, E., 42, 65, 66, 78, 87, 94 Adamson, H., 42, 65, 66, 78, 94 Adamson, J.K., 548, 549, 554, 565 Adelung, D., 62, 70, 100, 121 Adkisson, C.S., 213, 218 Aerts, R., 621, 627 Aey, W., 397, 398, 409 1

Page references to text are in roman type, to bibliographical entries in italics.

Agee, J.K., 456–458, 464, 713, 719 Ahlgren, C.E., 572, 573, 581 Ahlgren, C.F., 639, 652 Ahlgren, I.F., 572, 581, 639, 652 Ahlstrand, G.M., 62, 73, 78, 94 Ahmad, N., 472, 482 Aide, T.M., 256, 258, 266, 267, 477, 482 Aiello, A., 209, 220, 229, 231, 232, 234, 238, 239, 250 Aina, P.O., 124, 132 Ainley, D.G., 91, 101 Aiso, M., 65, 118 Akachuka, A.E., 199, 215 Akiyama, M., 54, 94 Al-Durrah, M., 513, 515 Al-Thagaﬁ, K.M., 388, 395 Alaback, P.B., 193, 222 Alaka, M.A., 235, 246 Albano, M., 345, 351, 359, 438, 447 Albert, P.J., 175, 182 Alberts, E.E., 126, 135 Albertson, F.W., 714, 722 Albrecht, W., 210, 219 Albritton, D.L., 64, 101, 112 Alcorn, S.M., 599, 608 Aldon, E.F., 378, 384, 526, 536, 537, 538, 540 Alexander, I., 472, 482 Alexander, L.T., 64, 104 Alexander, M., 127, 132, 530, 538, 540, 544 Alexander, S.K., 337, 347, 363 Alexander, V., 68, 90, 110 Alexandratos, N., 724, 745 Alfaro, R.I., 261, 266 Allan, C.J., 43, 94 Allan, G.E., 429, 431, 448 Allan, N.J.R., 57, 94 Allan, T., 309, 314, 316, 323, 325 Allanson, B.R., 39, 109 Allen, B.P., 190, 194, 195, 200, 201, 215, 221 Allen, E.B., 17, 26, 28, 34, 79, 94, 311, 320, 325, 325, 417, 418, 423, 451, 522–525, 527–533, 535, 537, 538, 539, 543, 681, 684, 709, 719 Allen, G., 444, 445 Allen, J.P., 660, 669 Allen, J.S., 64, 104

777

Allen, M.F., 79, 94, 153, 157, 415, 416, 444, 452, 522–533, 535–537, 538–544, 645, 652, 653, 681, 684 Allen, O.N., 506, 516 Allen, R.B., 679, 681, 684 Allen, S.E., 53, 94, 575, 581, 582 Allen, T.F.H., 4, 16, 149, 159, 597, 607, 708, 710, 712, 719, 721, 725, 737, 746 Allen, W., 131, 132 Alliende, M.C., 279, 280, 283 Allison, F.E., 123, 127, 128, 132, 562, 565 Allison, J.F., 66, 88, 94 Allison, S.K., 345, 347, 356, 356 Allison, T., 193, 215 Allmaras, R.R., 510, 515 Allnutt, F.C.T., 58, 112 Almasi, P., 66, 97 Alonso, C.V., 124, 132 Alpert, S., 129, 134 Alstad, D.N., 257, 258, 267 Alston, A.M., 527, 542 Alvarado, A., 60, 121 Alvarez, E., 536, 542 Alvarez, H., 376, 382 Alvarez, J., 555, 568, 637, 643, 652, 657 Alvarez, R., 496, 501 Alvarez-Buylla, E.R., 226, 249, 471, 477, 482 Amaral, P., 478, 486 Amaranthus, M.P., 694, 704 Ambrose, L.J.H., 51, 110 Ames, R.N., 530, 534, 541 Amiro, B.D., 64, 66, 116 Amman, G.D., 593, 603 Ammirati Jr, J., 591, 604 Amstrup, S.C., 65, 94 Amundson, R.G., 589, 603 Anable, M.E., 418, 423, 445, 449 Anders, K., 403, 411 Andersen, D.C., 79, 94, 153, 157, 528, 530, 538, 593, 600, 603 Anderson, A.B., 665, 669 Anderson, A.R., 548, 549, 554, 565 Anderson, B.J., 55, 114 Anderson, C., 341, 356 Anderson, D.C., 563, 564, 565 Anderson, J., 472, 485 Anderson, J.E., 294, 304, 307, 310, 329, 419, 420, 445, 676, 684 Anderson, J.M., 130, 131, 132, 487–489, 497, 500, 502, 593, 608

778 Anderson Jr, J.P., 595, 598, 607 Anderson, K., 277, 278, 282, 283 Anderson, K.L., 641, 655 Anderson, L.E., 590, 607 Anderson, N.H., 173, 184, 211, 218, 711, 720 Anderson, R.C., 288, 295, 301, 303, 638, 652, 713, 719 Anderson, R.L., 324, 327 Anderson, R.V., 524, 539 Anderson, S.J., 438, 442, 443, 445 Anderson, W.B., 508, 515 Anderson-Wong, P., 442, 445 Andersson, R., 548, 554, 569 Andresen, H., 350, 353, 356 Andrews, P.L., 456, 465 Andreyev, A., 84, 94 Angell, R.F., 337, 356 Angelstam, P., 162, 174, 182 Angermuller, S., 506, 515 Angino, E.E., 58, 112 Aniya, M., 66, 94 Anonymous, 44, 94, 137, 157 Anthony, R., 645, 653 Antibus, R.K., 90, 109 Antonovics, J., 374, 382, 403, 412 Antos, A.J., 142, 157 Antos, J.A., 140, 142, 144, 148, 157, 160, 587, 610, 618, 629 Aplet, G.H., 140, 142, 147, 157, 189, 204, 215, 438, 442, 443, 445, 587, 592, 603, 609, 675, 684 Apley, M.L., 340, 362 Apon, L.P., 340, 358 Appanah, S., 229, 230, 232–234, 240, 242, 250, 471, 482 Appleﬁeld, M., 223, 224, 239, 249 Appleyard, A., 508, 518 Arad´ottir, A.L., 52, 95 Aras, N.K., 123, 134 Arce, P., 274, 284 Archambault, S., 174, 175, 182 Archer, A.C., 30, 32, 33, 591, 593, 603 Archer, R.R., 198, 222 Archer, S., 55, 83, 94, 288, 303, 322, 325 Archer, S.A., 560, 565 Archibold, O.W., 453, 464 Arianoutsou, M., 271–273, 283 Arikaynen, A.I., 58, 94 Arines, J., 525, 539 Armesto, J.J., 4, 16, 31, 36, 139, 159, 188, 215, 223, 224, 236, 250, 585, 586, 596–598, 602, 607, 608, 611, 613, 616–618, 628, 630, 679, 686, 709–712, 721 Armstrong, A.J., 281, 282 Armstrong, D.P., 681, 684 Armstrong, F.A.J., 64, 117

AUTHOR INDEX Armstrong, T., 57, 94 Armstrong, T.D., 701, 704 Arnalds, A., 52, 92, 95 Arnalds, O., 52, 95 Arnborg, T., 163, 164, 170, 177, 182, 183 Arnebrant, K., 533, 539 Arno, S.F., 456, 464 Arnold, G.W., 282, 284, 708, 719 Aro, E.-M., 55, 111 Aronson, J., 276, 279, 280, 282–284, 325, 325, 674, 676, 679, 682, 684, 760, 765 ˚ 170, 182 Aronsson, K.A., Arp, P.A., 548, 554, 567 Arriaga, L., 189, 195, 215, 471, 482 Arrington, D.A., 682, 687 Arritt, S., 309, 325 Arroyo, M.K., 271–273, 283 Arsufﬁ, T.L., 432, 433, 446 Art, H.W., 235, 247 Artman, J.D., 529, 542 Artyshkova, L.V., 532, 544 Arya, L.M., 507, 515 Asbjornsen, H., 211, 212, 222 Asbury, C.E., 245, 249, 752, 753, 766 Ascaray, C., 589, 607 Aschmann, H., 271, 275, 282, 283 Ascorra, C., 478, 483 Ash, H.J., 406, 409, 592, 603 Ashbury, C.E., 211, 222 Ashley, J., 536, 539 Ashton, D.H., 75, 95, 147, 160, 429, 431, 450, 591, 609 Ashton, P.J., 432, 445, 715, 719 Ashton, P.M.S., 234, 247, 460, 466, 469, 481, 485 Ashton, P.S., 229, 242, 247 Ashton, P.W., 228, 229, 247 Ashton, S.R., 17, 27, 36 Asim, K.B., 526, 539 Aspi, J., 62, 104 Asquith, N.M., 147–149, 154, 160 Atkins, L., 415, 416, 421–425, 445, 448, 681, 685 Atkinson, A., 663, 669 Atkinson, I.A.E., 142, 157, 587, 603, 681, 687 Atlas, R.M., 63, 67, 68, 87, 89, 90, 95, 97, 103, 105, 109, 116, 526, 539 Atlavinyte, O., 130, 132 Attiwill, P.M., 174, 182 Attoe, O.J., 506, 516 Aubertin, G.M., 546, 565 Auble, G.T., 429, 431, 451 Audebert, R., 524, 540 Auen, L.M., 294, 305 Auerbach, N.A., 589, 603 Auerback, S.I., 40, 70, 95 Auerswald, K., 506, 515

Augspurger, C.K., 212, 213, 215, 221, 240, 241, 250, 251, 602, 603 Auhagen, A., 403, 404, 409 Auld, B.A., 442, 445 Aumen, N.G., 173, 184, 211, 218, 711, 720 Ausmus, B.S., 524, 541 Australian Environment Protection Agency, 379, 382 Avenant, N.L., 70, 99 Avenda˜no, J., 279, 280, 282, 284 Avery, D.T., 509, 515 Avis, A.M., 379, 383 Axelrod, D.I., 310, 325 Axelrod, R., 662, 669 Axelsson, G., 549, 565 Ayers, P.D., 507, 515 Ayers Jr, R.C., 68, 95 Ayling, R.D., 476, 482 Ayyad, M.A., 589, 599, 603 Azcon, R., 525, 539 Azkona, P., 343, 359 Baath, E., 533, 539 Babb, T.A., 47, 82, 83, 95 Babinskawerka, J., 529, 543 Babintseva, R.M., 167, 182 Bacci, E., 65, 101 Backshall, D.J., 424, 446 Baczocha, N., 529, 542 Badger, G.J., 190, 205, 220 Baeza, A., 64, 95 Bagley, C.F., 386, 396 Bahre, C., 279, 283 Bahre, C.J., 321, 326 Bahret, S., 415, 416, 421–423, 444, 452 Bai, X.F., 82, 121 Bailey, A., 455, 459, 466 Bailey, A.W., 556, 564, 568, 569, 572–576, 583, 644, 654 Bailey, D., 373, 383, 535, 539 Bailey, S.W., 682, 686 Bainbridge, D.A., 478, 483, 535–537, 539–541 Baird, A.M., 439, 445 Baker, A.C., 281, 285 Baker, A.J.M., 374–376, 382, 383, 524, 543 Baker, C.J., 510, 519 Baker, D.L., 296, 304 Baker, F.A., 253, 258, 260, 268 Baker, H.G., 493, 501 Baker, J.M., 337, 357 Baker, J.R., 536, 542 Baker, L.S., 528, 539 Baker Jr, M.B., 459, 464 Baker, P.E., 43, 95 Baker, R., 667, 669

AUTHOR INDEX Baker, S.R., 380, 381, 382 Baker, W.L., 88, 104 Baker-Blocker, A., 65, 95 Bakker, J.D., 619, 620, 623, 628 Bakker, J.P., 348, 350, 353, 356, 357, 497, 501, 502 Bakr, E.W., 528, 541 Bakshi, B.K., 646, 652 Balduzzi, A., 279, 283 Baldwin, A.D.J., 674, 684 Baldwin, A.H., 333, 336, 341, 345, 348, 351, 355, 357 Bale, J., 42, 96 Bales, R.C., 67, 121 Balisky, A.C., 162, 167, 183 Balke, K.D., 402, 410 Balks, M.R., 62, 63, 78, 98 Ball, E., 142, 157, 593, 603 Ball, E.D., 639, 655 Ballantyne, C.K., 21, 24, 33, 66, 95 Ballard, T.M., 52, 100 Baltz, D.M., 432, 445 Banchero, N., 44, 112 Bandyopadhyay, J., 56, 95 Banham, W.M.T., 474, 484 Bani, E., 636, 652 Banks, M.K., 369, 384 Bannister, P., 436, 445 Bao, X.K., 60, 122 Baranowski, S., 24, 33 Barbault, R., 681, 686 Barbero, N., 496, 501 Barbosa, P., 262, 267 Barbour, A.K., 365, 368, 369, 382 Barbour, M.G., 321, 326, 589, 594, 603 Barden, L.S., 190, 198, 202, 203, 205, 206, 215, 427, 428, 446 Barea, J.M., 525, 539 Barendregt, A., 682, 684 Bargagli, R., 64, 87, 95 Barik, S.K., 188, 202, 206, 215 Barinaga, M., 68, 95 Barkman, J.J., 74, 95 Barlow, S.B., 524, 527, 538 Barnes, G.G., 316, 328 Barnes, P.W., 87, 95 Barnett, T.P., 50, 95 Barnola, J.M., 66, 95, 113 Baron, J., 438, 443, 446 Baroni, C., 89, 95 Barreto, P., 467, 472, 478, 484, 486 Barrett, G.W., 40, 95 Barrett, J.W., 460, 464 Barrett, P.E., 51, 118 Barrett, S.W., 456, 464 Barrie, L.A., 64, 95 Barros, A.C., 478, 486

779 Barrow, C.J., 2, 15, 125, 132, 385, 394, 589, 591, 603, 660, 661, 663–665, 668, 669, 669, 689, 702, 704, 753, 765 Barrow, N.J., 646, 652 Barrows, H.L., 127, 132 Barry, R.G., 39, 66, 67, 95 Barsdate, R.J., 68, 90, 110 Barsky, T., 129, 134 Bartell, S., 661, 669 Bartell, S.M., 524, 541 Barth, R.C., 535, 539, 557, 565 Bartha, R., 67, 95 Bartlett, K.B., 89, 116 Bartolome, J.W., 293, 294, 304, 417, 419, 446, 448 Barton, A.M., 611, 612, 629 Barton, K.J., 62, 70, 121 Barton, R., 63, 114 Bartos, D.L., 555, 565 Bartus, W.S., 387, 394 Baskerville, G.L., 171, 182, 751, 765 Basnet, K., 231, 247 Bassett, P.A., 644, 652 Bassham, C.R., 288, 303 Batcheler, C.L., 438, 441, 446 Battaglia, L.L., 190, 192, 201, 215 Battan, L.J., 194–196, 215 Battisti, E., 64, 87, 95 Battles, J.J., 225, 226, 230, 252 Batzer, D.P., 343, 357 Batzli, G.O., 43, 53–55, 84, 95, 96, 106, 110 Bauder, J.W., 504, 516 Baudou, E., 170, 173, 182 Bauer, A., 510, 515 Bauer, G., 287, 306 Baugh, C.L., 390, 396 B¨aumler, E., 17, 36 Bawa, K.S., 472, 483, 485 Bayﬁeld, N.G., 78, 96 Bazely, D.R., 83, 96, 336, 345, 353, 357 Bazzaz, F.A., 189, 198, 206, 211, 215, 217, 219, 221, 307, 322, 325, 422, 449, 585, 592, 598, 603, 614–617, 619, 621–627, 628–632 Beach, H.F., 592, 599, 607 Beach, J.H., 474, 483, 642, 643, 653 Beale, O.W., 507, 509, 516 Beall, C.M., 56, 60, 65, 102 Beaman, J., 473, 482 Beaman, R., 473, 482 Beamon, J.H., 147, 157 Bear, G.D., 296, 304 Beard, E.B., 334, 337, 357 Beard, J.S., 142, 143, 157, 228–230, 240, 241, 247, 271, 281, 283, 284, 592, 603 Beare, M.H., 538, 539 Beare, P.A., 340, 345, 347, 349, 357, 363

Beatley, J.C., 422, 423, 446 Beatty, S.W., 190, 208–210, 212, 213, 215, 713, 719 Beatty, T.F., 90, 101 Beauchamp, E.G., 549, 566 Beauliew, J., 560, 567 Beaver, R.A., 647, 652 Becher, H.H., 388, 394 Beck, E., 43, 115 Beck, M.B., 660, 670 Becker, P., 229, 250, 602, 603 Beckett, C.L., 337, 361 Becquey, J., 189, 200, 215 B´edard, J., 342, 345, 357, 359 Bedford, B.L., 354, 357 Beerling, D.J., 430, 431, 446 Begin, C., 169, 175, 185 Begon, M., 593, 603 Behan-Pelletier, V.M., 523, 539 Behrens III, W.W., 723, 746 Beisner, B.E., 617, 619, 620, 630 B´eland, M., 164, 182 B´elanger, L., 342, 345, 357 Belcher, J.W., 617, 619, 620, 630 Bell, D.T., 680, 682, 684, 685 Bell, K.L., 78, 79, 96 Bell, S., 682, 684 Bell, S.S., 337, 357 Bellairs, S.M., 680, 684 Bellamy, D.J., 76, 96 Belleﬂeur, P., 614, 616, 618, 622, 624, 628 Bellingham, P.J., 223, 238–240, 242–244, 247 Belnap, J., 324, 325, 563, 564, 566 Belsky, A.J., 287, 299, 303, 305, 421, 446, 578, 581, 589, 603 Ben-Horin, R.A., 276, 285 Benarth, R.F., 646, 647, 652 Bendali, F., 436, 446 Bendotti, S., 681, 686 Benedict, F., 255, 267 Benedict, J.B., 51, 96 Bengtson, J.L., 91, 96 Bengtsson, G., 528, 541 B´enito-Espinal, D.B., 2, 15 B´enito-Espinal, E., 2, 15 Benn, D.I., 21, 33, 66, 95 Bennett, E., 471, 473, 482 Bennett, H.H., 126, 132 Bennett, I.L., 644, 645, 654 Benninghoff, A.S., 52, 53, 96 Benninghoff, W.S., 52, 53, 96 Bennington, S.L., 64, 96 Benson, C.S., 43, 106 Berendse, F., 621, 628 Berg, A., 43, 55, 112 ˚ 173, 182 Berg, A., Berg, A.B., 211, 216

780 Berg, B., 292, 304, 549, 565 Berg, D.R., 462, 465 Berg, W.W., 64, 96 Bergelson, J., 421, 422, 446 Berger, J.J., 674, 681, 684 Bergeron, Y., 161, 162, 164–166, 168, 169, 171, 172, 174–178, 181, 182, 182–185 Bergesen, H.O., 57, 96 Bergsma, B.M., 20, 33 Bergsteinsson, J.L., 167, 185 Bergstrom, D.M., 44, 89, 96, 116 Bergstrom, R., 579, 582 Bergstr¨om, R., 170, 182 Berish, C., 459, 465 Berish, C.W., 679, 685 Berkowicz, S.M., 131, 134 Berkowitz, A., 131, 134 Berkowitz, A.R., 598, 603 Berlinger, B.P., 389, 394 Berlow, E.L., 764, 765 Berner, P.O.B., 231, 232, 247 Bernstein, N.P., 70, 112 Berrisford, M.S., 21, 35 Berruti, A., 53, 54, 116, 121 Bertness, M.D., 333, 336–338, 345, 346, 351, 357, 359, 361, 362, 595, 603, 614–616, 618, 622, 624, 628 Berube, D.E., 418, 450 Best, T.R., 645, 652 Bester, M.N., 69, 119 Betancourt, J.L., 322, 325 Bethge, K., 210, 219 Bevier, G., 472, 484 Bevins, R.L., 126, 133 Beymer, R.J., 562, 566 Bezdicek, D.F., 488, 502 Bickerton, R.J., 18, 20, 33 Bidigare, R.R., 42, 66, 87, 96, 100 Bidin, K., 475, 483 Bidleman, T.F., 65, 96 Bidwell, T.G., 297, 303 Biebl, R., 41, 96 Bien, A., 226, 229, 231, 249 Bierbach, H., 403, 411 Bierregaard Jr, R., 259, 267, 467, 471, 475, 477, 482, 484 Bigelow, N.H., 590, 606 Bigger, C.M., 551, 552, 554, 566 Bignell, D.E., 474, 483 Bijlenga, G., 69, 98 Bilbao, B., 423, 446 Bilbrough, C.J., 318, 326 Billheimer, D.D., 318, 327 Billings, W.D., 45, 47, 55, 65, 87, 89, 96, 98, 111, 113, 590, 603 Bindschadler, R.A., 67, 122 Binggeli, P., 436, 441, 446, 599, 603

AUTHOR INDEX Bingham, R.L., 288, 302, 305, 311, 327 Binkley, D., 455, 457, 461, 462, 464, 466 Binkley, M.S., 225, 252 Binns, D., 70, 114 Biondini, M.E., 294, 298, 303, 306 Bird, P.M., 64, 96 Birks, H.J.B., 17, 28–31, 33, 599, 600, 603 Birney, E.C., 292, 298, 306 Bishop, S.C., 78, 96, 591, 603 Biswas, A.K., 660, 669 Biswell, H.H., 576, 581 Bj¨arnason, A.H., 142, 157 Black, A.L., 510, 515 Black, P., 131, 134 Black, R.A., 153, 159, 577, 581 Black, R.F., 51, 96 Blackburn, T.C., 278, 283 Blackburn, W.H., 561, 566, 569 Blaha, M., 524, 539 Blaikie, P.M., 664, 668, 669 Blair, J.M., 535, 539 Blair, R., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Blake, G.R., 507, 510, 515 Blake, J.G., 642, 643, 652 Blanchet, D., 724, 745 Blasco, D.E., 334, 357 Blehr, O., 43, 55, 112 Bleuler, M., 22, 36 Bleuten, W., 715, 720 Bliss, L.C., 28, 33, 39, 45, 47, 52, 62, 68, 78, 79, 82, 83, 90, 95, 96, 100, 102, 103, 113, 120, 139, 142, 144, 148, 149, 151, 157, 158, 577, 581, 591, 605 Block, W., 42, 43, 45, 96, 121 Blocker, H.D., 292, 298, 306 Blom, H.H., 21, 35 Blood, E.R., 195, 200, 201, 217 Bloomﬁeld, J., 212, 222 Bloss, H.E., 537, 539 Blotta, L., 496, 501 Blum, J.L., 343, 357 Blume, H.P., 131, 134, 400–405, 410, 411 Blumer, P., 55, 96 Blundon, D.J., 17, 33, 141, 157, 591, 599, 600, 603 Blydenstein, J., 439, 446 Boardman, R., 263, 267 Bobin, N.Y., 27, 32 B¨ocker, R., 401, 405, 410 Bockstoce, J.R., 58, 96, 110 Bodhaine, B.A., 51, 65, 96 Bodle, M.J., 433, 446 Bodman, G.B., 506, 516 Boeger, P., 45, 115 Boehm, A., 392, 396 Boeken, B., 708, 713, 716, 719

Boerner, R.E.J., 190, 203, 204, 206, 216, 593, 603, 713, 720 Boersma, O.H., 510, 517 Boesch, D.F., 40, 96, 337, 357 Boger, P., 42, 115 Bohac, J., 130, 132 Bohn, A., 64, 96, 97 Bokerman, R., 510, 515 Bokkestijn, A., 228, 231, 232, 249 Bolen, E.G., 334, 357 Bolin, S.B., 313, 326 Bollinger, M.J., 64, 101, 112 Bolton Jr, H., 557, 560, 563, 566 Bonan, G.B., 161, 182 Bonani, G., 66, 97 Bond, G., 17, 34, 66, 97, 600, 601, 606 Bond, J.J., 511, 518 Bond, W.J., 1, 5, 15, 233, 250, 281, 283, 418, 425, 451, 592, 603 Bonde, E.K., 51, 97 Bongers, F., 224–229, 231–234, 238, 242, 249, 250, 252, 471, 486 Bonis, A., 348, 357 Bonneau, M., 512, 515 Bonner, W.N., 55, 64, 69, 79, 91, 94, 97, 113 Bonniﬁeld, M.P., 668, 669 Boose, E.R., 187, 189, 191, 193–195, 198–201, 215, 217, 225, 238, 239, 242, 247, 248, 471, 482, 713, 719 Boot, R.G.A., 478, 485 Booth, J.A., 599, 608 Booth, R.G., 535, 543 Borgegard, S., 590, 608 Boring, L.R., 190, 197, 203, 206, 208, 216 Bormann, B.T., 17, 22, 26, 31, 33, 554, 566, 601, 603 Bormann, F.H., 193, 215, 546, 547, 552, 554, 566, 567, 711, 718, 719, 721, 764, 765 Bornemissza, G.F., 647, 652 Bornkamm, R., 401, 402, 410 Borror, D.J., 645, 652 Bosch, I., 42, 106 Boser¨up, E., 667, 668, 669 Bossard, C.C., 421–423, 445, 446 Boswall, J., 69, 97 Botkin, D.B., 1, 15, 549, 550, 565, 621, 628 Botkin, M.J.M., 555, 556, 566 Bottenheim, J.W., 64, 95 Bottner, P., 292, 304 Boucher, D.H., 235, 236, 238–242, 247, 252, 525, 539, 636, 657 Boukhris, M., 555, 566 Boulet, C., 274, 283 Boulton, A.J., 427, 451 Boulton, G.S., 19, 21–24, 33

AUTHOR INDEX Bourgoin, B.P., 65, 97 Bouss`es, P., 69, 98 Boutin, S., 69, 112 Boutron, C.F., 64, 97 Bowden, D.C., 296, 304 Bowden, R.D., 198, 217, 244, 245, 251, 435, 451 Bowen, G.D., 525, 539, 540, 542 Bower, C.A., 506, 515 Bowers, J.E., 310, 316, 317, 324, 326, 328, 415, 416, 446 Bowland, J.M., 351, 357 Bowles, D.E., 432, 433, 446 Bowles, M.L., 681, 684, 759, 765 Bowman, D.M.J.S., 438, 442, 443, 446, 451 Box, J.E., 507, 519 Box, J.H., 255, 268 Boyd, J.C., 70, 116 Boyer, D.C., 430, 446 Boyer, M.G., 56, 109 Boyle, J.R., 524, 539, 545, 547, 549, 551, 566 Boyle, M., 514, 515 Braat, L.C., 661, 670 Braatne, J.H., 592, 594, 603, 604 Bradford, J.M., 513, 514, 515 Bradley, L.F., 528, 540 Bradley, R., 66, 112 Bradshaw, A.D., 366–377, 382–384, 402, 403, 410, 535, 539, 585, 588, 592, 599, 600, 603, 607, 611, 613, 615, 628, 674, 684 Bradshaw, F.J., 715, 719 Bradshaw, G.A., 462, 463, 466 Bradshaw, R., 352, 362 Bradshaw, R.H.W., 161, 169, 182, 183 Bradstock, R.A., 271, 284 Brady, M.A., 682, 687 Bragg, E., 562, 568 Bragg, T.B., 293, 297, 298, 301, 304, 641, 652 Braham, H.W., 58, 97 Braithwaite, R.W., 429, 431, 441, 446 Braker, H.E., 225, 251 Brand, S., 131, 134, 713, 716, 719 Brand, T., 240, 247 Brandani, A.A., 26, 33 Brande, A., 403, 410, 411 Brandel, G., 200, 222 Brandes, D., 405, 410 Brandt, C.A., 317, 326, 415, 416, 419, 420, 446 Branson, F.A., 555, 557, 566 Bratton, S.P., 190, 193, 208, 215, 218, 438, 442, 446, 716, 720 Braun, M., 175, 186

781 Braun, O.P.G., 225, 228, 250, 471, 485, 713, 721 Braun-Blanquet, J., 17, 33, 52, 97 Bray, J.R., 25, 33, 189, 191, 203, 206, 215, 255, 267 Bray, M.P., 343, 353, 357 Braysay, M., 137, 158 Brazel, A.J., 19, 35 Breckle, S.W., 163, 167, 186, 309, 312, 317, 324, 326, 329 Breckwoldt, R., 681, 684 Breen, C.M., 715, 719 Breman, H., 292, 293, 304 Brenkley, D., 365, 367, 382 Brennan, B., 697, 698, 705 Brenner, A.C., 67, 122 Brenner, F.J., 378, 382 Bressan, M., 526, 542 Brewer, M.C., 76, 78, 94 Brewer, R., 189, 203, 204, 215 Brewerton, H.V., 65, 97 Bridgewater, P.B., 424, 446 Briggs, J.M., 288, 297, 302, 304 Brightman, R., 472, 484 Brinckmann, E., 311, 329 Brink, V.C., 30, 33 Brinkhurst, R.O., 87, 98 Brinson, M.M., 354, 357 Briske, D.D., 547, 566 Brisson, J., 166, 182 Britt, J.R., 618, 628 Britton, M.E., 45, 97 Broadgate, M., 172, 185 Broady, P.A., 27, 31, 33, 36, 43, 45, 88, 97, 105 Brock, T.D., 591, 603 Broecker, W.S., 66, 97 Broekman, R.A., 337, 361 Broggi, M., 57, 97 Broich, W.A., 68, 90, 103 Brokaw, N.V.L., 2, 16, 199, 202, 203, 215, 223–231, 233, 235–243, 247, 248, 250, 252, 333, 357, 478, 480, 486, 590, 605, 635, 641–643, 651, 652, 655, 656, 713, 717, 720, 722, 763, 767 Brongers, M., 350, 353, 356 Brookﬁeld, H., 668, 669 Brooks, C.M., 646, 656 Brooks, P.D., 67, 121 Brooks, R.H., 506, 515 Brooks, R.R., 372, 374, 382 Broom, J., 337, 357 Brothers, N.P., 69, 97 Brothers, T.S., 415, 416, 446 Brotherson, J.D., 428, 431, 446 Brouwer, L.C., 478, 485 Browder, J.O., 662, 670 Brown, A.D., 67, 121

Brown, A.V., 427, 450 Brown, B., 459, 465 Brown, C.J., 317, 326, 416, 446 Brown, D.E., 423, 439, 446 Brown, D.G., 147, 157 Brown, E.T., 590, 603 Brown, G.W., 546, 566 Brown, J., 62, 68, 76, 78, 82, 90, 94, 103, 114, 119, 390, 394 Brown, J.C., 532, 539 Brown, J.H., 319, 326, 327, 717, 719 Brown, J.R., 678, 685 Brown, L., 443, 446, 723, 745 Brown, L.R., 432, 446 Brown, M.C., 371, 383, 599, 604 Brown, M.J., 188, 211, 217 Brown, N., 243, 247 Brown, N.D., 237, 240, 247 Brown, P.J., 424, 451 Brown, R.J.E., 577, 581 Brown, R.W., 79, 94, 97, 98, 530 Brown, S., 230, 232, 233, 247, 471, 479, 482, 482, 678, 679, 684 Brown, T., 462, 464 Brown, V.B., 294, 306 Brown, V.K., 555, 557, 568, 579, 581, 593, 603 Brown, W.H., 587, 603 Brownsey, P., 87, 110 Brozka, R.J., 386–390, 396 Brubaker, H.W., 125, 130, 134, 729, 746 Brubaker, L.B., 66, 102 Bruce, K.A., 419, 420, 425, 446 Bruce, R.R., 128, 133 Bruenig, E., 225, 227, 233, 245, 247 Brugger, E., 57, 97 Bruggers, R.L., 337, 360 Bruijnzeel, L.A., 124, 132 Brum, G.D., 321, 329 Brundrett, M., 575, 581 Bruner, M.C., 715, 721 Brunk, K., 64, 97 Bruns, T., 538, 539 Brussard, P.F., 713, 719 Bryan, K., 47, 97 Bryan, R.B., 63, 97 Bryant, J.P., 55, 84, 97, 114, 169, 170, 182, 258, 267, 493, 501, 593, 599, 604 Bryant, W.S., 189, 195, 196, 201, 204, 205, 218 Bryson, M.I., 39, 97 Bubenzer, G.E., 514, 516 Buchanan, D.J., 365, 367, 382 Buchanan, R.A., 680, 684 Buchmann, N., 287, 306 Buchmann, S.L., 324, 328 Bucjli, P., 390, 391, 395 Buck, J., 645, 655

782 Buckley, G.P., 372, 382 Budd, W.F., 50, 67, 89, 97 Budde, P.E., 90, 119 Budowski, G., 467, 483 Budyko, M.I., 39, 97 Buell, M.F., 340, 357 Bufﬁngton, J.D., 640, 652 Bukkens, S.G.F., 726–728, 730, 731, 737, 738, 745 Bultot, F., 227, 247 Bunnell, F.L., 54, 96 Bunte, K., 513, 516 Bunza, G., 51, 59, 97 Burbaker, H.W., 500, 502 Burbanck, M.P., 590, 604 Burbank, D.H., 202, 215 Burckle, L.H., 42, 112 Burdick, D.M., 343, 344, 353, 354, 357, 359 Burdon, J.J., 530, 539 Buresh, R.J., 527, 540 Burger, A.E., 53, 54, 97, 116, 121 Burger, J., 337, 343, 345, 357 Burgers, C.T., 694, 695, 705 Burges, A., 531, 539 Burgess, A.D., 264, 268 Burgess, D., 545, 547, 550, 551, 554, 567 Burgess, R.L., 592, 606 Burgess, T.L., 324, 326, 415, 416, 446 Burgos, J.J., 323, 326 Burk, C.J., 345, 357 Burkart, S., 491, 492, 494, 502 Burke, I.C., 287, 288, 292, 294–296, 300, 302, 304–306, 557, 558, 560, 561, 563, 566, 567 Burke, M.J.W., 421, 445, 446 Burke, V.J., 682, 684 Burkhardt, J.W., 564, 566 Burnham, K.P., 511, 515 Burnham, R., 142, 145, 157 Burns, B.R., 613, 616, 630 Burns, R.G., 506, 515 Burns, R.M., 164, 165, 182, 470, 471, 483 Burns, S.F., 187, 221, 244, 251 B´urquez, A., 311, 317, 326, 418, 423, 446 Burrows, C.J., 17, 31, 33, 187, 215, 585–588, 596–599, 604, 753, 766 Burtner, D., 42, 99 Burton, B., 25, 36 Burton, D., 67, 116 Burton, H., 64, 100 Burton, H.R., 64, 102 Burton, I., 661, 671 Burton, P.J., 162, 167, 183, 614–616, 619, 623, 624, 626, 627, 628 Burwell, B., 479, 486 Burwell, R.E., 126, 135 Busby, W.H., 474, 483, 642, 643, 653

AUTHOR INDEX Busch, D.C., 427, 447 Busch, D.E., 428, 429, 431, 446, 590, 604 Buschbacher, R., 467, 486 Buschman, D.L., 427, 451 Busdosh, M., 68, 97 Bush, J.K., 598, 609 Bush, M.B., 2, 16, 142, 145, 147–149, 154, 157, 160, 467, 475, 483, 591, 593, 599, 609 Busing, R.T., 190, 193, 205, 216, 222 Busscher, W.J., 507, 509–511, 515–517, 519 Busse, K.K., 457, 459, 465 Bussler, B.H., 588, 604 Butler, J.H., 65, 103 Butler, R.G., 68, 112 Butterﬁeld, R., 480, 483 Butterworth, B.B., 639, 652 Butts, K.H., 390, 394 Byers, E.A., 56, 97 Byington, E.K., 126, 132 Bylund, H., 169, 183 Byres, W.R., 378, 382 Byrnes, W.R., 588, 604 Byttnerowicz, A., 65, 111 Caball´e, G., 234, 241, 248 Cabido, M., 418–420, 447 Cable, D.R., 423, 446 Cabrera, A.L., 495, 501 Cabrera Cano, E., 226, 228, 234, 235, 239, 241, 244, 252 Cadenasso, M.L., 715, 719, 721 Caesar, J.C., 478, 485 Cahoon, D.R., 332, 336, 337, 339, 345, 351, 355, 356, 359, 362 Cain, R.L., 340, 357 Caine, N., 47, 51, 67, 97, 98, 116, 121, 435, 451, 711, 721 Cairns Jr, J., 378, 383 Cairns, J.J., 370, 382 Cajal, J.L., 58, 92, 113 Cal, J., 472, 484 Calabrese, A., 87, 109 Calaby, J.H., 317, 324, 329 Calambokidis, J., 64, 113 Caldwell, B., 457, 464 Caldwell, B.A., 68, 90, 103 Caldwell, M.M., 42, 65, 87, 95, 98, 190, 215, 442, 446 Calkin, P.E., 66, 98 Calkins, D., 390, 391, 395 Calkins, J., 42, 66, 93, 98 Callaghan, T.V., 26, 33, 89, 104, 110, 121 Callaham, L.C., 535, 539 Callaway, R.M., 585, 595–597, 599, 600, 602, 603, 604, 613, 615, 628, 752, 766 Calvo, J.F., 354, 357

Cameron, G.N., 419, 420, 425, 446 Cameron, M.C., 233, 250 Cameron, R.D., 62, 117 Cameron, R.E., 27, 33, 39, 98, 529, 539 Camill, P., 715, 722 Camilo, G.R., 257, 269, 635, 637, 643, 656, 657 Cammack, L.R., 389, 394 Camp, C.R., 507, 516 Camp, P.D., 378, 379, 382 Campbell, A.J., 264, 266, 268 Campbell, B.D., 614, 616, 618, 622, 624, 628 Campbell, I.B., 53, 62, 63, 78, 98, 99 Campbell, J.B., 641, 654 Campbell, J.E., 208, 210, 220 Campbell, R.B., 507, 509, 516, 519 Campbell, R.E., 459, 464 Campbell, R.W., 701, 704 Campbell, T., 662, 670 Campbell, W.J., 66, 67, 102 Campion, M.K., 524, 539 Canaday, B.B., 47, 98 Canaday, C., 473, 483 Canadell, J., 287, 306 Canham, C.D., 187, 189–191, 193, 196, 197, 202–204, 206, 215, 223, 234, 241, 243, 247, 250, 598, 603, 620, 630, 709, 719 Cannell, M.G.R., 453, 464 Cannell, R.Q., 513, 516 Cannon, J.P., 523, 524, 539 Capehart, A.A., 343, 357 Caravello, G.U., 275, 283 Carballas, T., 533, 538 Carey, E.V., 230, 232, 233, 247 Carey, S., 155, 159 Cargill, S.M., 336, 345, 353, 357 Carley, M., 669, 669 Carloss, M., 336, 361 Carlquist, S., 593, 604 Carlson, D.H., 387–390, 396 Carlson, R.W., 623, 624, 628 Carlton, G.C., 198, 215 Carlton, J.T., 436, 447 Carman, J.G., 428, 431, 446 Carolin, V.M., 260, 267 Carpenter, A.T., 525, 527, 537, 539 Carpenter, D.E., 321, 326 Carpenter, R.A., 661, 669 Carpenter, S.E., 591, 604 Carroll, C.R., 261, 267, 690, 696, 699, 704, 704, 705 Carroll, J.B., 636, 652 Carson, C.E., 51, 98 Carson, W.P., 196, 209, 210, 212, 220, 240, 242, 250 Carter, D.L., 506, 507, 513–515, 516, 517

AUTHOR INDEX Carter, J.J., 264, 267 Carter, L.M., 510, 516 Carter, M.R., 510, 516 Carter, R.P., 385, 386, 395 Carter, V., 331, 358 Carter, V.G., 665, 669 Cartwright, B.W., 336, 358 Cary, J.F., 637, 643, 657 Casado, M.A., 472, 484 Casassa, G., 66, 94 Casasso, I., 188, 215 Case, R., 66, 112 Casey, R.J., 646, 652 Casini, S., 85, 101 Cassel, D.K., 510, 514, 515, 516–518 Cassels, D.M., 387, 395 Castaner, D., 345, 358 Castellini, M.A., 68, 107 Castro, M.S., 244, 245, 251, 435, 451 Castro, P.M., 574, 582 Caswell, H., 40, 111, 762, 766 Caswell, M.F., 1, 15 Cates, R.G., 258, 268, 533, 542, 594, 608 Catley, A., 436, 450 Catling, P.M., 434, 446 Catts, E.P., 647, 652 Caughley, G., 69, 98 Cavalieri, D.J., 66, 98 Cawich, A., 472, 484 Cayford, J.H., 164, 183 Cazares, E., 529, 539 Cederlund, G., 170, 183 Cederna, A., 69, 98 Center, D.M., 294, 304 Cern´ık, A., 44, 59, 98 Cernusca, A., 69, 98 Chabreck, R.H., 336, 337, 342, 343, 353, 355, 358, 360, 361, 363 Chadwick, H.W., 589, 604 Chadwick, M.J., 367, 370, 371, 373, 377, 382, 402, 403, 410, 535, 539, 588, 603 Chaieb, M., 555, 566 Challies, C.N., 438, 442, 447 Challinor, J.L., 63, 83, 98, 102 Chaloupka, M., 415, 447 Chambers, A.C., 153, 157 Chambers, J.C., 79, 94, 98, 318, 326, 530 Chambers, M.J.G., 47, 98 Chambers, R., 666, 669 Chanask, D.S., 644, 654 Chanasyk, D.S., 556, 568 Chance, N.A., 57, 98 Chancellor, R.J., 493, 501 Chancellor, W.J., 388, 394 Chandrashekara, U.M., 225, 247 Chaneton, E.J., 294, 304, 498, 502 Chaney, K., 128, 132, 505, 506, 516 Chaney, W.R., 378, 382, 588, 604

783 Chantal, J., 69, 98 Chanway, C.P., 614, 616, 618, 631 Chapin, D.M., 144, 157, 592, 594, 603, 604 Chapin III, F.S., 2, 15, 17, 18, 26, 33, 52, 55, 74, 78–80, 83, 89, 96, 98, 102, 114, 116, 120, 139, 157, 160, 169, 170, 182, 186, 258, 267, 435, 440, 447, 493, 501, 525, 539, 545, 566, 585, 588, 591–593, 596–602, 603, 604, 609, 611, 613–615, 626, 628, 631 Chapin, M.C., 80, 98 Chapman, H.M., 44, 99 Chapman, P.L., 417–419, 421, 450 Chapman, P.M., 87, 98 Chapman, R., 509, 510, 516 Chapman, R.A., 281, 284, 435, 439, 442, 449 Chapman, V.J., 331, 358 Chapman, W.L., 66, 98 Chappellaz, J., 66, 113 Chapuis, J.L., 69, 98 Charlebois, C.T., 64, 99 Charley, J.L., 557, 566 Charlton, C., 664, 670 Charney, J.G., 323, 326 Chase, M.R., 472, 483 Chatarpaul, L., 545, 547, 550, 551, 554, 567 Chaudhry, S., 480 Chazdon, R.L., 202, 215, 220, 223, 243, 247 Cheater, M., 437, 447 Checchini, L., 52, 100 Cheke, A.S., 636, 652 Chemello, G., 65, 101 Chen, J., 715, 719 Chen, T.-H., 188, 201, 216 Chen, T.W., 42, 45, 115 Cheney, N.P., 439, 441, 448 Cheng, G., 66, 103 Cherfas, J., 735, 745 Cherry, B.S.G., 66, 106 Cheseby, M., 66, 97 Chesnais, J.C., 724, 745 Chesson, P.L., 713, 719 Chester, D., 138, 155, 157 Chesterﬁeld, C.J., 681, 684 Chesters, G., 506, 516 Chevennement, R., 140, 142, 157 Chevrier, M., 69, 99 Chew, R.W., 639, 652 Chiba, N., 592, 604 Childers, D.L., 354, 358 Childs Jr, H.E., 639, 654 Childs, R.D., 558, 567 Chinea, J.D., 243, 251 Chinnow, D., 400, 404, 410

Chisholm, A.H., 668, 669 Chmielewski, W., 529, 539 Cho, D.-S., 190, 203, 204, 206, 216 Choi, Y.D., 599, 604 Chow, T.J., 64, 111 Chown, S.L., 89, 99 Christensen, M., 524, 538 Christensen, N.L., 202, 216, 240, 242, 250, 353, 363, 456, 459, 464, 465, 573, 575, 581, 597, 598, 607, 616–618, 625, 628, 630, 713, 719 Christensen, P., 572, 581 Christian, J.M., 623, 628 Christiansen, S., 555, 557, 569 Christie, I., 669, 669 Christy, E.J., 190, 211, 216 Church, M.A., 21, 33 Cicerone, R.J., 64, 96 Cincotta, R.P., 56, 60, 102 Ciudad, A.G., 288, 304 Claassen, V.P., 535, 539 Claesson, P.M., 524, 540 Clampitt, C.A., 153, 158 Clarholm, M., 524, 539 Claridge, G.G.C., 53, 62, 63, 78, 98, 99 Clark, B.C., 640, 652 Clark, D.A., 228, 236, 239, 242, 247, 478, 483 Clark, D.B., 223, 225, 228, 236, 239, 242, 247, 471, 478, 483 Clark, E.H., 131, 132 Clark, F., 522–525, 542 Clark, H.L., 702, 705 Clark, J.S., 174, 183, 193, 194, 216, 456, 464, 615, 628, 717, 719 Clark, K., 225, 227, 228, 231, 240, 242, 244, 251 Clark, M.J., 19, 34 Clark, R., 439, 448 Clark, W.C., 673, 687, 717, 721 Clarke, A., 41–43, 71, 99 Clarkson, B.D., 141–143, 145, 147, 148, 153, 154, 157, 593, 599, 604 Clarkson, B.R., 141–143, 147, 154, 157, 593, 604 Clarkson, R.W., 432, 447 Claupein, W., 510, 516 Clausen, T.P., 170, 182, 533, 542, 594, 608 Clauss, M.J., 579, 583 Clawson, R.G., 467, 472, 485 Clawson, W.J., 678, 685 Claypool, P.L., 297, 303 Clayton, J.S., 433, 452 Clearley, C., 175, 182 Clebsch, E.C., 438, 442, 451 Clebsch, E.E.C., 190, 197, 211, 220 Clebsch, E.E.L., 190, 205, 216 Cleef, A.M., 90, 99

784 Clelland, D.M., 525, 540 Clem, P., 646, 647, 656 Clements, F.E., 2, 15, 31, 33, 319, 326, 585, 586, 591, 593–595, 597–599, 602, 604, 659, 669, 713, 719, 752, 766 Cleveland, C.J., 736, 745, 746 Clifford, H.F., 646, 652 Climie, A., 66, 121 Cline, J.F., 557, 568 Cline, S.P., 173, 184, 211, 216, 218, 711, 720 Clinton, B.D., 190, 197, 203, 206, 208, 216 Clobert, J., 59, 120 Clubbe, C.P., 469, 483 Cluett, H.C., 525, 539 Coates, K.D., 618, 628 Coblentz, B.E., 443, 447, 452 Cochcroft, B., 506, 519 Cochrane, G.R., 52, 99, 438, 447 Cochrane, R., 676, 684 Cody, M.L., 311, 326 Cody, W.J., 415, 416, 421–423, 444, 452 Coe, M., 290, 304 Coe, M.J., 17, 30, 33, 56, 99 Coetzee, M.A.S., 590, 606 Cofﬁn, D.P., 288, 304, 620, 621, 630, 645, 652, 710, 719 Coffman, M.S., 613, 628 Cogbill, C.V., 165, 183, 194, 195, 216, 220 Cogley, J.G., 44, 110 Cohen, A.N., 436, 447 Cohen, J.E., 1, 15, 762, 766 Cohn, J.P., 683, 684 Coile, T.S., 212, 218 Colcote, R.R., 716, 720 Cole, C., 292, 305 Cole, C.V., 291, 292, 304, 306, 524, 539, 555, 566 Cole, D., 563, 566 Cole, D.R., 322, 326 Cole, D.W., 545, 547, 551, 552, 554, 566–569 Cole, E.C., 462, 465 Coleman, D.C., 294, 304, 488, 501, 522, 524, 530, 534, 535, 538, 539, 541, 542, 555, 556, 566, 578, 582 Coleman, R.L., 682, 684 Coley, P.D., 209, 220, 229, 231, 232, 234, 238, 239, 250, 255, 256, 258, 267, 493, 501 Colinvaux, P.A., 467, 475, 483 Collier, B.D., 84, 95 Collins, B., 434, 447 Collins, B.D., 141, 157 Collins, B.G., 674, 684

AUTHOR INDEX Collins, B.S., 187, 190, 198, 202, 208–210, 216 Collins, C.M., 390, 391, 395 Collins, H.P., 510, 518 Collins, N.J., 41, 52, 99 Collins, S.L., 4, 16, 31, 36, 139, 154, 157, 159, 193, 216, 223, 224, 236, 250, 288, 302, 304, 585, 586, 594, 596–598, 602, 607, 608, 611, 630, 679, 686, 709–713, 719, 721 Collins, W.B., 84, 99 Colpaert, J.V., 524, 539 Colwell, R.K., 522, 543 Colwick, R.F., 510, 516 Combs, G.S., 124, 132 Comerford, N.B., 461, 464 Comiso, J.C., 66, 98 Condit, R., 233, 234, 247, 475, 483 Conforti, P., 731, 745 Congdon, R.A., 472, 483 Conlan, K.E., 344, 358 Conn, D.B., 646, 652 Connell, A.D., 64, 115 Connell, J.H., 153, 157, 237, 239, 248, 333, 337, 350, 351, 358, 390, 394, 478, 483, 488, 489, 501, 596–598, 600, 602, 604, 611–613, 621, 626, 628, 694, 704, 757, 762, 766, 767 Conner, K.F., 530 Conner, W.H., 335, 336, 343, 358 Connor, E.F., 598, 602, 604 Connor, K.F., 79, 94 Connor, L.J., 128, 134 Connors, P.G., 64, 96, 419, 421, 449 Conrad, C.E., 272, 284, 575, 582 Conrad, J.C., 387, 388, 390, 391, 394 Conroy, D.W.H., 54, 99 Conroy, J.W.H., 91, 99 Constantinidou, H.A., 528, 539 Cook, A.E., 190, 195, 221 Cook, A.M., 142, 147, 149, 152, 158 Cook, G.D., 288, 304, 681, 684 Cook, J.G., 321, 326 Cooke, J.A., 375, 376, 382, 383 Cooke, R.C., 530, 539, 543 Cooper, A.C., 639, 652 Cooper, J., 69, 70, 99, 120, 436, 441, 443, 449 Cooper, J.C., 697, 704 Cooper, R., 141, 157 Cooper, S.M., 78, 96 Cooper, W.S., 17, 33, 162, 183, 591, 596, 598–601, 604, 755, 766 Cooperrider, A.Y., 675, 686 Copestake, P.G., 55, 91, 99 Copson, G.R., 69, 91, 97, 114 Cordell, C.E., 529, 542 Cordero, R.A., 594, 605

Cordray, S.M., 495, 501 Corke, C.T., 549, 566 Corkidi, L., 532, 539 Corlett, R.T., 471, 483 Cornejo, F., 478, 483 Cornelius, J.D., 387, 396 Cornell, W., 155, 159 Corona, M.E.P., 288, 304 Corte, A., 68, 99 Cortes, M., 17, 27, 36 Corzilius, D.B., 125, 130, 134, 500, 502, 729, 746 Cosentini, C.C., 646, 656 Costa, W.R., 646, 656 Costanza, R., 662, 669, 676, 683, 684, 736, 745 Costin, A.B., 56, 69, 99, 116 Coughenour, M.B., 287, 294, 304, 305, 555, 568 Coull, B.C., 337, 357 Council for Agricultural Science and Technology, 724, 745 Coupland, R., 287, 304 Courtin, G.M., 376, 383 Courtney, S.F., 63, 114 Cousens, R., 615, 628 Coutard, J.-P., 47, 99 Couteaux, M.M., 292, 304 Coutts, M.P., 187, 198, 216 Covich, A.P., 427, 450 Covington, W.W., 457, 462, 464 Coward, L.P., 162, 167, 183 Cowardin, L.M., 331, 358 Cowell, C.M., 675, 684 Cowie, I.D., 443, 447 Cowles, H.C., 595–597, 604 Cowles, S., 87, 89, 111 Cowling, R.M., 271–274, 281, 283, 285, 424, 442, 451, 694, 695, 705 Cox, C.S., 613, 616, 630 Cox, G.W., 529, 540 Cox, J.R., 423, 451 Cox, P.A., 131, 132, 636–638, 653, 655 Cox, S.B., 637, 643, 657, 756, 767 Coxhead, I., 695, 704 Coyne, P.I., 55, 89, 98, 99, 294, 305, 390, 394 Crafford, J.E., 89, 99 Craig, H., 42, 99 Craig, P., 636, 652, 653 Cramer, W., 175, 185 Crams, J.S., 459, 464 Crawford, C.S., 311, 318, 326 Crawford, H.S., 639, 654 Crawford, R., 153, 158 Crawford, R.L., 593, 604 Crawford, R.M.M., 44, 99

AUTHOR INDEX Crawley, M.J., 426, 447, 492, 501, 577, 578, 582, 751, 766 Cress, R.G., 52, 99 Creswell, L.L., 386, 390, 394 Crˆete, M., 54, 55, 99 Criado, B.G., 288, 304 Crisafulli, C., 529–532, 538, 645, 652 Crist, S., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Crist, T.O., 319, 326 Critchﬁeld, W.B., 165, 184 Crocker, R.L., 17, 33, 590, 596, 600, 601, 604, 605 Crofts, K.A., 536, 542 Cromack Jr, K., 26, 31, 33, 173, 184, 211, 218, 524, 540, 711, 720 Crome, F.H.J., 643, 653 Cronk, Q., 419, 420, 450, 758, 766 Cronon, W., 5, 15, 717, 719 Croome, R.L., 53, 99 Crosby, A.W., 2, 15, 154, 157 Cross, A.F., 314, 328 Crossley, D.A., 253, 255, 257, 258, 260–262, 268, 269, 488, 501, 524, 535, 539, 540, 552, 554, 569, 711, 721 Crossley, D.D., 538, 539 Crossley, M.E., 524, 540 Crosson, P., 129, 133 Crouch, H.J., 17, 33, 592, 604 Crowley, D.E., 524, 540 Croxall, J.P., 55, 59, 64, 89, 91, 99, 100 Crozier, C.R., 336, 361 Crump, A., 666, 669 Crutzen, P.J., 64, 95 Cruz, J., 44, 112 Cubbage, J.C., 64, 113 Cueto, L., 479, 486 Cuevas, E., 234, 248 Cuhel, R.L., 42, 105 Culik, B., 62, 70, 100, 121 Cullen, H., 66, 97 Culot, P.H., 492, 493, 495, 499, 501 Culver, D.C., 646, 653 Cumming, D., 290, 304 Cumming, J.R., 525, 540 Cumming, S.G., 162, 167, 183 Cummings, B.J., 666, 669 Cummings, D.L., 235, 251 Cummins, K., 682, 684 Cummins, K.W., 173, 184, 211, 218, 682, 683, 684, 685, 711, 720 Cundell, A.M., 89, 90, 100 Cunningham, G.L., 322, 328 Cunningham, G.M., 317, 324, 326 Curatolo, J.A., 62, 100 Curl, E.A., 523, 530, 540 Curran, M., 64, 100 Curran, W.J., 169, 186

785 Curtin, C.G., 590, 604 Cushman, D., 343, 353, 354, 357 Czuhai, E., 639, 654

Da Silva, J.M.C., 476, 485 Daae, F.L., 522, 543 Daehler, C.C., 336, 358 Dahaban, Z., 471, 473, 482 Dahir, S.E., 197, 216, 229, 235, 240, 248 Dahl, E., 41, 75, 100 Dahl, J.F., 636, 652 Dahlgren, M.A.G., 524, 540 Dahlgren, R., 151, 160, 545, 566 Dahlgren, R.A., 211, 212, 222 Dahlman, R.C., 641, 654 Dahm, C.N., 522, 528, 529, 538, 682, 684 Daiber, F.C., 336–338, 343, 347, 348, 358 Daily, G.C., 366, 382, 673, 686, 724, 734, 738, 745, 746 Dalal, R.C., 510, 516 Dale, J.E., 336, 361 Dale, M.B., 29, 33 Dale, M.R.T., 17, 33, 141, 157, 591, 599, 600, 603 Dale, P., 612, 631 Dale, T., 665, 669 Dale, V.D., 142, 145, 152–154, 157 Dale, V.H., 590, 591, 598, 603, 604, 716, 718, 721, 722 Dale, W.L., 228, 248 Dalke, P.D., 589, 604 Dallinga, J.H., 497, 501 Dalluge, R.W., 65, 113 Daly, H.E., 738, 745 Damman, A.W.H., 170, 183 Dan, J., 273, 275, 284 Danell, K., 169, 183, 579, 582 Danfeld, R., 70, 121 D’Angela, E., 420, 447 Daniel, P.E., 496, 501 Daniels, M.B., 512, 518 Daniels, R.B., 514, 516 Daniels, R.J.R., 472, 483 Daniels, W.L., 377, 378, 382 Danin, A., 316, 326, 589, 591, 594, 604, 606 Dansereau, P.R., 169, 175–177, 182, 183 D’Antonio, C.M., 278, 283, 413, 421–423, 425, 426, 431, 435, 440, 445, 447, 611, 628, 681, 684 Darby, H.C., 460, 464 Darby, M., 535, 539 Dargavel, J., 675, 684 Darley, W.M., 336, 361 Darmody, R.G., 51, 118 Darwin, C., 523, 540 Daschbach, N., 638, 653

Daubenmire, R., 288, 301, 304, 574–576, 582, 640, 653 Daugherty, C.H., 681, 687 Daugherty, D., 548, 554, 567 Davenport, J., 42, 96 Davenport, M.L., 312, 328 Davey, M.C., 26, 33, 591, 604 David, M.B., 457, 464 Davidson, A., 56, 100 Davidson, A.T., 42, 100, 110 Davidson, D.W., 169, 170, 183, 261, 267 Davidson, W., 472, 484 Davies, D.B., 682, 686 Davies, J.A., 506, 515 Davis, B.N.K., 371, 372, 383, 384, 599, 604 Davis, C.B., 332, 333, 336, 340, 348, 352, 363 Davis, F.W., 585, 604 Davis, G.E., 195, 220, 354, 361 Davis, J.J., 64, 114 Davis, J.R., 574, 583 Davis, M.B., 175, 183, 187, 191, 193, 213, 214, 215, 216, 716, 720 Davis, P.A., 64, 66, 116 Davis, P.B., 70, 100 Davis, R.W., 68, 107 Davis, S.M., 683, 685 Dawson, F.H., 433, 434, 447 Dawson, T.E., 52, 100 Day Jr, J.W., 352, 362 Day, T.A., 88, 103, 140, 158, 560, 566, 599, 604 Day, T.D., 294, 304 Dayton, P.K., 59, 100 De Bie, S., 497, 501 de Cock, N., 593, 609 de Grandpr´e, L., 175, 178, 183 de Groot, W.J., 163, 186 De Jong, D.J., 340, 352, 358 de Jong, T.M., 589, 594, 603 de la Cruz, A.A., 337, 342, 353, 358, 359 de la Fuente, E., 496, 501 De Leeuw, J., 340, 358 de Lima, O., 233, 249 De Luce, J., 674, 684 de Margerie, S., 62, 110 de Mol, F., 507, 516 de Mora, S.J., 43, 115 de Nicol´as, J.P., 147, 158 de Oliveira, P.E., 467, 483 de Orellana, J.A., 496, 502 de Ruiter, P.C., 522, 535, 542 De Seze, E., 189, 199, 216 de Silva, F., 66, 101 de Soyza, A.G., 311, 314, 326, 328 De Steven, D., 614, 616, 619, 628 de Vries, Y., 348, 357, 497, 501

786 de Wildt, E., 593, 609 de Wit, C., 292, 293, 304 Deacon, G.E.R., 39, 100 Deacon, H.J., 280, 283 Dealdana, B.R.V., 288, 304 Dean, W.R.J., 645, 653 DeAngelis, D.L., 4, 16, 597, 607, 678, 685, 712, 721, 725, 746 DeAngulo, J., 512, 518 Dearing, M.D., 55, 100 DeBano, L.F., 459, 464, 575, 582 deBary, A., 523, 526, 540 DeBoer, A.J., 662, 670 Debourdieu, J., 200, 216 Debussche, M., 427, 430, 452 DeCamps, H., 427, 429, 430, 450 Dee, P.E., 388, 390, 395 Deegan, L.A., 354, 358 DeFerrari, C., 427, 429, 430, 450 DeFerrari, C.M., 419, 420, 427, 447 DeGraaf, R.M., 714, 719 del Moral, R., 28, 33, 52, 121, 137, 139, 140, 142–145, 147–154, 158, 160, 590, 591, 599, 604, 610, 613, 629, 713, 720 del Pozo, A., 279, 280, 282, 284 DeLappe, B.W., 65, 114 DeLaune, R.D., 336, 349, 361 Delaune, R.D., 345, 358 Delcourt, H.R., 493, 501 Delcourt, P.A., 493, 501 Delille, D., 89, 100 Dell, B., 282, 283, 524, 542 DeLuca, T.H., 694, 705 DeLuisi, J.J., 65, 95 Delwaide, A., 169, 175, 185, 189, 193, 220 Delwiche, C., 523, 543 DeMars, B.G., 593, 603 deMenocal, P., 66, 97 Demeo, T.E., 169, 182 den Hartog, C., 403, 410 Denayer-de Smet, S., 399, 410 Deneke, F.J., 390, 394 Denevan, W.M., 471, 479, 483 Denny, C.S., 209, 210, 216 Denslow, J.S., 189–191, 202, 206, 213, 215, 216, 224, 225, 228, 229, 232, 234, 236, 240, 242–244, 247–249, 252, 333, 350, 358, 471, 472, 483, 642, 653, 709, 716, 719, 720 Densmore, R.V., 79, 100 Dent, D.L., 21, 23, 24, 33, 36 Department of Defense, 391, 394 Department of the Air Force, 391, 394 Department of the Army, 387, 394 Derenne, P., 69, 108 Deshler, T., 65, 105 Desideri, P.G., 52, 100

AUTHOR INDEX Desmet, P., 513, 516 Despain, D.G., 165, 183, 459, 466 Desponts, M., 161, 164, 165, 169, 175, 183, 185 Despres, F., 189, 194, 216 Dessens, J., 196, 200, 216 Detling, J.K., 288, 294, 295, 301, 304, 306, 546, 557, 558, 560, 566, 567, 580, 583 Detling, J.W., 557, 568 DeVries, A.L., 43, 100 Dexter, A.R., 507, 519 DeYoung, C.A., 387, 394 Dezzeo, N., 225, 227, 228, 231, 240, 242, 244, 251 Dhillion, S., 676, 684 Dhillion, S.S., 645, 647–649, 653, 654 di Castri, F., 271, 274, 283 Diaz, A., 642, 643, 657 Diaz, R.A., 496, 501 Diaz, R.J., 343, 358 D´ıaz, S., 418–420, 447 Dick-Peddie, W.A., 323, 326 Dickinson, A.B., 58, 115 Dickinson, M.B., 235, 245, 252 Dickinson, M.D., 226, 232–234, 248 Dickson, B.A., 590, 605 Dickson, J.F., 618, 628 Dieckmann, G., 89, 117 Dieckmann, G.S., 50, 105 Dieleman, P.J., 509, 517 Diem, H.D., 506, 516 Diemer, M., 55, 89, 96, 107 Diersing, V.E., 385, 386, 388–390, 394, 396, 416, 451 Dietz, R., 64, 100 Dighton, J., 532, 544 Dijkema, K.S., 337, 358 DiLabio, R.N.W., 337, 358 Dillenius, J.J., 495, 501 Ding, Y.H., 50, 114 Dinkelaker, B., 524, 540 Dinkins, W.C., 387, 394 Dinneford, B., 593, 605 Dionne, J.C., 336, 358 Dirzo, R., 714, 721 Dixon, J.S., 645, 653 Dixon, K.W., 680, 685 Dmuchowski, W., 65, 111 do Amaral, I.L., 225, 228, 250, 471, 485, 713, 721 Doake, C.S.M., 66, 100 Dobran, F., 155, 158 Dobson, A., 444, 447 Docters van Leeuwen, W.M., 137, 158 Dodd, J.L., 288, 304, 557, 567 Doelman, P., 661, 671 Doescher, P.S., 557, 566

Doiban, V.A., 57, 62, 63, 65, 100 Doing, H., 589, 605 Dolan, R., 441, 447 Dollar, K.E., 334, 358 Dollenz, O., 188, 215 Domarr, J.F., 644, 645, 654 Domm, S., 415, 447 Dommergues, Y.H., 506, 516 Domsch, K.H., 531, 533, 541 Donahue, R.H., 126, 133 Donahue, R.L., 126, 127, 129, 135 Donat, J., 507, 516 Donner, J., 174, 183 Donoso, C., 208, 222 D¨oo¨ s, Bo.R., 667, 669 Doran, J.W., 525, 540 Doran, P.T., 43, 94 Doren, R.F., 419, 420, 439, 441, 447 Dormaar, J.F., 555–557, 566, 569 Dorney, J.R., 190, 196, 197, 205, 208, 209, 216, 228, 242, 248, 398, 411, 713, 720 Doty, C.W., 510, 515 Doube, B.M., 646, 647, 653 Doucet, G.J., 55, 99 Douglas, E.L., 42, 104 Douglas, G.W., 52, 100 Douglas, I., 475, 483 Douglas, J.T., 510, 516 Douglas, M., 66, 112 Downing, J.R., 130, 135 Downs, A.T., 392, 396 Doyle, G., 39, 120 Doyle, T.W., 233, 235, 244, 245, 251, 333, 334, 336, 345, 347, 348, 352, 355, 358, 362 Draaijers, G.P.J., 715, 720 Drake, B.G., 322, 328 Drake, D.R., 142, 158 Drake, J.A., 278, 284 Drake, N.E.R., 333, 350, 361 Drawe, D.L., 337, 356 Drayton, B., 404, 410 Dregne, H.E., 126, 133, 312, 324, 329 Dreimanis, A., 19, 33 Drescher, H.E., 64, 115 Dreux, P., 69, 100 Drew, M.C., 512, 517 Drewry, D., 18, 34 Dreyer, G.D., 595, 598, 607 Driscoll, C.T., 646, 656 Drost, H.J., 352, 362 Drury, W.H., 595, 598, 605, 611, 621, 629 Druzhinina, O.A., 68, 100 Dryer, V., 229, 249 du Toit, P., 589, 607 Dubois, K.E., 76, 101 Dubois, P.C., 386–390, 396 DuBois, R., 472, 484

AUTHOR INDEX Dubos, R., 467, 469, 476, 483 Dubuc, M., 161, 168, 172, 175, 182 Duckham, F., 366, 383 Dudal, R., 512, 516 Dudley, L.M., 523, 524, 539, 541 Dudley, T., 434, 447 Dudley, T.L., 426, 428, 431, 432, 441, 447 Dueck, T.A., 530, 540 Duever, M., 336, 346, 360 Duever, M.J., 190, 195, 199, 200, 210, 216, 335, 336, 358 Duky, F.L., 130, 135 Duman, J.G., 43, 96, 100 Dumas, A.R., 125, 130, 134, 500, 502, 729, 746 Dunbar, M.J., 42, 64, 100 Dunlap, W.C., 42, 87, 106 Dunn, C.P., 190, 196, 197, 205, 208, 209, 216, 228, 242, 248, 337, 349, 358, 713, 720 Dunne, K.P., 187, 216 Dunne, T.D., 141, 157 Dupˆaquier, J., 662, 669 Durall, D., 524, 543 Durrell, L.W., 90, 116, 530, 540 Durrieu de Madron, L., 230, 248 Dutton, E., 65, 95 Duvigneaud, P., 399, 410 Duxbury, J.M., 525, 540 Dyck, W.J., 461, 464 Dye, P.J., 430, 447 Dyearsness, C.T., 211, 217 Dyer, C., 472, 484 Dyer, M.I., 293, 295, 306, 557, 566, 578, 582 Dyke, C., 740, 745 Dykyjova, D., 331, 363 Dymond, J.R., 124, 133 Dynesius, M., 188, 193, 194, 216, 218 Dyurgerov, M.B., 66, 100 Dzurec, R.S., 442, 446 Eakin, M., 599, 603 Early, M., 593, 596, 605 East, R., 288, 290, 304 Easter, M.J., 190, 202, 205, 216, 220 Eastham, J., 681, 686 Eaton, J.S., 547, 554, 567 Eaton, R.D., 64, 100 Ebelhar, S.A., 126, 133 Eberhard, I.E., 69, 97 Ecclestone, M., 43, 94 Eckert, G.E., 696, 704 Eckhart, V., 435, 440, 447 Eckholm, E.P., 692, 704 Edenius, L., 169, 170, 183, 579, 582 Edmonds, R.L., 664, 669 Edmunds Jr, G.F., 257, 258, 267

787 Edwards, C.A., 131, 133 Edwards, D.W., 264, 267 Edwards, F.S., 522, 528, 529, 538 Edwards, J.A., 41, 100 Edwards, J.H., 511, 517 Edwards, J.M., 337, 358 Edwards, J.S., 53, 100, 150, 151, 153, 158, 593, 604, 605 Edwards, N.T., 546, 552–554, 566, 567 Edwards Jr, T.C., 390, 395 Eggler, W.A., 141, 142, 144, 145, 153, 158, 587, 605 Eggleton, P., 474, 483 Egler, F.E., 595, 598, 605, 607 Ehleringer, J.R., 287, 306, 311, 312, 326, 562, 564, 565, 566 Ehlers, W., 507, 510, 516 Ehnstr¨om, B., 161, 163, 166, 167, 169, 173, 182, 183 Ehrenfeld, J.G., 190, 193, 197, 208, 212, 216, 345, 350, 358, 433, 434, 447 Ehresmann, J., 390, 395 Ehrlich, A.H., 1, 16, 673, 687, 723, 724, 738, 745, 746 Ehrlich, P.R., 1, 16, 673, 681, 686, 687, 723, 724, 734, 738, 745, 746 Ek, E.R., 545, 547, 549, 551, 566 Ek, R.C., 478, 485 Ekelund, N.G.A., 66, 100 Ekvall, R.B., 56, 60, 100 Ekwue, E.I., 506, 516 El-Atrach, F., 525, 539 El-Baz, F., 388, 390, 394 El-Sayed, S.Z., 42, 66, 87, 100, 105 El-Swaify, S.A., 124, 133 Eleuterius, L.N., 340, 358 Eliaˇs, P., 405, 406, 410 Elisseou, G.K., 294, 305 Elkins, C.B., 511, 516 Elkins, J.W., 65, 103 Elkins, N.Z., 526, 536, 540 Ellenberg, H., 59, 100, 228, 229, 248, 408, 410 Eller, B.M., 311, 329 Ellery, W.N., 379, 383 Elliot, C.C.H., 337, 360 Elliot, E.T., 294, 304 Elliot, R., 675, 685 Elliott, E.T., 524, 539 Ellis, B.A., 523, 540 Ellis, D.F., 127, 133 Ellis, D.V., 369, 383 Ellis, J., 324, 326 Ellis-Evans, J.C., 43, 54, 91, 100, 101 Ellison, A.M., 333, 336, 338, 345, 348–351, 357, 358, 361 Ellison, L., 557, 566 Ellsworth, D.S., 477, 483

Elmqvist, T., 131, 132, 636–638, 653, 655 Elton, C.S., 129, 133 Elven, R., 17, 20, 26, 28–30, 34 Elvers, H., 403, 411 Elwell, H.A., 126, 133 Emanuelsson, U., 62, 101 Emers, M., 76, 101 Emmingham, W.H., 618, 628 Emmons, L.H., 529, 540 Emslie, S.D., 89, 91, 101 Encyclopædia Britannica, Inc., 225, 226, 248 Endler, J.A., 710, 720 Endo, A.S., 388, 394 Eng, R.L., 390, 395 Engelhardt, F.R., 68, 90, 101 Engelmark, O., 161, 164, 166, 168–170, 174, 175, 177, 179–182, 183, 184 England, J., 69, 101 Engle, D.M., 297, 303 Englerth, G.H., 229–232, 252 Ennos, A.R., 198, 217 Enokido, M., 65, 118 Enzenbacher, D.J., 69, 101 Epstein, H.E., 296, 300, 305 Epstein, H.W., 288, 304 Erdman, J.E., 90, 116 Erickson, A.W., 59, 101 Erickson, H.E., 577, 580, 583 Erickson, L.E., 369, 384 Erickson, P., 65, 96 Erickson, P.M., 369, 383 Erickson, W.R., 198, 221 Ericson, L., 161, 163, 166, 167, 169, 173, 183, 579, 582 Ericsson, A., 166, 184 Eriksson, K., 170, 183 Erkkil¨a, R., 408, 410 Ernest, K.A., 319, 326 Ernst, A., 45, 115 Ernst, W., 65, 108 Ernst, W.H.O., 374, 383 Errington, P.L., 645, 653 Erwin, W.J., 639, 640, 653 Esler, K.J., 272, 283, 320, 329 Eslinger, D.H., 321, 329 Espejel, I., 532, 543 Espeland, M., 43, 55, 112 Espinosa, G., 279, 283 Espinoza, M., 480, 483 Esseen, P.A., 161, 163, 166, 167, 169, 173, 183, 184 Estbergs, J.A., 429, 431, 441, 446 Estes, J.E., 1, 15 Esteve, M.A., 354, 357 Estrada, C., 243, 251 Etheridge, D.M., 66, 67, 101, 111 Etienne, M., 276, 279, 280, 282, 283

788 Eubanks, C.S., 64, 112 Evans, C.C., 575, 582 Evans, D.G., 527, 542 Evans, E.W., 641, 653 Evans, J., 469, 470, 477–480, 483 Evans, M., 680, 685 Evans, M.E., 371, 384 Evans, R.A., 423, 452, 557, 569 Evans, R.D., 562, 564, 565, 566 Evans, S., 680, 685 Eve, M.D., 312, 328 Eveleigh, D.E., 599, 607 Evenari, M., 309, 312, 319, 320, 323, 326 Evenari, M.L., 536, 540 Evenson, J.P., 573, 582 Everett, K.P., 53–55, 110 Everett, K.R., 43, 45, 47, 65, 66, 68, 75, 101, 106, 119 Everham III, E.M., 187, 199, 201, 214, 217, 223, 224, 228–231, 235–243, 248, 252 Everitt, E.L., 428, 431, 447 Eviner, V.T., 545, 566 Evinger, W.R., 386, 394 Ewel, J., 239, 248, 459, 465 Ewel, J.J., 336, 351, 352, 358, 477, 483, 679, 685 Ewel, K.C., 336, 349, 355, 358 Eyles, N., 19, 33 Ezcurra, E., 599, 609

Facelli, J., 287, 305 Facelli, J.M., 288, 304, 420, 447, 498, 501, 614, 615, 629 Fahey, D.W., 64, 112 Fahey, T.J., 213, 218 Fahrig, L., 718, 720 Fairney, A., 42, 103 Falinska, K., 351, 359 Falinski, J.B., 188, 199, 210, 217, 405, 410, 709, 720 Falkenmark, M., 126, 133 Faller, A., 142, 147, 159 Fallis, B.W., 64, 96 Fand, H., 129, 134 Fanning, P., 443, 447 Fantechi, R., 276, 283 FAO, 453, 465, 470, 483, 489, 501 Farant, J.P., 64, 100 Farentinos, R.C., 529, 541 Fares, Y., 258, 267 Farina, A., 276, 283 Fari˜nas, M.R., 477, 485 Farman, J.C., 65, 101 Farnsworth, E.J., 348, 358 Farnworth, E.A., 476, 483 Farrington, P., 282, 284

AUTHOR INDEX Fastie, C.L., 2, 15, 17, 18, 33, 34, 139, 157, 588, 590–592, 594, 598–602, 604–606, 613, 628 Faulkner, H., 664, 669 Faulkner, S.P., 342, 353, 359 Fa´undez, L., 279, 283 Favarger, C., 42, 101 Fay, P.A., 639, 640, 653 Fayle, D.C.F., 174, 185 Feagley, S., 352, 362 Fearnside, P., 691, 704 Feder, W.A., 346, 361 Federer, C.A., 547, 566 Federle, T.W., 68, 90, 101 Feeny, P.P., 258, 267 Fehsenfeld, F.C., 64, 101, 112 Feinsinger, P., 474, 483, 642, 643, 653 Feistner, A.T.C., 681, 686 Felger, R.S., 323, 327 Felix, N.A., 76, 101 Feller, I.C., 336, 341, 355, 359 Felli, O.M., 496, 502 Fender, W.M., 524, 540 Fenner, M., 593, 594, 605, 622, 624, 629 Fenner, M.W., 350, 359 Fenner, R.L., 639, 654 Fenton, E.W., 59, 101 Fenton, J.H.C., 41, 101 Fergus, I.F., 594, 609 Ferguson, C.W., 592, 605 Ferguson, K.D., 369, 383 Ferm, L.M., 91, 96 Fernald, M.L., 52, 101, 590, 605 Fernandez, C., 343, 359 Fern´andez, D.S., 239, 243, 248, 594, 605 Fern´andez-Palacios, J.M., 147, 158 Ferreira, M.T., 430, 447 Ferreira, S.M., 379, 380, 383, 384 Ferris-Kaan, R., 675, 685 Fetcher, N., 90, 101, 124, 135, 223, 239, 243, 247, 248, 590, 592, 594, 605, 609, 717, 722 Ffolliott, F., 459, 464 Fægri, K., 17, 29, 30, 34 Fiard, J.P., 142, 143, 159 Fidelibus, M.W., 535–537, 539–541 Field, W.O., 600, 605 Fielder, P., 675, 686 Figueroa, C., 524, 527, 538 Filion, L., 169, 175, 185, 189, 193, 220 Finck, E.J., 639, 640, 654 Finegan, B., 598, 605, 611, 613, 621, 629 Finkl, C.W., 2, 15 Finlay, R.D., 526, 537, 542 Finley Jr, R.B., 645, 653 Finn, J.T., 39, 40, 111, 473, 486 Finney, B., 66, 112 Finzi, A.C., 202, 215

Firestone, M.K., 293, 294, 304, 507, 518, 525, 541, 542 Fischer, A., 188, 194, 197, 217 Fischer, B.C., 208, 218 Fischer, C.R., 472, 483 Fisher, D.A., 66, 107 Fisher, F.M., 645, 654 Fisher, R.F., 454, 465 Fisher, S.G., 427, 447, 451, 708, 715, 720, 722 Fiske, R.S., 2, 16, 587, 608 Fitch, H.S., 645, 653 Fitter, A., 17, 34 Fitton, L., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Flaccus, E., 590, 605 Flach, K., 292, 304 Flanders, N.E., 58, 101 Flannigan, M.D., 164, 166, 169, 174, 175, 182, 183 Fleeger, J.W., 345, 358 Fleming, R.A., 174, 175, 183 Fletcher, J.E., 535, 540, 562, 566 Fletcher, N., 211, 222 Flint, S.D., 87, 95 Flohn, H., 50, 101 Flores, J.S., 467, 483 Flores Paitan, S., 471, 479, 483 Floresroux, E.M., 421, 422, 446 Floret, C., 325, 325, 436, 446, 674, 676, 684 Flower, D.A., 65, 115 Fluet, M., 193, 195, 198, 215, 225, 247, 471, 482, 713, 719 Flynn, K.M., 342, 344, 345, 349, 359 Focardi, S., 64, 65, 85, 87, 95, 101 Fogel, R., 524, 540 Fogg, G.E., 53, 101 Foley, C.C., 387, 388, 390, 391, 394 Follett, R.F., 126–129, 133, 556, 557, 566 Follett, R.H., 126, 133 Follmann, E.H., 58, 70, 101 Fonda, R.W., 47, 98, 590, 593, 594, 599, 605, 606 Fondahl, G., 62, 101 Fonteyn, P.J., 313, 328 Food and Agriculture Organization of the United Nations, 723, 729, 745 Foord, S.H., 379, 383 Foote, A.L., 336, 339, 345, 351, 355, 356, 359 Forbes, B.C., 62, 75, 76, 78, 79, 101, 589, 605 Forcella, F., 415, 416, 420, 447, 488, 491, 501 Ford, H.A., 264, 267 Ford, M.A., 333, 342, 344, 345, 351, 359 Ford-Robertson, F.C., 469, 470, 483

AUTHOR INDEX Forman, R.T.T., 709, 711, 713, 715, 718, 719, 720, 762, 766 Forman, S.L., 56, 120 Formichi, P., 64, 87, 95 Forthman, C.A., 353, 359 Fosberg, R.F., 147, 158 Fossi, C., 85, 101 Foster, B.L., 619, 620, 629 Foster, D.R., 187, 189, 191, 193–196, 198–201, 214, 215, 217, 225, 229, 230, 238, 239, 242, 247, 248, 471, 482, 713, 719 Foster, G.R., 123, 124, 129, 133, 513, 514, 517, 518 Foster, J., 432, 451 Foster, J.R., 190, 202, 217 Foster, N.W., 549, 566 Foster, R., 260, 269 Foster, R.B., 224, 225, 229–231, 233–235, 242, 247–249, 252, 260, 269, 475, 483 Foster, R.C., 525, 542 Foster, W.P., 130, 133 Fountain, M.C., 436, 447 Fournet, S., 336, 339, 345, 351, 355, 356, 359 Fowler, B., 65, 96 Fowler, N., 493, 501 Fownes, J.H., 461, 462, 466 Fox, B.J., 281, 282, 283, 713, 720 Fox, L.R., 255, 267 Fox, M.D., 281, 282, 283, 713, 720 Fox, R.J., 64, 113 Fraley Jr, L., 555, 565 Frampton, C.M., 681, 686 Franceschini, G.A., 42, 105 Francis, G.R., 682, 683, 685 Francis, P., 137, 138, 156, 158 Francis, R., 526, 537, 542 Francou, B., 47, 99 Frangi, J.L., 238, 239, 242–245, 248 Frank, A.B., 556, 557, 566 Frank, D.A., 289, 290, 294–296, 304, 305 Franke, A., 87, 110 Franke, C., 390, 395 Frankenberger Jr, W.T., 514, 515 Frankland, J.C., 531, 540 Franklin, J., 196, 217, 454, 462, 465 Franklin, J.F., 137, 140, 142, 144, 151, 152, 158, 173, 184, 190, 198, 202, 203, 211, 216–218, 221, 229, 251, 464, 465, 641, 656, 711, 715, 716, 718, 719, 720 Franz, E.H., 39, 40, 111, 143, 153, 158, 159 Franz, H., 44, 51, 52, 101 Franzmeier, D.P., 514, 516 Fraser, W.R., 91, 101 Fraver, S., 235, 248 Frear, D.E.H., 65, 102

789 Freckman, D.W., 522, 535, 540, 542, 578, 582 Frederick, J.E., 93, 101 Fredriksen, R.L., 639, 654 Free, G.R., 513, 518 Freedman, B., 20, 33, 385, 394 Freeman, C.C., 639, 653 Freeman, M.M.R., 58, 101 Freese, D., 388, 390, 394 Frei, A., 67, 102 Freifelder, R., 440, 447 Frelich, L.E., 189, 192, 193, 203, 214, 217, 219, 716, 720 French, D.D., 74, 102 French, D.P., 64, 113 French, R.A., 460, 465 Frenkel, R.E., 279, 283 Frenot, Y., 23, 34, 69, 83, 99, 102, 118 Frenzen, P.M., 137, 139, 142, 144, 145, 148, 158, 592, 605 Fresh, K.L., 337, 362 Freskos, S., 226, 229, 231, 249 Frey, R.W., 336, 337, 358, 360 Freydin, I.L., 65, 102, 117 Fridriksson, S., 83, 92, 102, 142, 143, 147, 148, 151, 154, 158, 591, 605 Friedel, H., 17, 25, 34 Friedel, M.H., 546, 567, 675, 686 Friedman, J.M., 428, 451, 590, 599, 605 Friedmann, E.I., 45, 51, 102, 111 Friend, J.A., 681, 685 Friese, C.F., 528–533, 538, 540, 645, 647, 648, 652, 653 Frisancho, A.R., 44, 102 Frisby, K., 55, 114 Frisque, G., 161, 182 Frissell Jr, S.S., 193, 217 Froend, R.H., 681, 682, 685, 686 Froidevaux, L., 65, 115 Frost, K.A., 346, 361 Frost, P., 291, 304 Frumhoff, P.C., 473, 483 Frye, R.J., 590, 593, 605 Frye, W.W., 126, 133 Fryer, J.D., 493, 501 Fryrear, D.W., 514, 516 Fu, S., 238, 248 Fuentes, E., 279, 280, 283 Fuentes, E.R., 279, 283 Fuentes, M., 240, 251 Fuenzalida, P.H., 323, 326 Fuglsang, A., 682, 687 Fujita, N., 198, 218 Fujita, T.T., 196, 217 Fuller, D.O., 475, 483 Fuller, F., 525, 542 Fuller, J.L., 758, 766 Fulton, R.W., 387, 394

Funtowicz, S., 662, 669, 670 Funtowicz, S.O., 725, 741, 745 Furbish, C.E., 345, 351, 359, 438, 447 Furniss, R.L., 260, 267 Furrer, G., 57, 97 Furukawa, A., 589, 606 Furyaev, V.V., 161, 168, 183, 184 Futuyma, D.J., 259, 267 Gaare, E., 43, 55, 112 Gabrielson, F.C., 616, 622, 624, 630 Gade, D.W., 419, 420, 447 Gadgil, M., 472, 483 Gagen, P., 373, 383 Gaggi, C., 65, 101 Gagnon, D., 175, 178, 183 Gagnon, J., 164, 183 Gagnon, R., 174, 185 Gahler, A.R., 546, 566 Gajewski, K., 66, 112 Gale, M.R., 615, 629 Gallagher, G.J., 188, 194, 217 Gallepp, G., 524, 541 Gallimore, R.G., 196, 200, 217 Galloway, R.L., 366, 383 Galvin, K.A., 324, 326 Gamboa, R., 44, 112 Gambrell, R.P., 337, 359 Gams, H., 79, 102 Ganga-Visalakshy, P.N., 442, 449 Ganio, L.M., 255, 257, 269 Gannon, J.T., 530, 540 Gannon, M.R., 637, 638, 643, 653, 656, 657, 752, 767 Ganry, F., 506, 516 Ganskopp, D.C., 555, 557, 567 Garbaye, J., 525, 540 Garcia, R., 496, 501 Garcia-Barrios, R., 477, 482 Garc´ıa Ciudad, A., 294, 305 Garc´ıa Criado, B., 294, 305 Garcia-Moya, E., 574, 582 Garcia-Perez, J.D., 664, 670 Gardener, J.H., 380, 381, 382 Gardiner, B.G., 65, 101 Gardiner, R., 661, 669 Gardner, B.D., 64, 115 Gardner, I.C., 525, 540 Gardner, J.S., 20, 35 Gardner, L.R., 195, 200, 201, 217 Gardner, R.H., 716, 718, 722, 764, 767 Garﬁnkel, H.L., 66, 102 Garnick, E., 131, 134 Garnier, E., 623, 629 Garofalo, D., 336, 359 Garrett, H.G., 334, 358 Garrick, R.C., 64, 102 Garrison, R.W., 636, 653, 752, 766

790 Garrity, D.P., 527, 540 Garrity, S.D., 337, 359 Gartner, B.L., 79, 80, 102, 116 Garwood, N.C., 236, 237, 240, 248, 590, 605, 717, 720 Garza, N.E., 561, 569 Gathen, K.L., 333, 336, 341, 345, 348, 355, 357 Gaudet, C.L., 623, 624, 629 Gauthier, G., 560, 567 Gauthier, S., 164, 165, 183, 184 Gavin, T.A., 390, 395 Gavrilyuk, V.I., 532, 544 Gecy, J.L., 139, 158 Gee, G.W., 504, 516 Geertz, C., 660, 665, 670 Geiger, R., 715, 720 Gellatly, A.F., 24, 34 Gemma, J.N., 153, 159, 528, 540 Gemmell, R.P., 592, 603 Gemmill, B., 417, 446 Genskow, K.D., 390, 395 Gentry, A.H., 229, 234, 250, 467, 483 George, E., 524, 541 George, J.L., 65, 102 George, M.R., 678, 685 George, R.J., 282, 284, 682, 685 George, T., 527, 540 Georgiadis, N.J., 311, 327 Gepp, J., 407, 410 Geraci, J.R., 90, 102 Gerard, C.J., 507, 516 Gerdol, V., 343, 359 Gereben, B.A., 26, 34 Gerresheim, K., 556, 567 Gerry, A.K., 619–621, 623, 624, 629 Gersper, P.L., 63, 83, 98, 102 Gerwing, J., 478, 486 Gese, E.M., 387, 395 Gessel, S.P., 551, 566 Getz, L.L., 387, 395 Geyer, M.A., 618, 629 Ghersa, C.M., 491–495, 498, 499, 501, 502 Ghimire, K.B., 662, 670 Ghoda, A., 65, 118 Giacomin, F., 275, 283 Giampietro, M., 725–733, 735–738, 745, 746 Gibbens, R.P., 318, 327 Gibbons, J.W., 334, 362, 682, 684 Gibson, A.C., 278, 285, 316, 317, 328, 329 Gibson, A.D., 589, 608 Gibson, C.W.D., 593, 603 Gibson, D.J., 288, 304, 377, 383, 639, 641, 644, 645, 653 Gibson, J., 472, 484

AUTHOR INDEX Gibson, J.A.E., 64, 102 Gibson, N., 69, 102, 188, 211, 217 Gichuki, F., 662, 668, 671 Gies, H.P., 66, 114 Gifford, F.G., 563, 567 Gifford, G.F., 563, 567 Gifford, R.O., 506, 518 Gil, A., 495, 501 Gilbert, A.J., 661, 670 Gilbert, O.L., 397, 403, 405, 406, 408, 410 Gilbert, R., 44, 102 Gilbert, R.O., 557, 568 Gilichinsky, D., 27, 34 Giliomee, J.H., 281, 283 Gill, A.M., 439, 441, 448, 572, 582 Gill, D.S., 616, 617, 619, 627, 629, 631 Gill, W.R., 507, 516 Gillespie, A.J.R., 230, 232, 233, 247 Gillespie, B.M., 388, 395 Gillham, D.A., 680, 686 Gillham, M.E., 55, 102, 336, 337, 359 Gilliam, F.S., 478, 485 Gillon, Y., 639, 653 Gilman, R.D., 507, 515 Gilpin, M.E., 674, 686, 760, 766 Gimingham, C.H., 573, 576, 582, 622, 624, 629 Giroux, J.F., 345, 359 Gischler, C.E., 663, 670 Gist, C.S., 646, 653 Gitay, H., 613, 615, 616, 625, 627, 629 Given, D., 27, 33 Given, D.R., 87, 110 Gjerstad, D.H., 618, 628 Gladden, J.B., 591, 607 Glaeser, B., 663, 670 Glaeser, J.L., 68, 95 Glass, A.D.M., 545, 567 Glass, P.O., 636, 656 Glazener, W.C., 337, 359 Gleason, H.A., 585, 595, 598, 605 Gleeson, S.K., 190, 197, 199, 201, 219, 615–617, 622, 624, 627, 629 Glenn, S.M., 154, 157 Glenn-Lewin, D.C., 187, 217, 488, 501, 585, 605, 611, 629, 717, 720 Glinski, J., 512, 516 Glitzenstein, J.S., 190, 191, 193, 195, 196, 199, 201, 203, 217, 224, 248 Gloaguen, J.-C., 23, 34, 83, 102 Gloersen, P., 66, 67, 98, 102 Glossop, B.L., 381, 384 Glover, H.G., 368, 383 Gloyne, R.W., 189, 200, 217, 227, 228, 248 Glucksman, J., 142, 157, 593, 603 Glyphis, J.P., 281, 283 Gnaspini-Netto, P., 647, 656

Gochenaur, S.E., 530, 540, 544 Goddard, J., 585, 593, 596, 605 Godfrey, C.L., 506, 519 Godfrey, P.J., 441, 447 Godron, M., 573, 582, 711, 720 Goessling, N., 415, 416, 449 Goetz, H., 294, 305 Goff, L., 87, 113 Goff, M.L., 593, 596, 605, 647, 652, 653 Goheen, D.J., 259, 267 Gokosoyr, J., 522, 543 Gold, W.G., 45, 102, 591, 605 Goldammer, J.G., 161, 184, 436, 443, 448, 456, 465 Goldberg, D.E., 310, 311, 319, 327, 611, 612, 614–616, 619, 621, 622, 624, 625, 629, 710, 720 Goldblum, D., 208, 217 Goldfarb, G., 390, 391, 395 Goldfarb, W., 697, 705 Goldstein, M., 69, 103 Goldstein, M.C., 56, 60, 65, 102 Goldstein, R.A., 255, 268 Goldthwait, R.P., 600, 607 Golet, F.C., 331, 358 Golley, F.B., 255, 267, 476, 483 Golluscio, R.A., 294, 306 Gomez, J.R., 708, 720 Gomez Diaz, A.E., 213, 216 G´omez-Pompa, A., 467, 475, 478, 483 Gomez-Sal, A., 555, 557, 568 Gonnet, J.F., 27, 35 Gonz´alez, J.A., 43, 103 Good, R.E., 190, 197, 199, 201, 219 Goodall, D.W., 271, 283, 309, 326, 753, 766 Gooden, D.T., 510, 519 Goodin, J.R., 325, 327 Goodland, R.J.A., 738, 745 Goodland, T.C.R., 223, 239, 243, 244, 247 Goodlett, J.C., 193, 209, 210, 216, 217 Goodman, D., 666, 670 Goodman, M., 699, 704 Goodman, S.W., 385, 386, 390, 395 Goodwin, I.D., 67, 111 Goran, W.D., 387–390, 395, 396 Gorbunov, A.P., 66, 103 Gorchov, D.L., 478, 483 Gordon, A.G., 162, 184 Gordon, D.M., 337, 359 Gordon, M., 762, 766 Gore, A.J.P., 331, 359 Gore, J.A., 432, 448 Gorsira, B., 378, 383 Gosk, E., 661, 671 Goss, M.J., 510, 516

AUTHOR INDEX Gosselink, J.G., 331, 332, 334, 336, 343, 353–355, 357–359, 361, 590, 591, 599, 606, 608 Gosz, J.R., 70, 73, 119, 287, 304, 311, 326, 546, 547, 553, 569, 715, 720 Gottfried, M., 88, 102 Goudie, A., 314, 327, 673, 685 Goudie, A.S., 316, 327 Gough, L., 333, 351, 357 Goulet, F., 25, 34 Goulet, F.L., 199, 219 Gourp, V., 524, 540 Govers, G., 513, 516 Gowans, K.O., 575, 583 Grabherr, G., 58, 59, 69, 88, 102 Grable, A.R., 508, 512, 516 Grace, J., 187, 198, 216, 217, 227–229, 248, 336, 339, 345, 351, 355, 356, 359 Grace, J.B., 333, 342, 344, 345, 347, 351, 353, 359, 362, 617, 619, 620, 630 Gradwohl, J., 385, 395 Graetz, R.D., 561–563, 567 Graham, D.C., 528, 540 Graham, J.H., 525, 541 Graham Smith, L., 661, 670 Graichen, M., 510, 515 Gramenopoulos, N., 323, 328 Grandjean, P., 64, 102 Grange, A.C., 579, 581 Granstr¨om, A., 166, 185 Grant, M.A., 353, 359 Grant, M.C., 52, 67, 108 Grant, W.E., 70, 110 Granzow de la Cerda, I., 236, 238, 241, 252 Gras, J., 64, 102 Gratten, J., 137, 158 Grau, J., 62, 117 Gray, A.J., 591, 605 Gray, A.N., 198, 217 Gray, D.M., 131, 133 Gray, L.G., 427, 447 Grayson, D.K., 1, 16 Grear, J.S., 240, 247 Grebenik, E., 662, 669 Grechmann, R., 639, 652 Green, A.J.A., 88, 119 Green, G., 645, 653 Green, J.O., 646, 655 Green, P., 514, 518 Green, R., 530, 543 Green, T., 84, 97 Greenberg, R., 385, 395 Greene, D.M., 41, 100 Greene, R.S.B., 681, 685 Greene, S.E., 262, 267 Greenﬁeld, L.G., 27, 33 Greenland, D., 44, 66, 67, 102, 121

791 Greenland, D.J., 512, 516, 661, 670 Greenwood, D.J., 508, 516 Greer, J., 683, 684 Greer, T., 475, 483 Gref, R., 166, 184 Gregor, D.J., 65, 96 Gregory, M.R., 58, 63, 64, 103, 114 Gregory, S.V., 173, 184, 211, 218, 460, 465, 711, 720 Greif, F., 59, 103 Greller, A.M., 69, 103 Gremmen, N.J.M., 52, 54, 88, 103, 112 Grenke, W.C., 228, 229, 234, 235, 249 Grenon, D.A., 496, 502 Grenot, C.J., 314, 327 Grenzius, R., 401, 410 Gresham, C.A., 190, 195, 199–201, 217, 454, 465 Grice, S.S., 89, 116 Grier, C.C., 70, 73, 119, 546, 547, 553, 569 Grieve, I.C., 507, 518 Griffen, B.J., 56, 103 Grifﬁn, G.F., 429, 431, 448, 675, 686 Grifﬁn, J.R., 416, 449 Grifﬁths, J.F., 227, 247 Grifﬁths, P., 42, 103 Grifﬁths, R.P., 68, 90, 103 Grigal, D.F., 615, 629 Griggs, R.F., 78, 103, 140, 144, 153, 158, 587, 591, 605 Grigulis, K., 617, 619, 620, 630 Grillas, P., 348, 357 Grime, J.P., 26, 34, 40, 69, 80, 103, 223, 248, 307, 314, 327, 351, 359, 397, 410, 421, 445, 446, 488, 489, 493, 499, 501, 585, 592, 596, 598, 605, 611, 614, 616, 618, 621, 622, 624, 628, 629 Grimm, N.B., 427, 428, 432, 441, 447, 451, 708, 715, 720, 722 Grimshaw, H.M., 53, 94 Grinnell, J., 645, 653 Grip, H., 472, 485 Grishin, S.Yu., 137, 140, 142, 146, 148, 149, 158 Griswold, F.S., 193, 209, 219 Grobe, C.W., 88, 103 Groffman, P.M., 295, 306 Grootjans, A.P., 497, 502 Gross, K.L., 615, 616, 618–620, 622, 624, 629, 710, 720 Gross, L.J., 202, 220 Grossman, Y.C., 522, 543 Grove, J.M., 18, 34 Groven, R., 172, 185 Grover, R.F., 44, 103 Groves, R.H., 273, 274, 283, 439, 448, 681, 685

Grubb, P.J., 60, 103, 228, 229, 234, 236, 240, 248, 585, 591, 592, 594, 599, 605, 615, 625, 629 Grube, A.H., 131, 135 Gruber, A., 88, 102 Grulke, N., 66, 87, 89, 111 Grulke, N.E., 79, 89, 96, 103 Gryc, G., 63, 103 Gu, L.H., 82, 121 Guariguata, M.R., 229, 232, 248, 589, 590, 605 Gubb, A.A., 317, 326, 416, 446 Gudmundsson, O., 92, 103 Guevara, S., 475, 483 Guillerm, J.L., 274, 283, 573, 582 Gulland, J., 58, 103 Gulliksen, B., 50, 105 Gullison, R.E., 229, 248 Gulman, S.L., 571, 582 Gulmon, S.L., 279, 283, 421, 422, 445, 448 Gunatilleke, C.V.S., 234, 247 Gunatilleke, I.A.U.N., 234, 247, 472, 485, 594, 606 Gunderson, L., 683, 687 Gunderson, L.H., 336, 359, 683, 685, 686 Gunn, A., 83, 104 Gunn, A.S., 675, 685 Gunn, J., 373, 383 Guntenspergen, G.R., 190, 196, 197, 205, 208, 209, 216, 228, 242, 248, 336, 339, 345, 351, 355, 356, 359, 713, 720 Guo, H., 366, 383 Guo, Q., 319, 327 Gupta, S.C., 127, 133 Gurevitch, J., 611, 617, 619, 620, 629, 630 Gurnell, A.M., 19, 34 Gurney, S.E., 337, 359 Gurtz, M.E., 427, 450 Gustafsson, L., 173, 182 Gustin, P., 68, 109 Gutsell, S.L., 164, 165, 184 Gutterman, Y., 319, 320, 326, 713, 716, 719 Guyette, R., 386–390, 396 Guzm´an-Grajales, S., 211, 222, 240–242, 248, 252 Gyawali, D., 56, 95 Gysel, L.W., 203, 208, 217 Ha, P., 432, 449 Haack, R.A., 175, 184, 253, 258, 268 Haag, R.W., 47, 52, 83, 103 Haager, J., 142, 144, 148, 159 Haagerova, R., 142, 144, 148, 159 Haas, H., 491, 501 Habecker, M.S., 209, 217 Hack, J.T., 514, 516

792 Hackney, C.T., 343, 353, 357, 359 Haddad, N.M., 692, 705 Haddock, J.D., 591, 607 Hadley, E.B., 641, 653 Hadley, K.S., 169, 171, 186, 253, 259, 260, 267, 269, 455, 466, 716, 720, 751, 767 Haeberli, W., 66, 103 Haefner, J.H., 153, 157 Haefner, J.W., 762, 766 Haertling, J.W., 65, 103 Hagan, J.M., 474, 478, 480, 483, 486 Hagarty, E.P., 388, 390, 395 Hagedorn, H., 20, 37 Hagen, A., 43, 55, 112 Hagen, J.O., 67, 103 Haggis, M.J., 318, 328 Hagin, J., 506, 516 Hagin, L.J., 513, 516 H˚agvar, S., 43, 55, 112 Haines, B.L., 594, 605 Haines, J.R., 67, 103 Hairston Sr, N.G., 612, 629, 651, 653 Hajdas, I., 66, 97 Hajek, B.F., 514, 516 Hajek, E.R., 279, 283 Hajek, M., 407, 411 Hakkeling, R., 129, 134 Hall, A.J., 492, 493, 495, 499, 501 Hall, B.D., 65, 103 Hall, C.A.S., 736, 745, 746 Hall, C.M., 70, 103 Hall, D.K., 90, 103 Hall, G.F., 661, 670 Hall, K., 47, 118 Hall, K.J., 55, 103 Hall, P., 228–230, 232, 233, 247, 248, 250, 472, 483 Hallet, D., 68, 112 Hallingb¨ack, T., 173, 182 Halloy, S., 43, 103 Halpern, C.B., 137, 142, 144, 145, 158, 592, 605, 618, 629 Halvorson, J.J., 143, 153, 158, 159 Halwagy, R., 599, 605 Hamburg, S.P., 278, 284, 554, 566 Hamer, K.C., 474, 484 Hamilton, A.C., 436, 441, 446 Hamilton, R., 343, 353, 354, 357 Hamilton, W.D., 602, 603 Hamilton III, W.J., 324, 329 Hamman, K.C.D., 432, 448 Hammond, D.S., 478, 485 Hamr, J., 69, 103 Hancock, J.F., 621, 629 Handel, S.N., 680, 686 Hanes, T., 573, 582 Hanhim¨aki, S., 171, 184

AUTHOR INDEX Hanley, K.A., 557, 567 Hanley, T.A., 557, 567 Hansell, R.I.C., 174, 185, 198, 221 Hansen, A., 666, 670 Hansen, C.T., 64, 100 Hansen, E.M., 258, 259, 267, 269 Hansen, M.H., 460, 466 Hansen, M.M., 64, 100 Hansen, P.E., 260, 268 Hanski, I., 154, 159, 585, 606, 646, 653, 718, 720 Hanson, F., 175, 182 Hanson, H.C., 84, 103 Hanson, M.B., 59, 101 Hanson, W.C., 62, 64, 103, 114 Hansson, L., 55, 104 Hara, M., 188, 195, 207, 217, 222 Harasma, M.S., 337, 361 Harbeck, A.L., 613, 616, 631 Harborne, J.B., 258, 267 Harcombe, P.A., 190, 191, 193, 195, 196, 199, 201, 203, 217, 224, 248, 419, 420, 425, 446 Harden, C., 56, 104 Hardin, G., 662, 667, 670, 740, 746 Hardin, R.T., 556, 568, 644, 654 Hardy, D., 66, 112 Hardy, E.P., 64, 104 Hare, F.K., 39, 95, 104 Hargrave, B.T., 65, 96 Hargrove, W.W., 253, 258, 260–262, 269, 764, 767 Haripersaud, P.P., 478, 485 Harker, D., 680, 685 Harker, K., 680, 685 H¨ark¨onen, T.J., 91, 96 Harley, J.L., 525, 530, 540 Harlow, R.F., 639, 654 Harman, R., 724, 746 Harmon, M.E., 142, 145, 158, 173, 184, 190, 193, 211, 218, 228, 235, 239, 241, 244, 245, 248, 252, 592, 605, 711, 716, 720 Harms, W., 457, 465 Harmsen, R., 83, 111 Harper, J.L., 349, 359, 593, 603, 633, 644, 646, 653 Harper, J.R., 44, 104 Harper, K.A., 78, 79, 83, 104 Harper, K.T., 376, 382, 562–564, 565–568 Harrington, T.C., 189, 197, 199, 202, 219, 222 Harris, A.S., 189, 218 Harris, C., 24, 34 Harris, C.M., 58, 63, 65, 69, 70, 104 Harris, E., 144, 159, 462, 465 Harris, G.L., 682, 686 Harris, J., 462, 465

Harris, J.A., 402, 410 Harris, J.M., 51, 65, 96, 105 Harris, L.D., 390, 395 Harris, R.F., 506, 516 Harris, S.A., 66, 103 Harris, S.C., 683, 685 Harrison, P., 667, 670 Harrison, S., 718, 720 Harriss, R.C., 89, 116 Hart, R.H., 556, 557, 567, 568 Hart, S.C., 294, 304, 545, 569 Harte, J., 525, 540 Hartge, K.H., 510, 511, 516, 519 Hartley, A.E., 314, 328 Hartman, J.M., 336, 346, 359 Hartnett, B.B., 622, 629 Hartnett, D.C., 579, 583, 622, 629 Hartnick-K¨ummel, C., 400, 404, 410 Hartshorn, G.S., 197, 203, 211, 219, 224–229, 231–234, 237, 241, 242, 244, 248, 249, 251, 467, 469, 471, 472, 475, 477, 478, 484, 758, 766 Hartshorn, L., 472, 484 Harvey, A.E., 211, 218, 457, 465 Harvey, B., 162, 182 Harvey, C., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Harvey, D., 663, 670 Harvey, E.M., 615, 629 Harvey, H.R., 42, 96 Harvey, P.H., 492, 501 Harvey, S.J., 415, 416, 420, 447, 488, 491, 501 Harwood, R.S., 65, 115 Hasegawa, S., 167, 186 Haselwandter, K., 67, 104, 524, 533, 541, 543 Haskell, B.D., 676, 684 Haskins, J.L., 336, 346, 362 Hass, H., 521, 543 Hastings, J.R., 318, 327 Hastings, S.J., 66, 87, 89, 111 Hatch, D.A., 419, 448 Hater, G.R., 68, 90, 101 Hathaway, R.L., 441, 452 Hatheway, W.H., 228, 229, 234, 235, 249 Hattersley-Smith, G., 77, 104 Hatton, P.J., 392, 396 Hauﬂer, J.B., 390, 395 Haukioja, E., 54, 55, 84, 104, 111, 119, 171, 184 Haukos, D.A., 348, 359 Havstad, K.M., 312, 318, 327, 328 Havstr¨om, M., 89, 104 Hawes, I., 45, 91, 104 Hawkes, I., 17, 35 Hawkins, B.A., 488, 489, 497, 502 Hawksworth, D.L., 522, 540

AUTHOR INDEX Hawley, J.G., 124, 133 Hawley, W., 472, 484 Hayden, T.J., 387, 395, 396 Hayes, D.C., 640, 641, 655 Hayes, M.H.B., 661, 670 Hayes, M.P., 432, 448 Hayley, D.W., 62, 104 Hayman, D.S., 525, 540, 542 Haynes, R.J., 560, 567 Hayward, R.J.C., 88, 104 Heady, H.F., 278, 283, 417, 419, 448, 558, 567 Heal, O.W., 593, 608 Healey, J.R., 224, 235, 238–240, 242, 244, 247, 251 Healy, W.M., 390, 395 Heaney, D., 668, 671 Heard, D.C., 69, 112 Heath, J.P., 142, 145, 159, 590, 606 Heatwole, H.H., 255, 256, 263, 264, 267, 268 Hecht, S.B., 691, 704 Hechtel, J.L., 70, 101 Hedberg, O., 56, 104 Heddle, E.M., 682, 685 Hedin, L.O., 711, 720 Heffernan, T.D., 56, 109 Hegarty, E.E., 234, 241, 248 Hegg, O., 52, 104 Heide, O.M., 89, 104 Heidt, L.E., 64, 96 Heijne, T., 530, 540 Heikoff, J.M., 701, 704 Heil, G.W., 530, 540 Heilman, P.E., 577, 582 Heindl, B., 399, 411 Heindl-Tenhunen, B., 415, 416, 444, 452 Heine, C., 53, 104 Heinemann, H.G., 127, 135 Heinrich, J., 188, 200, 218 Heinrichs, D.H., 623, 630 Heinselman, M.L., 161, 164, 166, 168, 184, 193, 194, 218, 222, 456, 465, 577, 582 Heitschmidt, R.K., 557, 567 Hejkal, J., 593, 606 Hejny, S., 331, 363, 397, 411 Helal, H.M., 523, 540 Helbling, E.W., 87, 105 Held, M.E., 189, 195, 196, 201, 204, 205, 218 Helf, K., 646, 655 Helle, T., 62, 104, 169, 185 Hellriegel, T., 400, 404, 410 Helm, D.J., 17, 26, 28, 34, 522, 538 Helmeczi, B., 130, 134 Helmers, A.T., 126, 135 Hemmings, A.D., 62, 70, 104

793 Hemmingsen, E.A., 42, 104 Hempel, G., 59, 104 Henchi, B., 555, 566 Henderson, W., 762, 766 Hendrick, J.G., 511, 516 Hendricks, E.L., 29, 36 Hendrickson, O., 545, 547, 550, 551, 554, 567 Hendrix, L.B., 142, 159 Hendrix, P.F., 538, 539 Henry, G., 43, 83, 88, 106 Henry, G.H.R., 83, 104 Henry, J.D., 189, 193, 205, 209, 210, 218 Henry, R.F., 44, 104 Hepburn, L.R., 701, 704 Hepper, C.M., 506, 517 Hepple, S., 530, 540 Hepting, G.H., 239, 251 Herbein, S.A., 78, 104 Herbel, C.H., 318, 327 Herber, B.P., 94, 104 Herbert, D.M., 387, 396 Herbohn, J.L., 472, 483 Herman, D.J., 644, 654 Herman, P.M.J., 340, 358 Hermann, S.M., 226, 232–234, 248 Hermansson, J., 173, 185 Hermes, K., 39, 104 Hermesh, R., 167, 185 Herms, D.A., 493, 501 Hernandez, H., 62, 76, 78–80, 82, 94, 104 Hern´andez, M.J., 58, 92, 113 Herndon, A., 336, 346, 360 Herrera, L., 477, 482 Herwig, R.P., 43, 117 Herwitz, S.R., 263, 267 Hesjedal, O., 43, 55, 112 Heske, E.J., 319, 326, 717, 719 Hesse, I., 352, 362 Hessl, A.E., 88, 104 Hester, A.J., 416, 424, 425, 448, 622, 624, 629 Hester, M.W., 344, 349, 359, 360 Heusser, C.J., 17, 25, 34, 589, 599, 606 Hewett, S.W., 242, 252 Hewitt, K., 47, 104 Hey, M., 47, 110 Heydemann, B., 350, 353, 356 Heyligers, P.C., 436, 441, 448 Heywood, R.B., 45, 108 Heywood, V.H., 125, 126, 128, 130, 131, 133 Hibbs, D.E., 189, 202, 203, 205, 206, 208, 218 Hickson, D.E., 423, 425, 448 Higgins, G.M., 509, 517 Higgins, S.I., 590, 606 Higgins, W.D., 264, 268

Hik, D.S., 346, 353, 359, 579, 582, 583 Hilborn, R., 740, 746 Hill, A.R., 659, 670 Hill, J.K., 474, 484 Hillel, D., 389, 395, 506, 517 Hilliard, S.B., 334, 360 Hillyard, D., 674, 685 Hils, M.H., 613, 616, 629 Hinchman, R.R., 385, 386, 395 Hinsenveld, M., 388, 390, 395 Hinz, E.A., 128, 134 Hinzman, L.D., 43, 66, 106 Hipps, L.E., 528, 530, 538 Hirano, S.S., 528, 539 Hirose, T., 143, 153, 159, 592, 604 Hjalten, J., 579, 582 Hjeljord, O., 170, 182 Hjelm, U., 89, 103 Ho, I., 524, 540 Hoare, J., 572, 581 Hoare, J.R.L., 439, 441, 448 Hobbie, J.E., 90, 106, 646, 655 Hobbie, J.H., 169, 170, 185 Hobbs III, H.H., 432, 448 Hobbs, J.A., 126, 127, 129, 135 Hobbs, N.T., 288, 294, 296, 302, 304 Hobbs, R.J., 281, 282, 283, 284, 413–416, 421–425, 444, 445, 448, 573, 576, 582, 645, 653, 673–682, 685–687, 708, 714, 720, 721 Hobbs, V.J., 421, 422, 445, 448 Hodge, H., 44, 99 Hodgson, J.G., 40, 69, 80, 103, 371, 383 Hodson, R.E., 522, 543 Høeg, H.J., 57, 110 Hoekstra, T.W., 325, 327, 708, 710, 719 Hofer, H.R., 88, 104 Hoffert, M.I., 67, 102 Hoffman, C.A., 261, 267 Hoffman, G., 510, 517 Hoffman, J.A., 337, 359 Hoffmann, A.J., 279, 280, 283 Hoffmann, C., 401, 411 Hoffmann, R.S., 54, 55, 105 Hoffpauir, C.M., 353, 359 Hofgaard, A., 161, 162, 164, 166, 169, 170, 172, 173, 183, 184, 188, 211, 218 Hofman, L., 556, 557, 566 Hofmann, A., 67, 104 Hofmann, D.J., 65, 105 Hofstede, R.G.M., 60, 105, 555, 567 Hogan, G.D., 376, 383 H¨ogberg, P., 524, 541 Hogenbirk, J.C., 335, 336, 346, 348, 359, 360 Hoham, R.W., 42, 96 Hoinkes, H., 19, 20, 34 Holbrook, N.M., 230, 248

794 Holdgate, M.W., 41, 43, 53, 64, 69, 94, 95, 105, 109 Holdridge, L.R., 56, 105, 228, 229, 234, 235, 249, 469, 484 Hole, F.D., 131, 133 Holge, W., 389, 395 Holl, K.D., 378, 383 Holla, T.A., 169, 184 Holland, E.A., 322, 325, 557, 558, 560, 567 Holland, M., 644, 647, 649, 653 Holliday, R.J., 375, 383 Holling, C.S., 40, 70, 105, 169, 171, 184, 417, 419, 452, 660, 662, 670, 683, 685–687, 693, 696, 704, 751, 766 Holloway, J.D., 474, 484 Holm-Hansen, O., 42, 87, 105 Holmes, P.M., 273, 281, 283, 285, 441, 448 Holt, B.R., 614, 616, 619, 629 Holt, J., 492, 502 Holtman, J.B., 128, 134 Holtmeier, F.K., 59, 105 Holton Jr, B., 589, 606 Holway, J.G., 47, 105 Holzhauser, H., 25, 34 Holzmann, H.P., 67, 104 Holzner, W., 493, 501 Homann, P.S., 551, 552, 554, 569 Homfray, C.K., 469, 484 Honkala, B.H., 164, 165, 182 Honma, S., 188, 195, 222 Hood, I.A., 211, 218 Hood, L., 126, 133 Hoogland, J.L., 645, 654 Hook, D.D., 512, 517 Hook, P.B., 557, 567 Hooper, D.U., 694, 705 Hope, G.S., 17, 34 Hopkins, A.J.M., 673, 675, 679, 685 Hopkins, D.M., 80, 105 Hopkinson, C.S., 353, 360 Hoppes, W.G., 642, 643, 652 Horbert, M., 399, 401, 403, 410, 411 Horn, E.M., 142, 144, 159 Horn, H.S., 611, 612, 629, 709, 720, 762, 766 Horn, J.C., 212, 218 Horn, R., 403, 411, 510, 517 Horn, S., 471, 484 Hornbach, D.J., 432, 448 Hornbeck, J.M., 547, 567 Hornbeck, J.W., 545, 547, 552, 554, 566–568 H¨ornberg, G., 166, 173, 185, 186, 352, 362 Horne, D.J., 510, 519 Horner, R., 50, 105

AUTHOR INDEX Horng, F.W., 188, 201, 216 Horning II, W.B., 87, 108 Hornung, M., 548–551, 554, 565, 569 Horowitz, A., 87, 89, 105, 526, 539 Horrowitz, M.M., 668, 670 Horstmann, K., 188, 213, 221 Horton, J.S., 428, 431, 448 Hoshiai, T., 50, 105 H¨otzl, H., 402, 410 Houle, G., 590, 606 Houng, K.H., 512, 517 House, D., 311, 326 House, D.A., 52, 121 Houser, J., 129, 134 Houston, D.R., 443, 448 Hout, P. v.d., 478, 485 Howard, G.S., 535, 541 Howard, L.F., 429, 432, 448 Howard, R.V., 58, 63, 112 Howard, W.E., 639, 654 Howard-Williams, C., 17, 35, 42, 43, 45, 104, 105, 115 Howarth, F.G., 593, 606 Howe, H.F., 213, 221, 241, 251 Howe, W.H., 429, 431, 432, 448 Howell, M.R., 681, 686 Howse, P.E., 646, 654 Hsiao, S.I.C., 67, 105 Huang, C., 513, 514, 515 Huang, X., 125, 133 Hubbell, S.P., 203, 218, 225, 229, 231–235, 242, 247–249, 252, 260, 269, 475, 483, 602, 606 Huber, A., 397, 411 Hubert, E.E., 199, 218 Hubler, G., 64, 112 Hudes, E., 131, 134 Hudson, N.W., 125, 133 Huebert, B., 64, 112 Huenneke, L.F., 197, 218, 322, 328, 413, 414, 416, 426, 443, 448, 680, 685 Huettl, R.F., 260, 267 Huggett, R.J., 662, 670 Hughen, K., 66, 112 Hughes, F., 423, 425, 439, 440, 448 Hughes, H.G., 386, 388, 389, 395 Hughes, J., 713, 719 Hughes, J.W., 213, 218 Hughes, R.G., 343, 359 Hughes, R.J., 560, 567 Hughes, S., 548, 554, 568 Hughes, T.P., 762, 767 Huisman, J., 589, 596, 599, 607 Hulbert, L.C., 288, 297, 304, 575, 582, 639–641, 652–654 Hull, A.C., 439, 451 Humphrey, R.R., 318, 324, 327 Humphreys, F.R., 264, 267

Humphries, R.N., 371, 383 Humphries, S.E., 444, 448, 681, 685 Hungerford, R., 459, 460, 465 Hunst, M.A., 509, 517 Hunt, H.W., 294, 304, 522, 524, 530, 534, 535, 539, 541, 542 Hunt, P.G., 127, 133 Hunt, R., 40, 69, 80, 103 Hunter, I., 55, 91, 99 Hunter, J.C., 189, 193, 218 Hunter, J.V., 432, 448 Hunter, M.L., 162, 174, 184 Hunter Jr, M.L., 189, 197, 198, 206, 218, 462, 465 Hunter, R., 419–423, 448 Hunter, R.B., 562, 568 Hunter, R.D., 340, 362 Hunter, S., 54, 105 Huntington, H.P., 58, 105 Huntley, B., 52, 54, 105, 288, 304 Huntley, B.J., 318, 327 Huntly, N., 254, 267, 294, 304, 580, 582, 645, 654 Huot, J., 54, 99, 110 Hupp, C.R., 427, 448 Hurni, H., 125, 133 Hurst, G.A., 641, 654 Huso, M., 613, 616, 630 Hussey, K.M., 51, 98, 114 Husted, W.M., 56, 105 Huston, M., 307, 327, 538, 541, 593, 596–599, 602, 606, 611, 621, 625, 630, 633, 654, 713, 720 Huszar, P.C., 131, 132, 133 Hutchings, H.R.W., 230, 232, 250 Hutchings, P.A., 429, 434, 450 Hutchings, R.W., 467, 475, 477, 482 Hutchinson, C.F., 324, 327 Hutchinson, S.E., 354, 358 Hutchinson, T.C., 369, 383 Hutchison-Benson, E., 64, 87, 105, 118 Hutnick, R.J., 209, 210, 218 Hutnik, R.J., 378, 384 Hutte, P., 188, 200, 218 Hwang, T.C., 62, 105 Hyt¨onen, M., 174, 184 Hytteborn, H., 172, 184, 188, 199, 204, 218, 220

Iacobelli, A., 346, 349, 351, 360 Ide, G., 510, 517 Ihlenfeldt, H.-D., 311, 329 Iida, S., 142, 145, 159, 188, 199, 208, 219 Imagawa, T., 141, 151, 159 Imbert, D., 236, 249, 476, 484 Inbar, M., 143, 159 Ingersoll, R.C., 554, 566

AUTHOR INDEX Ingham, E.R., 294, 304, 530, 534, 535, 541 Ingham, R.E., 534, 541 Ingraham, N.L., 428, 431, 446 Inoue, T., 471, 475, 484 Inouye, R., 294, 304, 580, 582, 645, 654 Insam, H., 531, 533, 541 Inselberg, A., 202, 219 International Atomic Energy Authority, 666, 670 International Erosion Control Association, 124, 133 Irmler, U., 350, 353, 356 Irwin, L.L., 321, 326 Ishikawa, S.-I., 589, 606 Ishizuka, M., 196, 221 Isichei, A.O., 576, 583 Iturriaga, R., 42, 96 Ivanovici, A.M., 39, 105 Iverson, L.R., 588, 606 Ives, J.D., 47, 69, 71, 105, 120 Ives, R.L., 58, 105 Iwaki, H., 172, 184, 198, 218 Jackowiak, B., 403, 410 Jackson, J.B.C., 611, 630 Jackson, J.R., 618, 630 Jackson, L.E., 278, 284, 293, 294, 304, 525, 541, 542 Jackson, L.L., 674, 685 Jackson, M.B., 512, 513, 516, 517 Jackson, M.T., 142, 147, 159 Jackson, P.W., 419, 420, 450 Jackson, R., 56, 60, 107 Jackson, R.B., 287, 306 Jacober, F.C., 511, 518 Jacobi, J.C., 435, 438, 443, 448 Jacobo, E., 420, 447 Jacobs, H., 701, 704 Jacobsen, T., 663, 670 Jacobson, S., 131, 134 Jacoby, G., 66, 112 J¨adicke, B., 407, 410 Jaeger, C., 89, 103 Jaffe, D.A., 64, 105 Jager, A., 510, 517 J¨ager, E., 404, 410 Jakeman, A.J., 660, 670 Jaksic, F., 279, 283 Jalas, J., 408, 410 James, H., 426, 450 James, H.F., 426, 435, 448 James, M.E., 300, 304 James, P.W., 590, 607 James, S.W., 294, 306, 436, 443, 448, 449, 640, 641, 654 Jameson, D.A., 572, 578, 582 Jane, G.T., 188, 196, 200, 208, 218

795 Janetschek, H., 26, 34 Janos, D.P., 472, 483, 484, 590, 593, 605, 606, 717, 720 Jans, L., 225–229, 231–234, 238, 242, 249, 250 Jansson, B.-O., 683, 685 Janzen, D.H., 471, 478, 484, 602, 606 Jaramillo, M., 478, 483 Jaramillo, V.J., 294, 304 Jarnot, R.F., 65, 115 Jarrell, W.M., 322, 328, 537, 543 Jasper, D.A., 381, 383, 527, 541 Jass, J.P., 432, 448 Jastrow, J.D., 535, 542 Jayanth, K.P., 442, 449 Jayasuriya, S., 695, 704 Jeakins, S.L., 645, 652 Jebb, M., 472, 484 Jefferies, R.L., 43, 53, 55, 73, 83, 88, 96, 105, 106, 108, 333, 335, 336, 345–351, 353, 357, 359, 360, 362, 579, 582, 583 Jefferson, R.G., 372, 384 Jeffrey, D.W., 52, 106 Jeffries, D.L., 562, 567 Jeggo, D.F., 635, 654 Jen´ık, J., 43, 49, 51, 106, 117 Jenkin, J.F., 52, 106 Jenkins, J., 196, 218 Jenkins, M., 66, 89, 111 Jenkins, M.B., 537, 543 Jenkins, S., 529–532, 538 Jennings, A., 66, 112 Jennings, M.R., 432, 448 Jenny, H., 17, 29, 33, 34, 52, 97, 594, 606 Jensen, A., 333, 350, 360 Jensen, K.C., 70, 110 Jeppson, L.R., 528, 541 Jepsen, M., 593, 603 Jetten, V.G., 478, 485 Jevons, F., 735, 746 Jevons, W.S., 735, 746 Jewett, K., 548, 554, 567 Jhilmit, S., 469, 483 Jobb´agy, E., 287, 295, 304–306 Jochimsen, M., 17, 18, 29, 30, 34, 75, 106 Joffre, R., 294, 305 Johannesson, K., 592, 606 Johansen, J.R., 536, 541, 543, 555, 562–564, 567, 569 Johansson, M.E., 617, 619, 620, 630 John, B.S., 18, 36 Johns, A.D., 471, 473, 484 Johns, A.G., 473, 484 Johns, H.O., 68, 95 Johns, J., 478, 486 Johns, J.S., 467, 484 Johns, R.J., 223, 229, 249 Johnson, A.F., 589, 606

Johnson, A.H., 124, 135, 243, 244, 251, 590, 592, 609, 717, 722, 752, 766 Johnson, B., 132, 134 Johnson, B.J., 65, 105 Johnson, C.W., 332, 337, 355, 360 Johnson, D.A., 79, 89, 97, 98, 442, 446 Johnson, D.J., 187, 221 Johnson, D.L., 244, 251, 665, 670 Johnson, D.W., 545–547, 552–554, 567, 568 Johnson, E.A., 75, 106, 161, 164, 165, 174, 184, 457, 465 Johnson, F.L., 377, 383, 390, 395 Johnson, H.B., 321, 329, 388, 394 Johnson, L., 90, 103, 106 Johnson, L.A., 52, 62, 90, 113, 116 Johnson, M.S., 368–371, 375, 376, 383, 384 Johnson, N., 675, 684 Johnson, N.C., 525, 541 Johnson, N.F., 645, 652 Johnson, N.M., 546, 547, 567 Johnson, P.L., 80, 106 Johnson, S.R., 340, 346, 360 Johnson, W., 508, 518 Johnson, W.B., 590, 599, 606 Johnson, W.C., 213, 218, 592, 606 Johnston, A., 555, 557, 569, 644, 645, 654 Johnston, B.R., 62, 64, 65, 106 Johnston, C., 532, 542 Johnston, C.A., 169, 170, 184, 577, 580, 582, 583, 646, 655, 681, 685, 717, 720 Johnston, D.W., 474, 483 Johnston, G.C., 69, 106 Johnston, M.E., 70, 103 Johnston, R.S., 79, 97 Joliffe, P.A., 626, 631 Jolley, P.M., 506, 517 Jolly, A., 662, 670 Jonasson, S., 89, 104 Jones, B.R., 387, 396 Jones, C.B., 662, 670 Jones, C.G., 131, 134, 580, 582, 680, 685, 717, 720, 750, 756, 762, 764, 766 Jones, C.S., 313, 328 Jones, E.W., 241, 249 Jones, J.B., 59, 106 Jones, J.T., 594, 603 Jones, M., 524, 543 Jones, M.B., 87, 106, 294, 304, 573, 582 Jones, N., 467, 479, 485 Jones, P.D., 66, 106, 113 Jones Jr, S.B., 340, 358 Jonsell, M., 173, 182 Jonsson, B.G., 173, 184, 188, 193, 194, 216, 218 Jonsson-Ninniss, S., 346, 360 Jordan, E., 17, 34

796 Jordan, M.J., 90, 106 Jordan, P.W., 311, 327 Jordan III, W.R., 760, 766 Jordan, W.R.I., 674, 686 Jorgensen, E.E., 334, 360 Jorgensen, J., 550, 569 Jorgenson, J.C., 76, 101 Joshi, M.C., 579, 582 Joshi, N.V., 472, 483 Josselyn, M.N., 337, 357 Jou, A.S.R., 680, 687 Jouventin, P., 59, 70, 106, 120 Joyal, C., 171, 182 Joyce, G.D., 455, 465 Joyce, L.A., 287, 303, 306 Julkunen-Tiitto, R., 84, 97 Juma, N., 525, 542 Jung, H.G., 55, 95, 106 Jurgensen, M.F., 211, 218, 457, 465 Jurinak, J.J., 523, 524, 539, 541 Juritz, J., 441, 448 Justice, N.L., 456, 465 Kachanoski, R.G., 513, 517 Kadlec, J.A., 342, 346, 362 Kadomura, H., 141, 151, 159 Kainm¨uller, C., 43, 106 Kainz, M., 506, 515 Kaiser, J., 89, 106 Kaldor, N., 742, 746 Kalela, O., 55, 106, 118 Kalinin, M.I., 189, 200, 218 Kalliola, R., 590, 592–594, 606 Kalluri, S.N.V., 300, 304 Kamnalrut, A., 573, 582 Kanazawa, Y., 196, 221 Kanda, H., 39, 54, 94, 106 Kane, D.L., 43, 66, 106 Kane, T.C., 646, 655 Kantrud, H.A., 332–334, 360 Kapitsa, A.P., 63, 65, 114 Kaplan, D., 276, 284 Kaplan, R.D., 739, 746 Kapos, V., 224–226, 228, 229, 231, 235, 244, 249–251, 467, 471, 475, 477, 482, 485, 713, 721 Karamysheva, Z.V., 581, 582 Kareiva, P., 259, 267 K¨arenlampi, L., 62, 106 Karentz, D., 42, 87, 92, 106, 113 Karl, D.M., 52, 68, 106 Karlen, D.L., 510, 511, 514, 515–517, 519 Karpa, D.M., 142, 154, 159 Kartawinata, K., 226, 251 Kasenene, J.M., 475, 485 Kasparov, A., 58, 109 Kassas, M., 323, 327 Kates, R.W., 673, 687, 717, 721

AUTHOR INDEX Katz, J., 337, 359 Katz, T.K., 718, 722 Kauffman, J.B., 235, 251, 456, 465, 473, 485, 486 Kaufman, D.W., 639, 640, 652, 654 Kaufman, G.A., 639, 640, 654 Kaufmann, M.R., 512, 519 Kaufmann, R., 736, 745, 746 Kauhanen, H., 188, 199, 200, 218 Kaur, J., 124, 134 Kautto, A., 62, 106 Kawasaki, K., 758, 766 Kay, F.R., 312, 329 Kazaklis, A., 276, 284 Kazempour, M.K., 417–419, 421, 450 Kazmierczak, E., 153, 159 Keage, P.L., 66, 88, 94 Keammerer, W.R., 592, 606 Keddy, P.A., 340, 342, 351, 360, 363, 598, 599, 606, 612, 617, 619–621, 623, 624, 626, 629–631 Kedziora, A., 682, 686 Keeling, C.D., 508, 517 Keen, N.A., 528, 541 Keenan, R.J., 460–462, 465 Keeney, D.N., 436, 443, 451 Keifer, H.H., 528, 541 Keighery, G.J., 282, 284 Keith, J.O., 337, 360 Keith, J.R., 90, 116 Kelley, J.J., 89, 99 Kellman, M., 225, 249, 477, 484 Kellman, M.A., 478, 485 Kellom¨aki, S., 172, 185 Kelly, G.J., 42, 110 Kelly, J., 337, 362 Kelly, J.M., 547, 567 Kelly, P.M., 66, 106 Kelly, V.R., 598, 603 Kelman, S., 702, 704 Kelty, M.J., 189, 195, 218, 460, 466, 469, 481, 485 Kemp, E.M., 317, 327 Kemp, P., 535, 539 Kemp, P.F., 353, 360 Kemp, P.R., 311, 328 Kemper, W.D., 507, 508, 515, 517, 561, 567 Kendall, H., 724, 729, 746 Kendall, M.A., 347, 360 Kendrick, B., 532, 541 Kennedy, D.N., 236, 237, 240, 242, 249 Kennedy, H.E., 511, 517 Kennedy, J.C., 57, 106 Kennedy, R.J., 635, 654 Kenney, D.S., 529, 542 Kennicutt II, M.C., 42, 96 Kenworthy, J.B., 226, 251

Keough, M.J., 333, 337, 358 Kerbes, R.H., 55, 106, 333, 346, 348, 350, 360 Kerkstra, K., 682, 687 Kerp, H., 521, 543 Kerr, R.A., 41, 66, 107 Kerry, E., 90, 107 Kershaw, G.P., 63, 78, 79, 83, 104, 107 Kershaw, K.A., 645, 654, 659, 670 Kershaw, L.J., 78, 107 Kertell, K., 78, 118 Kessler, E., 196, 218 Ketcheson, J., 127, 133 Ketcheson, J.W., 127, 133, 507, 518 Ketner, P., 324, 328 Key, C.H., 415, 416, 421, 452 Keynhans, C.J., 434, 449 Khalil, M.A.K., 64, 65, 107, 113 Khan, Z., 478, 485 Khazzoom, J.D., 735, 746 Khen, C.V., 474, 484 Khoshoo, T.N., 660, 669 Khosla, A., 660, 669 Khosla, P.K., 471, 484 Khurana, D.K., 471, 484 Kido, M.H., 432, 449 Kidron, G.J., 563, 569 Kieckhefer, B.J., 641, 653 Kieft, T.L., 325, 327 Kielland-Lund, J., 169, 184 Kiener, W., 78, 107 Kienholz, R., 212, 221 Kightley, S.P.J., 55, 107 Kiilsgaard, C.W., 262, 267 Kikkeri, S.R., 388, 390, 395 Kile, G.A., 260, 263, 268 Killham, K., 546, 554, 557, 558, 561, 567, 645, 654 Kilmer, V.J., 127, 132 Kilpel¨ainen, L., 55, 106 Kim, K.T., 64, 108 Kim, S.H., 64, 108 Kimball, A.J., 189, 197, 198, 206, 218 Kimball, B.A., 337, 360 Kimberley, M.O., 211, 218 Kimmins, J.P., 262, 267, 460–462, 465 Kimura, K., 145, 155, 159 Kimura, W., 592, 606 Kindell, C.E., 613, 630 King, B., 56, 60, 107 King, D.A., 204, 218, 228, 229, 249 King, J.C., 50, 107 King, J.M., 432, 448 Kingsbury, C.M., 43, 94 Kinler, N.W., 336, 343, 360 Kinler, Q.J., 336, 360 Kinnell, P.I.A., 681, 685 Kinter, W.B., 68, 112

AUTHOR INDEX Kinzie, R.A.I., 432, 449 Kinzig, A.P., 525, 540 Kira, T., 228, 234, 249 Kirby, R.E., 336, 360 Kirch, P., 426, 451 Kirchgeorg, A., 399, 403, 410, 411 Kirk, R.M., 64, 103 Kirk-Spriggs, A.H., 474, 484 Kirkby, A., 514, 517 Kirkby, M.J., 514, 517 Kirkman, L.K., 333, 360 Kirkpatrick, J.B., 68, 69, 115 Kitajima, K., 602, 603 Kitayama, K., 142, 159 Kitazawa, T., 130, 133 Kitazawa, Y., 130, 133 Kitchell, J.F., 524, 541 Kittel, T.G.F., 291, 292, 294, 306 Kitzberger, T., 193, 196, 208, 220, 222, 253, 260, 269 Kjelvik, S., 43, 55, 112 Kjerve, B., 195, 200, 201, 217 Klant, R., 701, 704 Klausner, S.D., 127, 133 Klausnitzer, B., 408, 410 Kl¨ay, J-R., 52, 121 Klee, G.A., 125, 133 Klein, D.R., 44, 62, 69, 84, 107, 169, 182 Klein, E., 614, 616–620, 630, 631 Klein, R.M., 190, 205, 220 Klein, W.N., 90, 116 Kleiner, E.F., 320, 327, 563, 567 Klemedtsson, L., 549, 554, 568 Klemm, G., 405, 410 Klemmedson, J.O., 423, 440, 449, 557, 565 Klemow, K.M., 372, 383 Kley, D., 64, 101 Klinger, L.F., 62, 107 Klopate, J.M., 354, 360 Klopatek, J.M., 562, 566, 567 Klopsch, M., 190, 203, 221 Klotz, S., 404, 410 Klug-P¨umpel, B., 69, 107 Knapp, A.K., 294, 297, 302, 304, 306, 340, 346, 360, 645, 654 Knapp, G., 58, 107 Knapp, G.W., 57, 94 Knapp, P.A., 317, 327 Kneeshaw, D.D., 162, 167, 183 Kniffen, F.B., 334, 360 Knight, D., 196, 217 Knight, D.H., 190, 198, 212, 220, 225, 226, 230, 252, 320, 325, 455, 465, 681, 684, 713, 716, 719, 720 Knight, W.G., 524, 541 Knopf, F.L., 390, 395, 428, 429, 431, 432, 448, 449

797 Knops, J.M.H., 416, 449 Knott, D.M., 340, 360 Knowles, O.H., 467, 479, 485 Knowles, P., 169, 184 Knutson, H., 641, 654 Kobayashi, M., 140, 142, 144, 146, 149, 160 Koch, J.M., 674, 687 Koch, M.S., 344, 355, 359, 360 Kock, K.-H., 59, 107 Kodama, H., 457, 465 Koebel, J.W.J., 682, 686 Koerner, R.M., 66, 107 Kohl, R.A., 513, 518 Kohm, K.A., 464, 465 Kohyama, T., 198, 218 Kok, K., 90, 119 Kol, E., 45, 107 Kolasa, J., 223, 250, 709–712, 721, 750, 762, 764, 766 Kolbe, D., 480, 485 Kolodny-Hirsch, D.M., 261, 269 Komarek, E.V., 2, 15 Komarek Sr, E.V., 639, 654 Kom´arkov´a, V., 39, 41, 45, 51, 70, 72–76, 80–83, 85, 88, 90, 106, 107 Komhyr, W.D., 64, 112 Koner, Ch., 493, 501 Konopka, J., 189, 218 Koonce, J.F., 524, 541 Kooyman, G.L., 68, 107 Koponen, T., 55, 106 K¨oppen, W., 39, 107 Korn, H., 710, 720 Kornas, J., 430, 431, 449 K¨orner, C., 41, 45, 52, 71, 73, 87–89, 107, 108, 113, 115 Korotkevich, Y.S., 66, 95, 113 Korstian, C.F., 212, 218 Koske, R.E., 153, 159, 528, 540 Koskela, H., 646, 653 Koster, E.A., 66, 108 Kostopoulou, S., 506, 518 Kotanen, P., 83, 108 Kotanen, P.M., 55, 106, 333, 346, 348, 350, 360, 438, 442, 443, 449 Kotarba, A., 57, 108 Koteja, J., 639, 654 Kothari, S.K., 524, 525, 541 Kotter, M.M., 529, 541 Kottke, I., 524, 541, 543 Kottmeier, S.T., 42, 117 Kouwenhoven, J.K., 510, 517 Kovrov, M., 611, 630 Kowarik, I., 397, 405, 406, 408, 409, 410, 411 Kozlowski, T.T., 511–513, 517 Kraft, J.M., 428, 431, 448

Krainer, K., 20, 34 Krajina, V., 75, 108 Krapu, G.L., 332–334, 360 Krasney, M.E., 139, 142, 145, 148, 158 Krasny, M.E., 224, 233, 249, 592, 605 Kratz, T.K., 435, 451, 711, 721 Kratzer, A., 399, 410 Krause, H.H., 548, 554, 567 Krebs, C.J., 189, 204, 219, 454, 465, 633, 654 Krestov, P., 137, 140, 148, 149, 158 Krichevsky, M., 526, 539 Krissek, L.A., 467, 483 Kritzinger, J.J., 379, 380, 384 Krivolutskii, D.A., 532, 541 Kriwoken, L.K., 59, 108 Krog, H., 90, 116 Krogh, A., 508, 517 Kronzucker, H.J., 545, 567 Kropp, B.R., 531, 544 Kruckeberg, A.R., 143, 149, 159 Kruess, A., 259, 267 Kruger, F.J., 272, 285 Kruse, A., 639, 654 Krzysik, A.J., 387–390, 395 Ku, M.S.B., 144, 159 Kub´ıkov´a, J., 43, 117 Kubowitz, F., 508, 519 Kuc, M., 28, 30, 34 Kucera, C.L., 289, 291, 297, 298, 305, 574, 582, 641, 654 Kuch, P.J., 461, 466 Kudryashov, B.B., 27, 32 Kuepper, F., 433, 449 Kuepper, H., 433, 449 Kullman, L., 44, 66, 88, 108, 161, 164, 166, 169, 172, 174, 181, 182, 183, 184 Kulman, H.M., 579, 580, 582 Kumar, A., 579, 582 Kummerow, J., 523, 540 Kundell, J.E., 701, 704 Kunick, W., 401, 402, 404, 405, 407, 410, 411 Kunz, T.H., 646, 647, 652 Kupfer, J.A., 190, 203, 218 Kupferberg, S., 429, 432, 449 Kurz, D., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Kushima, H., 196, 221 Kushlan, J.A., 336, 360 Kutiel, P., 589, 606 Kuttler, W., 399, 410 Kuuluvainen, T., 172, 184 Kuzucuoglu, C., 277, 284 Kvˆet, J., 403, 410 Labbe, P., 236, 249, 476, 484 Lace, L.A., 474, 484

798 Lacey, J.R., 578, 582 Lacey, R.M., 389, 395 Lachenbruch, A.H., 47, 66, 108 Lacher Jr, T.E., 5, 16 Ladd, P.G., 681, 686 Ladha, J.K., 527, 540 Ladiges, P.Y., 433, 434, 449 Ladin, M.C., 386, 388, 389, 395 Laﬁte, P.W., 70, 99 Laﬂen, J.M., 513, 514, 517, 518 Lago, P.K., 585, 593, 596, 605 Laidly, P.R., 164, 185 Laitenen, T., 188, 200, 219 Lake, S., 426, 449 Lakhani, K.H., 130, 135, 371, 383, 599, 604 Lal, R., 124–126, 129, 131, 132, 132, 133, 661, 670 LaManna, J., 151, 160 Lamb, F.B., 472, 484 Lambers, H., 571, 582 Lambert, D.H., 524, 541 Lambert, F.R., 473, 484, 758, 766 Lambert, L., 211, 218 Lamont, B.B., 271–273, 283, 439, 450 Lamoureux, S., 66, 112 Lancaster, N., 589, 606 Land, M.C., 42, 87, 106 Land Use Consultants, 372, 383 Landa, K., 614, 616, 622, 624, 625, 629 Landh¨ausser, S.M., 174, 184 Landin, M.C., 386, 388, 389, 395 Landsberg, E., 399, 411 Landsberg, J., 264, 267 Lane, A.M., 415, 416, 449 Lane, J., 513, 517 Lane, L.J., 514, 518 Lang, G.E., 51, 114, 211, 218, 225, 226, 230, 252 Langdale, G.W., 128, 133 Langer, A., 405, 411 Langlands, J.P., 644, 645, 654 Lanier-Graham, S.D., 385, 395 Lanly, J.P., 470, 484 LaPlante, D., 345, 358 Laprise, D., 169, 170, 185 Lara, R.J., 65, 108 Larcher, W., 43, 44, 52, 71, 87, 107, 108, 113, 115 Lari, L., 85, 101 Larigauderie, A., 89, 108 LaRocque, G., 618, 628 LaRoe, E.T., 331, 358 Larsen, C.P.S., 174, 184, 457, 465 Larsen, J.A., 331–334, 360 Larsen, M.C., 589, 590, 603, 605, 606 Larsen, M.J., 211, 218, 457, 465 Larson, B.A., 702, 704

AUTHOR INDEX Larson, B.C., 161, 164, 172, 185, 187, 219, 223, 240, 250, 454, 456, 460, 465, 466, 469, 481, 485 Larson, M.M., 592, 600, 606 Larson, R.J., 386, 388, 389, 395 LaSalle, M.W., 343, 360 Lasca, A., 66, 112 Lashko, T.N., 532, 544 Latham, R.E., 680, 686, 710, 721 Lathrop, E.W., 388–390, 395 Latkin, J.D., 211, 218 Lattin, J.D., 173, 184, 711, 720 Latz, P.K., 439, 449 Lauenroth, W.K., 287, 288, 294–296, 298–303, 304–306, 417–419, 421, 450, 555–557, 561, 567, 568, 620, 621, 630, 645, 652, 679, 686, 710, 719 Lauer, W., 224, 249 Laundre, J.W., 318, 327 Laurance, W.F., 475, 484 Lavado, R.S., 294, 304 Lavelle, P., 131, 133 Laven, R.D., 189, 204, 215, 428, 451, 456, 465 Lavin, A., 280, 282 Lavoie, K.H., 646, 655 Lavorel, S., 154, 159, 344, 361, 418, 449 Lavrenko, E.M., 581, 582 Lawler, D.M., 45, 108 Lawrence, B., 87, 89, 111 Lawrence, D.B., 17, 22, 25, 30, 31, 34–36, 599–601, 606–608 Lawrence, E.G., 17, 34 Lawrence, M.B., 195, 221 Lawrence, W.T., 594, 605 Laws, R.M., 91, 92, 108 Lawson, D.E., 20, 35 Lawton, J.H., 474, 483, 580, 582, 680, 685, 717, 720, 756, 762, 766 Lawton, R.O., 225–229, 231–234, 242, 249 Lay, D.W., 336, 337, 343, 360 Laycock, W.A., 310, 327, 417, 419, 452, 546, 567 Le, Y.Z., 60, 122 Le Bissonais, Y., 507, 513, 517 Le Floc’h, E., 274, 276, 283, 325, 325, 436, 446, 674, 676, 679, 682, 684, 760, 765 Le H´enaff, D., 54, 110 Le Hou´erou, H.N., 273, 274, 284, 288, 302, 305, 308, 311, 323–325, 327, 573, 576, 582 Le Maitre, D.C., 274, 281, 284, 285, 424, 435, 439, 442, 449, 451 Leach, J.H., 432, 449 Leader-Williams, N., 54, 69, 108 Leaf, A.L., 549, 567

Leake, J.R., 524, 543 Learner, M.A., 432, 451 Leatherwood, S., 59, 114 Lebouvier, M., 69, 118 Lebreton, J.D., 154, 159 Leck, M.A., 212, 219, 350, 360 Leduc, A., 164, 168, 171, 182 Lee, C.A., 555, 561, 567 Lee, E., 130, 133 Lee, J.A., 89, 121 Lee, K.E., 131, 133, 509, 517, 646, 654 Lee, L.W., 602, 603 Lee, S.H., 64, 108 Lee, S.Y., 341, 356 Lee, T.D., 189, 197, 198, 202, 206, 219 Lee, W.G., 679, 681, 684 Leege, T.A., 644, 654 Leemans, R., 188, 204, 219 Lefkovitch, L.P., 434, 446 Legendre, L., 50, 105 Legg, C.J., 333, 336, 348, 360, 361 Lehre, A.K., 141, 157 Lehrsch, G.A., 506, 517 Leicach, S., 495, 501 Leicht, P.N., 593, 603 Leigh, E.G., 255, 267 Leigh Jr, E.G., 475, 484 Leighton, M., 224, 226, 229, 234, 249, 473, 484, 717, 720 Leiser, A.T., 131, 133 Leisman, G., 592, 606 Leitner, L.A., 398, 411 Leiva, J.C., 25, 37 Lemcoff, J.H., 294, 304 Lemme, G.D., 513, 517 Lenihan, J.M., 163, 184 Lennartsson, T., 167, 174, 185 Lenschow, D.H., 64, 101 Lenski, R.E., 522, 543 Lentz, R.D., 507, 513, 514, 517 Lenzano, L., 25, 37 Le´on, R.J.C., 287, 288, 295, 304–306, 418, 420, 447, 451, 491–494, 497, 498, 501, 502 Leonard, H.J., 662, 670 Leonard, J., 472, 484 Leonzio, C., 85, 101 Leopold, A., 723, 746 Leopold, D.J., 211, 221, 378, 383, 588, 591, 606, 680, 686 Lepage, S., 62, 110 Lepart, J., 348, 357 Lepri, L., 52, 100 Lepˇs, J., 611, 612, 625, 630 Leresche, B., 62, 117 LeResche, R.E., 55, 108 Lertzman, K.P., 189, 202, 204, 219, 454, 465

AUTHOR INDEX Lesel, R., 69, 108 Lesica, P., 694, 705 Lessmann, J.M., 333, 336, 341, 345, 348, 349, 355, 357, 360 Letey, J., 508, 517 Lettau, H.H., 53, 108, 196, 200, 217 Letzsch, W.S., 336, 360 Leuchs, W., 403, 411 Leung, Y., 617, 619, 620, 630 Levanon, D., 507, 518 Levenson, J.B., 715, 721 Levey, D.J., 213, 221, 240, 241, 249, 251 Levin, S.A., 333, 361, 386, 395, 707, 720, 722, 764, 766 Levings, S.C., 337, 359 Levy II, H., 64, 112 Lewis, B.G., 530, 543 Lewis, D.H., 523, 526, 541 Lewis, K., 530, 543 Lewis, L.A., 665, 670 Lewis, M.C., 83, 110 Lewis, P.A., 87, 108 Lewis, S.J., 336, 360 Lewis Jr, W.M., 52, 67, 108, 590, 599, 605 Lewis Smith, R.I., 591, 592, 606 Lewkowicz, A.G., 47, 108 Ley, E., 226, 234, 252 Leyton, L., 126, 133 Lezberg, A., 189, 193, 217 Lægaard, S., 60, 108 Li, D.W., 532, 541 Li, H., 716, 720 Li, H.W., 427, 450 Li, J.H., 56, 109 Li, M., 234, 249 Li, T.-X., 164, 182 Li, X.L., 524, 541 Liang, J.R., 56, 109 Liang, T., 228, 229, 234, 235, 249 Licht, L.A., 369, 384 Lichter, J., 589, 606 Lieberman, A.S., 276, 277, 284, 595, 607 Lieberman, D., 190, 197, 203, 211, 219, 224, 225, 228–230, 233, 234, 237, 242, 249, 251, 642, 654 Lieberman, M., 190, 197, 203, 211, 219, 224, 228–230, 233, 237, 242, 249, 642, 654 Liegel, L.H., 479, 484 Lien, L., 43, 55, 112 Lienkaemper, G.W., 173, 184, 211, 218, 711, 720 Light, J.J., 45, 108 Light, S.S., 683, 685, 686 Lightfoot, D.C., 318, 327 Likens, G.E., 193, 215, 231, 247, 546, 547, 552, 554, 566, 567, 711, 718, 719–721, 764, 765

799 Liljeroth, E., 525, 541 Lima, A.P., 233, 249 Lin, Q., 341, 346, 353, 360 Lin, Z.Y., 59, 108 Lindeboom, H.J., 54, 97, 108 Linden, O., 666, 670 Linderman, R.G., 472, 483 Linderman, S.A., 55, 108 Lindley, D., 68, 95 Lindner, H., 510, 517 Lindo, L.S., 240–242, 249 Lindqvist, R., 528, 541 Lindr¨oth, C.H., 20, 35 Lindsay, D.C., 69, 108 Lindstrom, M.J., 510, 513, 517, 518 Linhart, Y.B., 474, 483, 642, 643, 653 Link, S.O., 557, 560, 562, 563, 566, 567 Linkins, A.E., 68, 79, 90, 109, 116 Linscombe, R.G., 336, 337, 343, 360, 363 Linsdale, J.M., 645, 653, 654 Linthurst, R.A., 337, 346, 362 Lippitt, L., 537, 541 Lipps, J.H., 58, 109 Lipscomb, D.J., 190, 195, 199–201, 217, 454, 465 Liro, A., 529, 543 Liston, G.E., 43, 106 Litaor, M.I., 51, 109 Litchﬁeld, J.H., 666, 670 Little, P., 668, 670 Little, S., 457, 465 Littman, S., 131, 134 Liu, B., 514, 517 Liu, F., 67, 121 Liu, J.K., 56, 109 Liu, K.-B., 475, 483 Liu, Q.-H., 188, 199, 218 Liu, S.C., 64, 101, 112 Livingston, S.J., 128, 134 Lizarralde, N.S., 432, 449 Lizotte, M.P., 42, 117 Ljungqvist, H., 170, 183 Llano, G.A., 45, 64, 109 Lo, A., 124, 133 Lobb, D.A., 513, 517 Lochmiller, R.L., 387, 394 Lockaby, B.G., 131, 134, 467, 472, 485 Lockwood, J.L., 530, 541 Lodge, D.J., 2, 16, 211, 222, 224, 235, 236, 238, 241, 243–245, 249–252, 435, 449, 594, 605, 635, 636, 651, 656, 657, 713, 722, 752, 753, 763, 766, 767 Lodge, D.M., 336, 360 Lodge, T.E., 337, 340, 360 Loehle, C., 615, 631 Logan, I.B., 696, 704 Lohmann, L., 665, 670 Lohse, K.A., 140, 149, 159

Lojan Idrobo, L., 479, 480, 484 Londo, G., 403, 411 Long, J., 462, 465 Long, J.N., 462, 466 Long, S.P., 573, 582 Longhurst, R., 666, 669 Longrigg, C., 56, 109 Longton, R.E., 43, 45, 68, 88, 95, 109 L¨onnberg, E., 70, 109 Lonsdale, W.M., 415, 416, 429, 431, 439, 441, 446, 449 Looman, P.E., 623, 630 Loope, L., 195, 220, 354, 361 Loope, L.L., 324, 327, 443, 451 Loope, L.M., 336, 346, 360 Loope, W.L., 324, 327, 563, 567 L´opez Ornat, A., 235, 249 Lopoukhine, N., 674, 685 Lorence, D., 426, 449 Lorenz, D.C., 17, 27, 36 Loreti, J., 287, 306 Lorimer, C.G., 189, 190, 193, 194, 197, 199, 200, 205, 216, 217, 219, 229, 230, 235, 240, 248, 249 Lorius, C., 64, 66, 95, 97, 113 Losleben, M., 67, 121 Lotan, J.E., 165, 184 Lotti, R., 66, 97 Loucks, O.L., 190, 193, 196, 197, 215, 707, 709, 710, 713, 721, 722 Lovari, S., 69, 98 ´ 42, 109 L¨ove, A., L¨ove, D., 39, 42, 51, 109 Love, F.G., 41, 45, 112 Lovegrove, B., 309, 316, 327 Lovejoy, T.E., 230, 232, 250, 259, 267, 467, 471, 475, 477, 482, 484 Lovelock, C.E., 472, 484 Lovett, G.M., 261, 262, 267 Lovig, D., 602, 610 Lowe, C.H., 310, 317, 319, 329, 599, 607 Lowe, E.F., 351, 360 Lowery, B., 514, 516, 561, 568 Lowman, M.D., 253–257, 260, 261, 263, 264, 266, 267, 268, 694, 704 Lu, K., 209, 220, 229, 231, 232, 234, 238, 239, 250 Lubchenco, J., 333, 360 Lubin, D., 87, 93, 101, 105 Lubke, R.A., 379, 383 Lucas, H.L., 646, 655 Lucas, R.E., 128, 134 Luck, R.F., 260, 268 Luckman, B.H., 51, 109 L¨udi, W., 17, 35 Ludwig, D., 417, 419, 452, 740, 746 Ludwig, J.A., 376, 382, 680, 686, 687 Lugg, D.J., 66, 111, 121

800 Lugg, D.L., 66, 114 Lugo, A.E., 70, 109, 223–226, 228, 230–234, 236, 238–240, 242–245, 247–251, 352, 360, 435, 451, 471, 476, 479, 482, 482, 484, 485, 590, 608, 633, 656, 678, 679, 684, 716, 721, 755, 766 L¨uhrte, A. v., 403, 411 Luken, J.O., 1, 15, 87, 89, 96, 378, 383, 415, 416, 449, 585, 590, 594, 599, 606, 679, 686, 717, 721, 758, 766 Lukkari, A., 55, 112 Lukschanderl, L., 57, 109 Lund, J.A., 546, 568 Lund, L.J., 314, 329 Lund, L.L., 591, 599, 609 Lundberg, P., 169, 183 Lundgren, L., 590, 606 Lundmark-Thelin, A., 550, 554, 568 Lundquist, L., 200, 222 Lunn, N.J., 65, 70, 109 Lunt, I.D., 424, 449 Lussenhop, J., 532, 541 Lutjeharms, J.R.E., 39, 109 Luttich, S., 54, 110 Lutz, H.J., 25, 35, 193, 209, 219 Luxmoore, R.J., 322, 328 Lwanga, J.S., 475, 485 Lyck, L., 58, 109 Lyford, W.H., 193, 195, 209, 219 Lyles, L., 504, 517 Lynch, J.F., 235, 245, 246, 252, 636, 654 Lynch, J.J., 336, 337, 343, 353, 360 Lynch, J.M., 488, 502, 525, 541, 562, 568 Lyon, L.J., 639, 654 Lyons, W.B., 17, 35 Mabin, M.C.G., 64, 103 MacAller, R., 537, 539 MacCleery, D.W., 456, 460, 465 MacCracken, M.C., 67, 102 MacDiarmid, B.N., 646, 654 MacDonald, G., 66, 112 MacDonald, I.A.W., 154, 159, 281, 285, 415, 416, 436, 440, 441, 443, 448, 449 Mace, G.M., 681, 686 Macfayden, A., 599, 603 Machin, J.L., 390, 395 MacInnes, J.R., 87, 109 MacIsaac, D.A., 17, 29, 33, 141, 157, 591, 599, 600, 603 Mack, R.N., 144, 159, 190, 211, 216, 500, 501, 522, 543, 546, 556, 568, 611, 630 Mackay, D., 63, 109 Mackay, J.R., 45, 52, 90, 109 MacKay, W.P., 645, 657 MacKenzie, M.D., 190, 193, 205, 222 Mackintosh, N.A., 91, 109 Macko, S.A., 42, 96

AUTHOR INDEX MacLean, D., 161–164, 186 MacLean, D.A., 168, 183, 577, 582, 583 MacLean, D.W., 193, 195, 209, 219 MacLean Jr, S.F., 84, 95 MacMahon, J.A., 31, 35, 79, 94, 98, 140, 142, 151–153, 157, 158, 309, 311, 314, 318–320, 325, 326–328, 526–531, 537, 538, 543, 593, 600, 603, 607, 713, 717, 721 Macmillan, B.H., 30, 33, 591, 593, 603 Maddock, L., 556, 568 Maggi, O., 472, 484 Maggio, R., 288, 303 Magnuson, C.E., 258, 267 Magnuson, J.J., 2, 16, 764, 766 Magnusson, B., 142, 143, 147, 148, 154, 158 Magnusson, W.E., 233, 249 Mahall, B.E., 425, 447 Mahaney, W.C., 17, 35, 56, 66, 83, 109 Mahar, D.J., 691, 705 Maheswaran, J., 594, 606 M¨ahr, E., 88, 102 Maillet, J., 274, 283 Mainguet, M., 323, 327, 665, 670 Maizels, J., 19, 35 Maizels, J.K., 19, 21, 35 Majer, J.D., 325, 327, 381, 383, 593, 606, 674, 678, 680, 686, 752, 766 Major, J., 17, 33, 596, 600, 601, 604 Major, J.A., 67, 122 Makeyev, V., 58, 109 Makurat, A., 402, 410 Malaisse, F., 372, 374, 382 Malajczuk, N., 282, 283 Malanson, G.P., 164, 184 Mallik, A.U., 346, 351, 360 Malloch, D.W., 521, 541 Mallona, M.A., 238–242, 247, 252 Mallory, F.F., 56, 109 Malmer, A., 472, 475, 485 Malo, A.R., 312, 328 Malo, D.D., 515, 517, 518 Maloney, K.A., 198, 219 Maltby, E., 333, 334, 336, 337, 348, 360, 361 Malthus, T., 723, 746 Mamolos, A.P., 294, 305 Manabe, S., 66, 67, 109, 110 Manders, P.T., 274, 284 Mani, M.S., 43, 110 Manion, P.D., 437, 449 Manley, J.T., 556, 568 Mann, D.H., 31, 36, 590, 600, 606, 608 Mann, L.K., 545, 547, 552, 554, 568 Manney, G.L., 65, 115 Manning, A.E., 58, 101 Manning, M.M., 555, 568

Manning, S.J., 429, 449 Mannion, A.M., 666, 670 Manseau, M., 54, 99 Mansikkaniemi, H., 188, 200, 219 Manske, L., 294, 303 Manzanares, A.R., 311, 329 Maquirino, P., 225, 227, 228, 231, 240, 242, 244, 251 Marathe, K.V., 535, 541 Marble, J.R., 562–564, 567, 568 Marchand, P.J., 199, 219 Marchant, H.J., 42, 88, 89, 100, 110 Marcus, L., 69, 103 Mares, M.A., 5, 16 Margalef, R., 595, 607 Margaris, N., 276, 283 Margules, C.R., 714, 721 Marion, G.M., 594, 609 Mark, A.F., 17, 28, 36, 78, 90, 110, 112, 114, 589, 590, 607, 608 Markgren, G., 170, 183 Markham, K.R., 87, 110 Marks, B., 464, 466 Marks, P.L., 187, 205, 215, 219, 223, 247, 615, 616, 618, 619, 627, 629, 630 Marlette, G.M., 419, 420, 445 Marneweck, G.C., 590, 606 Maron, J.L., 419, 421, 449 Marquette, W.M., 58, 110 Marquez, L., 67, 116 Marquis, R.J., 261, 268 Marr, J.W., 51, 59, 69, 79, 110, 120, 121 Marrs, R.H., 372, 382, 585, 588, 599, 600, 607 Marschner, B., 401, 411 Marschner, H., 523–525, 540–543 Marsden, I.D., 349, 361 Marsh, C., 473, 482 Marsh, J.G., 67, 122 Marsh, P., 47, 110 Marsh, P.C., 432, 449 Marshall, B.V., 66, 108 Marshall, S.A., 646, 652 Marshall, W.H., 640, 656 Marsili, L., 85, 101 Marston, R.A., 388–390, 395 Marston, R.B., 546, 566 Martens, H.E., 79, 121 Marti, E., 708, 720 Martin, B., 646, 654 Martin, C.W., 545, 547, 552, 554, 566–568 Martin, G.L., 189, 203, 217 Martin, J.P., 506, 517 Martin, M.H., 337, 361 Martin, R., 457, 465 Martin, R.E., 456, 465 Martin, T.H., 683, 685

AUTHOR INDEX Martin, W.C., 378, 384 Martin, W.H., 198, 220 Martin, W.P., 535, 540, 562, 566 Martin-Ruiz, P., 664, 670 Martinez-Ghersa, M.A., 492, 501 Martinez-Meza, E., 314, 326 Mart´ınez-Ramos, M., 224, 226, 228, 242, 249, 250 Martinovic, L., 510, 517 Martiocorena, E., 44, 112 Martore, R.M., 340, 360 Maruta, E., 143, 159 Marx, D.H., 529, 542 Mary, F., 691, 705 Masaki, T., 188, 199, 208, 215, 219 Maser, C., 210, 211, 219, 675, 686 Mason, C.R., 681, 684 Mason, R.R., 260, 268 Massman, W., 187, 222, 228, 252 Masters, K.A., 429, 431, 448 Masuzawa, T., 142, 143, 147, 159 Matelson, T.J., 227, 228, 231, 232, 234, 238, 249 Mathews, J.T., 673, 687 Mathews, W.H., 45, 109 Matlack, G.R., 190, 197, 199, 201, 219 Matson, P.A., 1, 16, 153, 154, 160, 527, 533, 542, 543, 549, 554, 569, 599, 609, 673, 687, 723, 746 Matsumae, A., 65, 118 Matsuoka, N., 51, 110 Mattheck, C., 210, 219 Mattheis, P.J., 83, 110 Matthews, J.A., 17, 18, 20, 21, 24, 27–30, 33–35, 83, 110, 588, 589, 591–594, 596–599, 602, 607, 752, 766 Matthews, J.R., 257, 268 Matthews, J.T., 717, 721 Matthews, R.W., 257, 268 Mattson, L.E., 20, 35 Mattson, W.J., 171, 175, 184, 253, 258, 268 Mattsson, J.O., 314, 328 Matveyeva, N.V., 68, 110 Matzdorf, K.D., 515, 517, 518 Maunder, M., 681, 686 Mauret, M., 522, 542 Maurette, M.T., 522, 542 Mausbach, M.J., 514, 518 Maxson, S.J., 70, 112 Maxwell, A., 341, 361 May, J., 456, 465 May, R.M., 492, 501, 522, 542 May, S.A., 417, 449 May, S.W., 292, 298, 306 Mayer, R.J., 510, 516 Mayer, W., 673, 687 Mayes, P.R., 66, 113

801 Mayumi, K., 729, 746 Mazia, C.N., 498, 502 Mazurack, A.P., 535, 539 Mazzarino, M.J., 679, 685 Mazzoleni, S., 140, 142, 159 McAleer, M.J., 660, 670 McAuliffe, J.R., 309, 311, 319–321, 328, 599, 607 McBride, N.J., 56, 113 McCabe, T.R., 343, 346, 361 McCaffrey, R.J., 51, 64, 113 McCalla, T.M., 130, 135 McCann, S.B., 44, 110, 336, 361 McCann, T.S., 59, 64, 89, 91, 97, 99, 110 McCarroll, D., 20, 35 McCarthy, B.C., 196, 209, 210, 220 McCartney, A.P., 59, 110 McClanahan, T.R., 592, 607, 680, 686 McClaran, M.P., 418, 419, 423, 445, 446, 449 McClellan, M.H., 26, 31, 33 McClure, J.W., 189, 197, 198, 202, 206, 219 McClure, M.S., 260, 268 McCollom, J.M., 190, 195, 199, 200, 210, 216, 335, 336, 358 McComb, A., 343, 349, 363 McComb, A.J., 682, 685 McConnaughay, K.D.M., 211, 219, 422, 449, 614, 617, 630 McConnell, J., 599, 603 McCool, D.K., 513, 518 McCormick, J.F., 590, 608 McCown, B.H., 74, 120, 390, 394 McCune, B., 149, 159 McDonald, K.W., 388, 390, 395 McDonald, N.H.E., 438, 447 McDonald, R.B., 223, 224, 239, 249 McDonnell, M.J., 150, 159, 213, 219, 224, 236, 250, 397, 401, 411, 585, 598, 599, 607, 614, 616, 617, 628, 630, 717, 721 McDowell, W.H., 243–245, 249, 435, 449, 471, 486, 752, 766 McElroy, R.O., 64, 97 McElroy, W.K., 594, 603 McEuen, F.S., 42, 87, 106 McEvoy, P.B., 421, 422, 449, 613, 616, 630 McEwen, L.J., 21, 35 McFarland, M., 64, 101 McFarlane, D.J., 282, 284, 682, 685 McFayden, A., 639, 654 McGee, P.A., 529, 542 McGhee, R., 56, 110 McGinley, M.A., 645, 647–649, 653, 654 McGinnes, W.J., 535, 542 McGonigle, T.P., 527, 542 McGraw, J.B., 79, 110, 617, 619, 620, 630

McGregor, J., 667, 670 McGregor, M.A., 126, 134 McInnes, S.J., 55, 91, 99 McIntyre, S., 344, 361, 418, 433, 434, 449, 452 McIntyre, T., 437, 449 McIsaac, G.F., 127, 134 McKay, C.P., 42, 99 McKee, K.L., 332, 341, 342, 344–346, 348, 349, 357, 359–361 McKell, C.H., 536, 542 McKelly, D.H., 281, 284, 435, 439, 442, 449 McKelvey, K., 718, 721 McKelvey, K.S., 457, 459, 465 McKendrick, J.D., 52–55, 73, 74, 83, 84, 89, 90, 98, 107, 110, 297, 305, 591, 607 McKenna Neuman, C., 44, 102 McKenzie, G.D., 600, 607 McLachlan, A., 589, 607 McLachlan, J.S., 214, 217 McLaren, I.A., 43, 110 McLaughlin, S.P., 317, 328 McLauglin, L., 124, 134 McLean, I.G., 681, 684 McLendon, T., 319, 320, 328 McMahon, E.A., 752, 766 McMahon, R.F., 349, 361 McManus, M.L., 437, 449 McMillan, C., 336, 362 McMullen, K.G., 385, 386, 395 McMurtry, G., 64, 116 McNabb, K.L., 467, 472, 485 McNair, D., 123, 125, 126, 128, 129, 131, 132, 134 McNair, M., 729, 746 McNamara, T.M., 68, 90, 103 McNaughton, S.J., 287–291, 294–298, 301, 303, 304–306, 488, 494, 502, 555–558, 560, 568, 578, 582 McNeilly, T., 374, 376, 383, 384 McPherson, C., 87, 98 McRae, D.J., 164, 183 McSweeney, K., 209, 217 McSwiney, C.P., 752, 766 McTaggart, A.R., 64, 102 Meadows, D.H., 664, 670, 723, 746 Meadows, D.L., 664, 670, 723, 746 Meaney, J.J., 125, 130, 134, 500, 502, 729, 746 Means, J.E., 137, 142, 144, 158 Mech, S.J., 513, 518 Medina, E., 295, 305, 439, 449 Meeder, J.F., 335, 336, 358 Meeder, L.C., 335, 336, 358 Meentemeyer, V., 292, 305 Mees, C.A., 461, 464

802 Meeus, J., 277, 284 Meffe, G.K., 690, 696, 699, 704, 705 Mehrhoff, L.A., 625, 631 Mehta, A.J., 337, 357 Meier, M.F., 66, 100 Meigs, P., 307, 328 Melack, J.M., 67, 121 Melchers, G., 42, 110 Melick, D.R., 65, 114, 429, 431, 450 Melillo, J.A., 435, 451 Melillo, J.M., 70, 73, 89, 110, 119, 169, 170, 185, 198, 217, 244, 245, 251, 546, 547, 549, 550, 553, 555, 556, 565, 566, 569, 646, 655 Mello, M.J., 343, 362 Mellor, A., 23, 35 Melnikov, I.A., 50, 105 Mendel, A.C., 124, 135 Mendelssohn, I.A., 341, 342, 344–346, 348, 349, 353, 357, 359–361 Mendonca, B.G., 51, 65, 96 Mendoza, J.A., 343, 361 Meng, W.W., 475, 483 Menge, J.A., 530, 542 Menges, E.S., 596, 599, 607, 714, 721 Menke, J.W., 439, 450, 681, 686 Mentis, M.T., 379, 383 Menzie, C.M., 65, 116 Mercer, D.C., 337, 361 Merrell Jr, T.R., 64, 110 Merrens, E.J., 189, 194, 195, 219, 239, 250 Merrill, E.D., 587, 603 Merritt, P.G., 189, 203, 204, 215 Mertens, S.K., 415, 416, 444, 450 Mesch, M.R., 319, 328 Mesnick, L., 129, 134 Messer, A.C., 23, 35, 83, 110, 594, 607 Messerli, B., 57, 97 Messerli, P., 57, 97 Messier, D., 62, 110 Messier, F., 54, 110 Metcalfe, W.S., 351, 361 Metz, L.J., 639, 654 Metzger, L., 507, 518 Meurk, C.D., 90, 121 Mexal, J.G., 529, 542 Meyer, M., 55, 114 Meyer, M.W., 64, 104 Meyer, W.B., 435, 450, 717, 721 Meyer-Arendt, K.J., 334, 362 Meyers, N.L., 209, 217 Miall, A.D., 21, 35 Mian, A.A., 388, 395 Michener, W.K., 195, 200, 201, 217 Michon, G., 691, 705 Middleton, J., 346, 360 Middleton, N.J., 665, 671

AUTHOR INDEX Midgley, J.J., 233, 250 Migenis, L.E., 469, 485 Mika, J.S., 346, 361 Mikkelsen, E., 57, 110 Mikkola, K., 169, 170, 186 Mikola, P., 169, 185 Milberg, P., 439, 450 Milchunas, D.G., 294, 296, 298, 299, 301, 305, 306, 417–419, 421, 450, 556, 557, 567, 568, 679, 686 Miles, J., 370, 383, 585, 595, 599, 607, 611, 622, 624, 629, 630, 713, 717, 721 Miller, D.E., 506, 518 Miller, D.L., 343, 346, 353, 361 Miller, D.S., 68, 112 Miller, F.P., 661, 670 Miller, G., 388, 390, 395 Miller, G.C., 429, 431, 451 Miller, G.R., 573, 582 Miller, I.L., 439, 441, 449 Miller, J.C., 260, 268 Miller, J.M., 51, 65, 96 Miller, K.K., 257, 268 Miller, M.C., 65, 68, 90, 101, 110, 117, 467, 475, 483 Miller, M.H., 513, 517, 527, 542 Miller, M.S., 467, 472, 485 Miller, M.W., 70, 110 Miller, N.G., 51, 110 Miller, P.C., 55, 74, 89, 98, 118, 120 Miller, R.D., 25, 36 Miller, R.F., 417–419, 450, 557, 564, 566, 568 Miller, R.I., 714, 719 Miller, R.M., 535, 542 Miller, S.L., 190, 198, 212, 220 Miller, T., 337, 357 Miller, T.E., 613, 615, 616, 630 Miller, W.P., 128, 133 Milne, R., 188, 219 Milner, A.M., 25, 36 Milton, K., 230, 232, 233, 250 Milton, S.J., 281, 283, 645, 653 Minchin, P.R., 190, 192, 201, 215 Minckler, L.S., 202, 205, 219 Minckley, W.L., 432, 449 Minello, T.J., 343, 363 Mingelgrin, U., 507, 518, 530, 540 Mining Annual Review, 365–367, 383 Minnich, R., 429, 432, 448 Minnich, R.A., 278, 284, 317, 328, 423, 439, 446 Minore, D., 190, 211, 219 Minshall, G.W., 294, 304, 427, 450, 713, 719 Miotke, F.-D., 52, 111 Mirmanto, E., 142, 149, 154, 159 Mir´o, C., 64, 95

Mirreh, H.F., 507, 518 Mirsky, A., 17, 35 Mitchel, P., 419, 448 Mitchell, C., 369, 384 Mitchell, D.S., 681, 685 Mitchell, E.D., 58, 111 Mitchell, G.A., 52, 110 Mitchell, J.K., 127, 134 Mitchell, P.S., 432, 445 Mitchell, R.J., 618, 628 Mitchell, S.J., 172, 185 Mitchell, W.W., 52, 90, 110 Mitsch, W.J., 331, 332, 334, 336, 355, 358, 361, 428, 450 Mizuno, N., 145, 155, 159 Mladenoff, D.J., 189, 204, 205, 212, 213, 219 Moad, A., 469, 478–480, 485 Moe, A., 57, 96 Moeller, R.E., 193, 215 Moen, J., 169, 185 Moermond, T.C., 240, 249 Moffat, D., 666, 670 Moffett, M., 255, 268 Mohammed, D., 513, 518 Mohammed, G.H., 615, 629 Mohler, C.L., 615, 616, 618, 630 Moiroud, A., 27, 35 Mokma, D.L., 126, 134 Moldenhauer, W.C., 124, 128, 133, 134 Molina, R., 522, 538 Molina, R.J., 529, 542 Molion, L.C.B., 323, 326 Moll, J.B., 379, 383 Molofsky, J., 240, 250 Moloney, K.A., 386, 395 Molope, M.B., 507, 518 Molski, B., 65, 111 Monastersky, R., 587–589, 607 Monda, M.J., 343, 346, 361 Mondrag´on Castillo, M.X., 60, 105 Monger, H.C., 322, 326 Monk, C.D., 616, 622, 624, 630 Monroe, W.H., 590, 607 Monson, R.K., 307, 310, 329 Montaldo, N.H., 498, 502 Montalvo, A., 209, 220, 229, 231, 232, 234, 238, 239, 250 Montenegro, G., 274, 284 Montpetit, D., 531, 544 Mooney, H.A., 271, 272, 278, 283, 284, 287, 306, 322, 328, 421, 422, 445, 448, 488, 502, 571, 582, 645, 653, 681, 685 Moore, D.M., 69, 99 Moore, J., 66, 112 Moore, J.C., 522, 530, 534, 535, 541, 542 Moore, M.M., 457, 464 Moore, M.R., 208, 219

AUTHOR INDEX Moore, P., 599, 607 Moore, P.N., 711, 720 Moore Jr, T.S., 524, 538 Moore-Landecker, E., 530, 542 Moorhead, D.L., 645, 654, 756, 767 Moran, E.F., 467, 485, 667, 670 Morehouse, T.A., 58, 107, 111 Moreira, I.S., 430, 447 Morelli, F.A., 27, 33 Moreno, J.M., 272, 284, 425, 450 Moreno-Casasola, P., 589, 599, 607 Morgan, J.K., 126, 133 Morgan, V.I., 67, 111 Mori, S.A., 229, 250 Morin, H., 167, 169–171, 182, 185, 751, 766 Morita, R.Y., 68, 90, 103 Moriwaki, K., 51, 110 Morley, C.R., 530, 534, 541 Morneau, C., 169, 175, 185 Morrell, T.E., 636, 653 Morrey, D.R., 375, 376, 382, 383 Morris, E.C., 617, 619, 620, 630 Morris, H.M., 130, 134 Morris, J.T., 337, 357 Morris, M., 599, 607 Morris, P.A., 386, 388, 389, 395 Morris, S.J., 524, 542 Morris, W.F., 141, 152, 153, 159, 160, 592, 594, 610 Morrison, R.G., 151, 152, 160, 591, 610 Morrison, S.J., 45, 88, 111 Morrison, W.H., 510, 519 Morrow, L.A., 611, 629 Morrow, P.A., 255, 267 Mortensen, D.A., 87, 89, 96 Mortimer, M., 615, 628 Mortimore, M., 662, 668, 671 Morton, S.R., 429, 431, 448, 675, 686 Moser, M.B., 323, 327 Moser, W., 40, 43, 111 Moskal, W., 43, 115 Moskalenko, N.G., 78, 111 Moss, A.J., 514, 518 Mosse, B., 525, 528, 542, 543 Mott, K.A., 202, 220 Motzkin, G., 333, 350, 361 Mounts, F.L., 428, 431, 448 Mousain, D., 524, 542 Moya, S., 243, 251 Moyle, P.B., 432, 445, 446, 450 Muckenhausen, E., 510, 517 Mueller-Dombois, D., 142, 152–154, 158–160, 260, 263, 264, 266, 267, 268, 435, 438, 443, 450, 451, 577, 582, 599, 609 Muir, D.C.G., 64, 111 Muir, P.S., 462, 466

803 Mulder, C.P.H., 83, 111 Muller, C.H., 459, 464, 573, 581 Muller, H.K., 66, 111, 121 M¨uller-Wille, L., 62, 111 Mullinax, B., 90, 101 Mun, H.T., 645, 654 Munn, C.A., 474, 486 Munsee, J.R., 646, 647, 656 Munshower, F.F., 680, 686 Murakami, Y., 530, 544 Murawski, D.A., 472, 485 Murdock, L.W., 126, 133 Murie, A., 529, 542 Murozumi, M., 64, 111 Murphee, C.E., 126, 134 Murphy, D., 718, 721 Murphy, E.C., 58, 111 Murphy, K.J., 432, 450 Murphy, P.C., 64, 101, 112 Murphy, P.G., 239, 250, 471, 485 Murphy, S.M., 62, 100 Murray, K.G., 474, 483, 642, 643, 653 Murray, R.S., 527, 542 Muscutt, A.D., 682, 686 Musgrave, G.W., 127, 134 Musil, C.F., 274, 284 Musselman, R., 187, 222, 228, 252 Muszick, D.A., 682, 687 Mutch, R.W., 164, 185, 615, 630 Muzika, R.M., 591, 607 Myers, J.H., 418, 450, 579, 580, 582 Myers, K., 438, 443, 450 Myers, N., 123, 125, 131, 134, 667, 670 Myers, R.L., 336, 361, 433, 434, 450 Myrcha, A., 53, 113 Myster, R.W., 124, 135, 590, 592, 609, 613, 615, 616, 622, 624, 625, 630, 717, 722 Mysterud, I., 43, 55, 112 Mytton, W.R., 387, 395 Nadelhoffer, K.J., 546, 565 Nadian, H., 527, 542 Nadkarni, N.M., 227, 228, 231, 232, 234, 238, 249 Nadolny, C., 266, 268 Nadvornyi, V.G., 532, 541 Naeth, M.A., 556, 568, 644, 654 Nagy, K.A., 646, 656 Naiman, R.J., 169, 170, 184, 185, 419, 420, 427, 429, 430, 447, 450, 577, 580, 582, 583, 646, 655, 681, 685 Naka, K., 188, 195, 202, 219 Nakamura, K., 471, 475, 484 Nakamura, T., 142, 159 Nakashizuka, T., 142, 145, 159, 188, 195, 199, 204, 208, 209, 211, 215, 219, 222 Nakata, J.K., 316, 328, 388, 396

Nanson, G.C., 592, 599, 607 Napolitano, S., 131, 134 Naruse, R., 66, 94 National Academy of Sciences, 692, 705 National Foreign Assessment Center, Central Intelligence Agency., 57–59, 62, 63, 66, 90, 111 National Research Council, 702, 705 National Science Foundation, 70, 111 National Soil Erosion-Soil Production Research Planning Committee, 126, 134 Nauman, L.E., 334, 360 Navaratnam, S.J., 436, 450 Navarro, E., 64, 95 Navas, M.L., 274, 285 Naveh, Z., 273, 275–277, 284, 595, 607 Nay, S.M., 26, 31, 33 Neal, J.L., 78, 104 Neal, J.T., 313, 328 Neal, O.R., 127, 134 Nebeker, G.T., 536, 541, 563, 564, 567, 569 Nebeker, J.T., 543 Nechaev, A.P., 592, 599, 607 Negi, G.C.S., 82, 83, 111, 114 Negri, G., 17, 35 Nel, E.M., 253, 260, 269 Nelson, B.W., 225, 228, 250, 471, 485, 713, 721 Nelson, L.E., 462, 466 Nelson, W.W., 510, 513, 515, 517, 519 Nepal-Australia Forestry Project, 480, 485 Nepstad, D., 614, 631 Nepstad, D.C., 467, 476, 477, 485, 486, 665, 670 Neri, A., 155, 158 Neris, L.E., 592, 609 Neubert, M.G., 40, 111 Neumann, J.C., 645, 647–649, 654 Neuvonen, S., 84, 119, 171, 184 New York Times, 702, 705 Newcomb, W.W., 312, 324, 329 Newell, S.Y., 353, 360 Newman, E.I., 575, 582 Newman, J.A., 421, 422, 446 Newman, L.J., 639, 655 Newman, P., 735, 746 Newson, M., 666, 670 Newton, M., 462, 465, 466 Ng, A., 64, 111 Ngakan, P.O., 140, 142, 144, 146, 149, 160 Ngoile, M.A.K., 337, 361 Ngondi, J.G., 337, 360 Nicholas, D., 523, 525, 543 Nicholas, P.J., 535, 542 Nichols, L., 58, 111 Nichols, O., 381, 383, 674, 684

804 Nichols, O.G., 674, 687 Nicholson, S.E., 312, 328 Nicholson, T.H., 530, 532, 542 Nickling, W.G., 19, 35 Nicolait, L., 472, 484 Nicolait, R., 472, 484 Nielsen, B.O., 255, 268 Nielsen, C.O., 64, 100 Niemel¨a, P., 55, 84, 97, 111, 119, 171, 184 Niemel¨a, T., 408, 410 Nieminen, M., 62, 106 Nienow, J.A., 45, 111 Niering, W.A., 595, 598, 599, 607 Niesenbaum, R.A., 680, 686, 710, 721 Nihlen, T., 314, 328 Niiyama, K., 188, 199, 208, 219 Nilssen, A., 172, 186 Nilsson, A., 169, 185, 549, 554, 568 Nilsson, C., 170, 185, 617, 619, 620, 630 Nilsson, M.C., 166, 186 Nilsson, S., 460, 466 Nisbet, I.C.T., 595, 598, 605, 611, 621, 629 Nissley, S., 457, 465 Niswander, S.F., 428, 450 Nitecki, M.J., 42, 111 Nizeyimana, E., 126, 128, 134 Nobel, P.S., 310, 311, 327, 328 Noble, I.R., 596, 597, 604, 607, 613, 616, 631, 678, 680, 686 Noble, J.C., 271, 284 Noble, M.G., 22, 30, 31, 35, 600, 607 Noble, R.R., 236, 238, 240, 242, 250 Noel, J.M., 341, 361 Noest, V., 153, 159 Nohrstedt, H., 549, 554, 568 Noland, T.L., 615, 629 Noon, B., 718, 721 Nordheim, E.V., 193, 219 Nordlander, G., 548, 554, 569 Norgaard, R.B., 743, 746 Norman, M.J.T., 646, 655 Norstrom, R.J., 64, 111 Northington, D.K., 325, 327 Norton, A., 263, 268 Norton, B.G., 676, 684 Norton, D.A., 209, 219, 674–676, 678, 685 Norton, M.L., 188, 209, 210, 215 Norton, R.B., 64, 112 Norton, T.W., 417, 449 Noss, R.F., 675, 686 Nothnagle, P.J., 716, 721 Nowak, R.S., 317, 329, 442, 446 Noy-Meir, I., 287, 302, 305, 306, 309–311, 326, 328, 329, 675, 678, 687 Nriagu, J.O., 64, 111 Nuhn, W.W., 151, 160

AUTHOR INDEX Nullet, D., 140, 149, 159 Nulsen, R.A., 281, 284 Numata, M., 397, 405, 411 Nu˜nez-Farfan, J., 714, 721 Nussbaum, R., 472, 485 Nuttall, M., 58, 111 Nuttle, W.K., 337, 357 Nyk¨anen, M.-L., 172, 185 Nyland, R.D., 460, 465 Nyman, J.A., 336, 337, 342, 349, 353, 355, 361 Oades, J.M., 506, 519 Obara, H., 407, 411 Oberbauer, S.F., 87, 89, 111, 594, 605 Oberwinkler, F., 524, 543 Oboyski, P., 257, 268 O’Cinn´eide, M.S., 188, 200, 219 O’Connor, I., 625, 627, 630 O’Connor, K.F., 90, 111 Odion, D.C., 278, 283, 423, 425, 426, 445, 447 Odland, A., 21, 35 O’Dowd, J., 426, 449 Odum, E.P., 39, 40, 111, 129, 134, 295, 305, 341, 361, 487, 502, 531, 533, 535, 538, 539, 542, 595, 597, 607, 755, 766 Odum, H.T., 235, 250, 255, 268, 731, 732, 746 Odum, W.E., 336, 346, 362, 441, 447 Oechel, W.C., 52, 66, 87, 89, 98, 103, 111, 118, 272, 284, 322, 328, 425, 450 Oesterheld, M., 287–291, 295, 303, 305, 306, 418, 451, 497, 501, 502 Ofﬁcer, C., 2, 16 Oﬂas, S., 142, 159 Ogden, J.C., 683, 685 Ohlson, M., 173, 185 Ohmann, J., 457, 465 Ohmart, C.P., 255, 260, 263, 268 Ohsawa, M., 142, 143, 147, 159 Ohtani, S., 54, 94 Ohtonen, R., 169, 170, 186 Ohyama, Y., 54, 94 Oikawa, T., 589, 606 Ojima, D.S., 243, 245, 251, 288, 291, 292, 294, 302, 304–306, 435, 450 Oke, T.R., 25, 35, 399, 411 Oksanen, J., 170, 186 Oksanen, L., 55, 84, 111, 112, 169, 183, 185 Oksanen, T., 55, 112 Okuda, S., 141, 159 Olah-Zsupos, A., 130, 134 Old, K.M., 260, 263, 268, 530, 542 Oldeman, L., 129, 134 Oldeman, R.A.A., 231, 239, 242, 250 O’Leary, J.F., 40, 120, 317, 328

Olech, M., 58, 88, 112 Olff, H., 589, 596, 599, 607, 617, 619, 620, 630 Olin, G., 599, 608 Oliveira, P.D., 475, 483 Oliveira, W.J., 225, 228, 250, 471, 485, 713, 721 Oliver, C.D., 17, 27, 29, 35, 161, 164, 172, 185, 187, 189, 193, 219, 223, 240, 250, 454, 456, 460, 465 Oliver-Smith, A., 666, 670 Olmsted, I., 228, 235, 239, 241, 244, 245, 248, 252 Olney, P.J.S., 681, 686 Olson, A.M., 618, 629 Olson, D.L., 199, 221 Olson, J., 456, 465 Olson, J.S., 596, 600, 607 Olson, J.T., 675, 684 Olson, K.R., 126, 128, 134 Olson, M.S., 344, 361 Olson, S., 426, 450 Olson, T.E., 428, 431, 449 Olsson, B.A., 551, 554, 569 Olsson, K., 665, 670 Olszewski, J.L., 213, 220 Oltmans, S.J., 64, 65, 105, 112 Onderdonk, J.J., 127, 133 Ondrusek, M.E., 42, 66, 87, 96, 100 O’Neil, J.A.S., 500, 502, 729, 746 O’Neil, L.J., 386, 388, 389, 395 O’Neil, T., 336, 337, 343, 360 O’Neill, J., 125, 130, 134, 226, 234, 252 O’Neill, R., 661, 669 O’Neill, R.V., 4, 16, 524, 541, 597, 607, 712, 716, 718, 721, 722, 725, 746 Oner, M., 142, 159 Ong, C.K., 488, 489, 497, 502 Onsi, D.E., 125, 130, 134, 500, 502, 729, 746 Onstad, C.A., 123, 124, 129, 133 Oosting, H.J., 590, 607, 713, 721 Opdam, P., 681, 687 Oppenheimer, M.J., 618, 630 Orheim, O., 66, 112 Orians, G.H., 244, 250, 476, 485, 611, 630, 694, 705 Orio, A.A., 1, 15 Ormsby, J.P., 90, 103 Orombelli, G., 89, 95 Orozco-Segovia, A., 212, 222 Orwig, D., 190, 193, 194, 221 Orwig, D.A., 169, 182, 190, 205, 214, 217, 220 Osawa, A., 196, 221 Osborn, H., 639, 655 Osbornov, J., 611, 630 Osherenko, G., 57, 112

AUTHOR INDEX Osmond, C.B., 472, 484 Ossemerct, C., 510, 517 Østbye, E., 43, 55, 112 Ostera, H.A., 143, 159 Osterkamp, W.R., 427, 448, 590, 599, 605 Osterman, S., 170, 183 Ostfeld, R.S., 718, 721 Østreng, W., 57, 96 O’Toole, R., 699, 705 Ott, M.L., 337, 361 Ott, V.J., 52, 110 Ouellet, J.-P., 69, 112 Ovalle, C., 279, 280, 282, 284, 325, 325, 674, 676, 684 Overpeck, J., 66, 112 Ovington, J.D., 453, 465 Øvstedal, D.O., 88, 112, 117 Owens, S., 386, 395 Owensby, C., 641, 656 Owensby, C.E., 288, 294, 297, 302, 304, 305, 435, 450, 640, 641, 655 Oxford World Atlas, 6, 16 Pacala, S.W., 202, 215, 615, 630 Pace, L., 341, 361 Pace, M.L., 336, 361 Pacenza, M., 724, 746 Pacey, A., 666, 669 Packham, J.R., 172, 184 Pacovsky, R.S., 525, 542 Padbury, G.A., 167, 185 Padney, C.B., 555, 568 Padoch, C., 471, 479, 483 Page, E.R., 507, 518 Page, J., 2, 16 Pahl-Wostl, C., 678, 686 Paige, K.N., 579, 582 Paine, R.T., 131, 134, 333, 361, 707, 720 Paine, T.D., 253, 258, 260, 268 Painter, E., 299, 305 Palaniappan, V.M., 599, 607 Palazon, J.A., 354, 357 Palber, T., 58, 112 Paling, E., 343, 349, 363 Pallant, L., 226, 229, 231, 249 Pallardy, S.G., 334, 358, 512, 517 Palmer, J.P., 427, 429, 450 Palmer, M.E., 368, 384 Palmer, R.S., 514, 518 Palmiotto, P.A., 195, 220 Palmisano, A.C., 42, 45, 112, 117 Panayiotopoulos, K.P., 506, 518 Pandey, C.B., 296, 305 Pandey, H.N., 188, 202, 206, 215 Panﬁl, S.N., 229, 248 Paniagua, J.M., 64, 95 Panton, W.J., 438, 443, 446 Paoletti, M.G., 526, 542

805 Paone, J., 508, 518 Papanastasis, V., 276, 284 Papanastasis, V.P., 298, 305 Papendick, R.I., 488, 502 Papp, R.P., 53, 112 Pargney, J.C., 524, 540 Parica, C.A., 143, 159 Pariona, A.W., 469, 478, 484 Parisella, S., 175, 182 Park, C., 659, 670 Park, D.G., 371, 383, 599, 604 Parker, B.C., 41, 45, 58, 63, 112, 120 Parker, C.A., 681, 686 Parker, G., 244, 250 Parker, G.G., 258, 262, 268 Parker, I.M., 415, 416, 421, 422, 444, 450 Parker, L.W., 526, 536, 540 Parker III, T.A., 474, 486 Parker, V.T., 189, 193, 212, 218, 219, 240, 247, 675, 676, 686 Parkinson, C.L., 66, 98 Parkinson, D., 532, 543 Parkinson, J.A., 548, 554, 568 Parks, C.G., 262, 268 Parmelee, D.F., 70, 112 Parmenter, R.R., 319, 328, 593, 607 Parnell, J.A.N., 419, 420, 450 Parodi, L.R., 489, 491, 495, 497, 499, 502 Parrington, J.R., 123, 134 Parrish, D.D., 64, 101, 112 Parrish, J.A.D., 615, 622, 624, 625, 630 Parrotta, J.A., 211, 222, 238, 243, 245, 250, 467, 479, 485 Parsons, A.J., 313, 328 Parsons, A.N., 89, 121 Parsons, D.A., 423, 425, 450 Parsons, R.F., 681, 684 Parsons, W.F.J., 17, 34, 190, 198, 212, 216, 220 Parsons, W.J.F., 532, 543 Partomihardjo, T., 140, 142, 147–149, 154, 157, 159, 160, 599, 608 Parton, W., 292, 305 Parton, W.J., 243, 245, 251, 287, 291, 292, 294, 298, 303, 304, 306, 435, 450, 555, 568 Paruelo, J., 295, 304 Paruelo, J.M., 287, 296, 300, 302, 305 Passel, C.F., 50, 116 Pastor, J., 89, 112, 169, 170, 184, 577, 580, 582, 583, 646, 655 Pastore, G., 729, 746 Pate, J.S., 271, 284, 680, 685 Paternoster, M., 17, 35 Patric, J.H., 546, 565 Patrick Jr, W.H., 332, 336, 345, 349, 358, 361 Patten, D.T., 428, 431, 438, 450, 451

Patterson, C.C., 64, 97, 111 Patterson, S., 344, 359 Patterson, W.A., 333, 350, 361 Patton, D.R., 462, 465 Patton, G.W., 65, 96 Paul, E., 522, 542 Paul, E.A., 294, 304, 522–525, 542 Paul, M.A., 22, 33 Pauley, E.F., 190, 194, 195, 197, 200, 201, 211, 215, 220, 221 Pauli, H., 88, 102 Paulitz, T.C., 530, 542 Paulsen, G.M., 297, 305 Paulsen Jr, H.A., 59, 112 Pavlik, B.M., 589, 594, 603, 607 Payette, S., 161, 164, 165, 169, 174, 175, 183, 185, 189, 193, 220 Payne, J., 596, 607 Payne, M.R., 91, 112 Payne, W.J., 353, 359 Payton, I.J., 90, 112 Peakall, D.B., 68, 112 Pearcy, R.W., 190, 215 Pearman, G.I., 66, 101 Pearsall, W.H., 26, 36 Pearson, H.A., 574, 583, 639, 654 Pearson, M.L., 386, 388, 389, 395 Pearson, S.M., 354, 361 Peart, D.E., 438, 450 Peart, D.R., 189, 194, 195, 219, 220, 239, 250, 419, 421, 422, 450, 645, 655 Peat, A., 45, 112 Pecarsky, J., 724, 746 Peck, A.J., 281, 284 Peckham, G.E., 65, 115 Pedersen, J.M., 70, 116 Pederson, N., 190, 194, 221 Pedraza, C.R., 238, 248 Peek, J.M., 713, 719 Peel, D.A., 67, 112 Peet, R.K., 169, 185, 187, 217, 333, 363, 487, 488, 502, 572, 583, 585, 597, 598, 605, 607, 611, 616–618, 625, 628–630, 632, 717, 720 Pegau, R.E., 55, 82, 84, 112 P˛ekala, K., 24, 33 Pekelder, J.J., 337, 357 Pelaez Menendez-Riedl, S., 89, 108 Peltola, H., 172, 185 Pemble, R.H., 640, 641, 656 Pemesal, M.B., 143, 159 Penaﬁel, S.R., 436, 443, 448 Penaloza, D., 44, 112 Pendry, C.A., 229, 250 Peppers, L.L., 637, 655 Peralta, R., 190, 197, 203, 211, 219, 224, 228, 229, 233, 237, 242, 249, 642, 654 Percy, J.A., 68, 90, 112, 113, 120

806 Percy, R.W., 202, 220 Pereira, C.A., 476, 485 Perez, J.M., 212, 222 P´erez Corona, M.E., 294, 305 Perez-Trejo, F., 275–277, 284 Perfect, E., 25, 36 Perfecto, I., 236, 238–242, 247, 252 Perkins, A.L., 228, 251 Perkins, R.D., 190, 205, 220 Perrin, M.R., 351, 357 Perry, D.A., 472, 483, 694, 704, 753, 766 Perry, R.A., 753, 766 Persiani, A.M., 472, 484 ˚ 17, 28, 36 Persson, A., Persson, H., 90, 116 Perumpral, J.V., 507, 515 Pescatore, T., 155, 159 Petch, J.R., 19, 35 Peterken, G.F., 174, 185 Peterman, R.M., 417, 419, 452 Peters, A.J., 312, 328 Peters, I.D., 278, 284 Peters, R.W., 388, 390, 395 Petersen, R.G., 646, 655 Peterson, B.J., 90, 106, 354, 358 Peterson, C.J., 190, 191, 195, 196, 199, 203, 205, 208–212, 220, 240, 242, 250, 613, 630, 713, 721 Peterson, D.L., 616, 617, 621, 623, 624, 630 Peterson, K.M., 45, 87, 89, 96, 113 Peterson, M.A., 153, 158 Peterson, N.J., 640, 641, 655 P´etillon, Y., 618, 628 Petit, N.E., 681, 686 Petraitis, P.S., 624, 631, 680, 686, 710, 721 Petrov, M.P., 309, 323, 328 Petty, W.H., 239, 252 Pfeiffer, W.J., 636, 655, 752, 766 Philipson, J., 290, 304 Philipupillai, J., 349, 361 Phillips, D., 55, 114 Phillips, D.L., 132, 134, 259, 269 Phillips, J.J., 545, 547, 549, 551, 566 Phillips, O.L., 229, 234, 250 Phillips, S.L., 312, 326 Pi, N.L., 58, 92, 113, 121 Piche, Y., 531, 544 Pickard, J., 31, 36 Pickart, A., 441, 452 Pickering, S.P.C., 59, 64, 91, 100 Pickett, S.T.A., V, VI , 2, 4, 16, 31, 36, 139, 159, 161, 185, 187, 190, 191, 195, 196, 198, 199, 202, 203, 205, 208–212, 216, 217, 220, 222, 223, 224, 236, 250, 252, 307, 329, 333, 334, 339, 340, 349, 355, 361, 363, 386, 389, 395, 396, 397, 411,

AUTHOR INDEX 412, 414, 452, 454, 466, 467, 485, 488, 502, 585, 586, 594–598, 602, 607, 608, 611–613, 615–618, 622, 624, 625, 627, 628, 630, 633, 635, 642, 655, 656, 675, 676, 679, 680, 686, 707, 709–713, 715, 717, 718, 719, 721, 750, 762, 764, 766, 767 Pickup, G., 675, 686 Pickup, J., 45, 113 Piehl, J.L., 639, 654 Pielou, E.C., 588, 608 Pieper, R.D., 417, 419, 452 Pierce, R.S., 546, 547, 552, 554, 566, 567, 711, 721 Pierpont, N., 474, 486 Pierson, E.D., 131, 132, 636–638, 653, 655 Pierson, J.M., 56, 120 Pierzynski, G.M., 369, 384 Piest, R.F., 127, 135 Piest, R.G., 127, 134 Pieterse, A.H., 432, 450 Pietr, S.J., 53, 113 Pignatelli, O., 22, 36 Pike, J., 378, 382 Pilatti, M.A., 496, 502 Pilkey, O.H., 2, 15 Pimentel, D., 123, 125, 126, 128–132, 134, 724, 726–731, 737, 738, 745, 746 Pimentel, D.A., 500, 502 Pimentel, M., 724, 746 Pimm, S.L., 40, 113, 195, 220, 354, 361, 678, 686 Pinard, M.A., 478, 485 Pinay, G., 577, 580, 582, 583, 646, 655, 682, 687 Pinchot, G., 477, 485 Pinder, D.A., 337, 361 Pinder III, J.E., 618, 630 Pineda, F.D., 472, 484 Pi˜nero, D., 226, 249 Pinto-Correia, T., 276, 284 Piotrowska, H., 589, 608 Piper, S.L., 131, 132, 133 Piperno, D.R., 467, 483 Pirozynski, K.A., 521, 541 Pisarski, B., 408, 411 Pisek, A., 43, 113 Pitelka, F.A., 84, 95 Pitelka, L.F., 322, 328 Pitul’ko, V., 58, 109 Pity, B., 480, 485 Place, I.C.M., 211, 220 Planty-Tabacchi, A.M., 427, 429, 430, 450 Plassard, C., 524, 542 Platt, K.H., 681, 686 Platt, R.B., 590, 604

Platt, W.J., 189–191, 202–206, 215, 220, 234, 239, 243, 247, 250, 333, 336, 341, 344, 345, 348, 355, 357, 361, 615, 617, 625, 630, 631, 641, 645, 655, 709, 719 Platter, Z.J.B., 697, 705 Pletsch, C., 674, 684 Plice, M.J., 90, 113 Pluth, D.J., 556, 568, 644, 654 Poblete, V., 274, 284 Poesen, J., 513, 516 Poff, N.L., 427, 450 Poiani, A., 279, 280, 283 Poissonet, J., 573, 582 Poissonet, P., 573, 582 Pokarzhevskii, A.D., 532, 541 Pokarzhevsky, A., 130, 132 Polak, A.M., 478, 485 Polis, G.A., 318, 328 Polley, H.W., 557, 568 Pollock, M.M., 577, 580, 583 Pollock, W.H., 64, 96 Polunin, O., 271, 272, 284 Ponnamperuma, F.N., 512, 518 Pons, T.L., 478, 485 Pontanier, R., 325, 325, 436, 446, 674, 676, 684 Pool, D.J., 223, 224, 239, 249 Poole, G., 58, 63, 113, 116 Poore, D., 700, 705 Poore, M.E.D., 226, 250 Poorter, H., 571, 582 Poorter, L., 225–229, 231–234, 238, 242, 249, 250 Pope, P.E., 378, 382, 588, 604 Popma, J., 224, 228, 242, 250 Popov, G.B., 318, 328 Porcella, D., 562, 568, 591, 608 Porteous, T., 681, 686 Porter, M.R., 264, 266, 268 Porter, S.C., 588, 589, 608 Portier, R., 354, 358 Porto, M.E., 525, 539 Poscher, G., 20, 34 Post, A., 27, 36 Post, W.M., 89, 112, 321, 328 Postel, S., 128, 134 Postel, S.L., 673, 686, 734, 746 Potter, J.A., 89, 121 Potter, N., 69, 113 Potts, M., 45, 112 Poulson, T.L., 189, 202–205, 220, 641, 646, 655 Poulter, A.G., 430, 447 Pounder, E.J., 62, 113 Pounds, W.Z., 474, 483, 642, 643, 653 Powell, A.J., 90, 111 Powell, C.L., 525, 542 Powell, J.M., 664, 670

AUTHOR INDEX Power, J., 599, 603 Power, J.F., 514, 516 Powers, B.V., 509, 517 Prach, K., 430, 431, 450, 592, 593, 599, 608, 611, 630, 758, 766 Prather, R.J., 506, 518 Pratt, R.M., 69, 108 Preiss, E., 129, 134 Prentice, I.C., 175, 185, 492–494, 502 Prescott, C.E., 532, 543 Press, M.C., 89, 121 Preston, E.M., 354, 357 Pretes, M., 57, 58, 62, 63, 65, 100, 113, 114 Pretty, J.N., 666, 670 Priano, L.J., 496, 502 Price, L.W., 44, 51, 53, 59, 60, 62, 113 Price, N., 459, 465 Price, P.W., 254, 268 Pridmore, R.D., 45, 105 Primack, R.B., 404, 410 Prince, P.A., 55, 69, 89, 91, 99, 108 Prince, S.D., 475, 483 Prins, H.H.T., 324, 328 Priore, P., 66, 97 Pritchard, D., 668, 671 Pritchett, W.L., 454, 465 Pritts, M.P., 621, 629 Prober, S.M., 418, 450 Prock, S., 89, 113 Proctor, J., 229, 250, 374, 376, 382 Proctor, M.C.F., 333, 336, 348, 360, 361 Prodon, R., 272, 277, 285 Profﬁtt, A.P.B., 681, 686 Prose, D.V., 385, 388, 389, 395 Provost, M.W., 337, 361 Prudhomme, T., 87, 89, 111 Prystupa, P., 496, 501 Publicover, D., 524, 543 Publicover, D.A., 212, 222 Puechcostes, E., 522, 542 Pugesek, B.H., 351, 359 Puhakka, M., 590, 592–594, 606 Pullar, D.M., 561, 569 Pulles, J., 478, 485 Pulliainen, E., 62, 113 Pulliam, H.R., 692, 705, 718, 721 Pulliam, W.M., 211, 222 Purvis, J.C., 195, 220 Putman, R.J., 647, 655 Putnam, R.D., 704, 705 Putt, M., 42, 114 Putwain, P.D., 375, 383, 680, 686 Putz, F.E., 191, 194, 195, 199, 200, 203, 209, 210, 220, 223–234, 237–242, 248–251, 336, 346, 361, 478, 485, 620, 630 Pyatt, D.G., 548, 549, 554, 565

807 Pye, T., 69, 113 Pyne, S.J., 161, 185, 456, 465, 713, 719 Pyˇsek, A., 407, 411 Pyˇsek, P., 404, 411, 430, 431, 450, 592, 608, 758, 766 Qinghong, L., 188, 204, 220 Qualls Jr, C.W., 387, 394 Quesada, A., 42, 87, 113, 119 Quigley, N.C., 56, 113 Quinby, P.A., 164, 166, 185 Quine, C.P., 172, 185 Quinn, C.E., 528, 540 Quinn, J.A., 590, 593, 605 Qui˜nones-Orﬁla, V., 211, 222 Quintana, M., 418, 423, 446 Quintana, M.A., 311, 317, 326 Quirk, W.A., 167, 185 Quispel, A., 17, 34, 600, 601, 606 Raaimakeers, D., 478, 485 Rabassa, J., 17, 27, 36 Rabie, M.A., 366, 384 Rabinovich, J.E., 58, 92, 113 Rabinowitz, D., 348, 361 Racine, C.H., 52, 62, 73, 78, 90, 94, 113, 116, 390, 391, 395 Rackham, O., 460, 465 Rada, A., 444, 445 Radforth, J.R., 76, 96 Radforth, N.W., 76, 96 Radosevich, S.R., 492, 495, 501, 502, 615, 618, 622, 624, 628, 630, 631 Radtke, L.L., 388, 395 Ragsdale, H.L., 590, 594, 608 Rahn, K.A., 51, 64, 113 Raich, J., 459, 465 Raich, J.W., 240, 242, 243, 250, 592, 608 Raikes, J.A., 314, 328 Raillard, M., 43, 55, 83, 88, 106, 113 Railsback, S.F., 65, 113 Rainey, D.G., 645, 653, 655 Rainey, W.E., 131, 132, 636–638, 653, 655 Raisbeck, G.M., 590, 603 Raison, R.J., 575, 583 Rajasilta, M., 590, 592–594, 606 Raju, G., 523, 543 Ralston, C., 457, 465 Ram, J., 82, 114 Ramakrishnan, P.S., 225, 247, 488, 489, 497, 502 Ramsey, P.R., 343, 360 Rance, B.D., 679, 681, 684 Rand, A.S., 475, 484 Randers, J., 664, 670, 723, 746 Rands, G.F., 129, 135 Rankevich, D., 276, 285 Rankin de Merona, J.M., 230, 232, 250

Ranney, J.W., 715, 721 Rannie, W.F., 88, 113 Ranta, E., 84, 111 Ranwell, D.S., 336, 337, 361 Rao, P., 188, 202, 206, 215 Raper, S.C.B., 66, 106, 113 Rapport, D.J., 676, 686 Rasmussen, K.J., 510, 518 Rasmussen, P.E., 510, 518 Rasmussen, R.A., 64, 65, 95, 107, 113 Ratcliffe, D.A., 371, 384 Ratti, J.T., 343, 346, 361 Rauch, T.J., 333, 336, 341, 345, 348, 355, 357 Raup, D.M., 40, 70 Raup, H.M., 47, 113, 209, 210, 220, 595, 608 Rauser, W.E., 376, 383 Raven, P.H., 521, 541 Ravetz, J., 662, 669, 670 Ravetz, J.R., 725, 741, 745 Raw, F., 130, 134 Raynal, D.J., 372, 383 Raynaud, D., 66, 95, 113 Raynolds, M.K., 76, 101 Read, D., 524, 543 Read, D.J., 67, 104, 526, 530, 537, 542 Read, W.G., 65, 115 Reader, R.J., 617, 619, 620, 630, 645, 655 Reagan, D.P., 636, 655, 752, 766 Reardon, R.C., 444, 450 Rebar, C., 645, 655 Rebele, F., 397, 411 Rebella, C.M., 492, 493, 495, 499, 501 Rebelo, A.M., 273, 283 Rebertus, A., 171, 186 Rebertus, A.J., 169, 171, 186, 188, 193, 196, 198, 199, 204, 208, 220, 222, 455, 466, 613, 616, 630, 751, 767 Recher, H., 572, 581 Recher, H.F., 429, 434, 450 Redclift, M., 666, 670 Reddy, K.R., 355, 360 Reddy, M.V., 641, 655 Redente, E.F., 319, 320, 328 Redford, K.H., 473, 485 Redhead, C.S., 60, 121 Reeburgh, W.S., 50, 89, 105, 120 Reed, A., 560, 567 Reed, B.C., 312, 328 Reed, D., 669, 671 Reed, D.J., 332, 334, 337, 362 Reed, J.C., 63, 113 Reed, M., 64, 113 Reeder, W.G., 639, 640, 655 Rees, W., 1, 16, 663, 671 Rees, W.G., 50, 63, 65, 113, 114 Reeve, R.C., 506, 515

808 Reeves, D.W., 511, 517 Reeves, R.R., 58, 59, 111, 114 Regal, P.J., 522, 543 Regier, H.A., 682, 683, 685 Reice, S.R., 427, 450, 714, 721 Reich, P.B., 192, 193, 217, 477, 483 Reich, R.J., 69, 114 Reichardt, P., 84, 97 Reichardt, P.B., 55, 114, 170, 182, 533, 542, 594, 608 Reichel, W.L., 65, 116 Reichholf, J.H., 316, 328 Reichle, D.E., 255, 268 Reichman, G.A., 511, 518 Reichman, O.J., 645, 655 Reicosky, D.C., 510, 518 Reid, C.P., 294, 304 Reid, M., 253, 260, 269 Reid, M.S., 169, 171, 186, 455, 466, 751, 767 Reid, W., 593, 600, 608, 647, 655 Reid, W.S., 130, 134 Reid, W.V.C., 691, 705 Reidd, C.P.P., 555, 566 Reimold, R.J., 337, 346, 362 Reinbold, K.A., 387, 395 Reiners, W.A., 17, 36, 51, 70, 73, 114, 119, 190, 202, 211, 217, 218, 313, 328, 531, 535, 543, 546, 547, 553, 557, 560, 561, 566, 567, 569, 600, 608, 621, 630, 711, 721 Reinert, S.E., 343, 362 Reisigl, H., 88, 102 Reiter, E.R., 50, 60, 67, 114 Reitsma, L., 390, 391, 395 Rejm´anek, M., 40, 43, 68, 69, 114, 117, 142, 144, 148, 159, 278, 284, 413, 426, 450, 590, 591, 608, 611, 618, 631, 758, 766 Remmert, H., 69, 119, 716, 721 Rempel, G., 57, 114 Remy, W., 521, 543 Renard, K.G., 513, 514, 518 Rencz, A.N., 337, 358 Renger, M., 401, 411 Renvall, P., 173, 185 Resh, V.H., 343, 357, 427, 450 Resosudarmo, P., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Resvoll, T.R., 43, 114 Retelle, M., 66, 112 Rex, R.W., 51, 114 Reynolds, B., 548, 554, 568 Reynolds, H., 435, 440, 447 Reynolds, J.F., 311, 322, 328 Reynolds, T.D., 555, 565 Rheney, J.W., 618, 630 Rhoades, D.F., 258, 268

AUTHOR INDEX Rhoades, J.D., 506, 518 Ricciardi, M., 140, 142, 159 Rice, K.J., 418, 421, 451 Rice, L.A., 640, 655 Rice, S.D., 68, 114 Rich, P.M., 223, 239, 243, 244, 247 Richard, J.L., 17, 30, 36 Richards, B.N., 130, 134, 523–525, 542 Richards, G.C., 643, 653 Richards, J.F., 337, 362, 717, 721 Richards, J.H., 318, 326, 442, 446, 555, 568 Richards, J.R., 673, 687 Richards, K., 2, 16, 142, 145, 154, 160, 591, 599, 609 Richards, L.A., 507, 518 Richards, P.W., 225, 227–229, 231–234, 238, 239, 242, 245, 251, 469, 485, 596, 608 Richardson, C.J., 546, 568 Richardson, D.M., 69, 114, 272, 274, 281, 284, 285, 418, 424, 425, 436, 439–442, 449, 451, 452 Richardson, J.A., 371, 384 Richardson, S.G., 536, 542 Richter, B.D., 428, 431, 451 Richter, D., 457, 464, 465 Richter, M., 390, 395 Rickard, W., 390, 394 Rickard Jr, W.E., 62, 82, 114 Rickard, W.H., 64, 114, 317, 326, 415, 416, 419, 420, 446, 557, 568 Ricou, G.A.E., 130, 134 Ridley, D.M., 647, 656 Riechers, G., 66, 87, 89, 111 Riechers, G.H., 87, 89, 103, 111 Riechert, S.E., 639, 640, 655 Riecke, F., 403, 411 Riekerk, H., 545, 547, 552, 554, 568 Ri´era, B., 233, 251 Riesbeck, F., 388, 390, 394 Rietkerk, M., 324, 328 Rietsma, C.S., 336, 338, 363 Riewe, R.R., 58, 114 Rifﬂe, J.W., 529, 542 Riggins, R.E., 387–391, 394–396 Riggs, S.C., 599, 607 Rigney, L.P., 139, 142, 145, 148, 158, 592, 605 Rignot, E.J., 66, 114 Rikhari, H.C., 82, 83, 111, 114 Riley, C.V., 367, 384 Rinc´on, E., 522, 532, 538, 539 Ring, E., 549, 554, 568 Rinker, H.B., 255, 268 Riou-Nivert, P., 189, 200, 215 Ripley, E., 292, 305 Ripple, W.J., 462, 463, 466

Risch, S., 259, 268 Risebrough, R.W., 64, 65, 96, 114 Risenhoover, K.L., 378, 383 Risk, M.J., 65, 97 Risley, L.S., 257, 262, 268, 711, 721 Risser, J., 125, 134 Risser, P., 291, 295, 305 Risser, P.G., 288, 292, 294, 298, 306, 377, 383, 555, 568 Riswan, S., 226, 251 Ritcey, G.M., 369, 384 Rittenhouse, L.R., 92, 95, 557, 567 Riviere, A., 144, 151, 154, 159 Rivkin, R.B., 42, 114 Roa, A., 511, 518 Robberecht, R., 65, 98 Robblee, M.B., 233, 235, 244, 245, 251, 333, 334, 336, 345, 347, 348, 352, 355, 358, 362 Roberson, E.B., 507, 518 Roberts, D.W., 154, 157 Roberts, H.H., 349, 361 Roberts, M.J., 573, 582 Roberts, M.R., 478, 485 Roberts, R.D., 372, 382 Roberts, R.J., 264, 266, 268 Robertson, F., 291, 304 Robertson, G.P., 615, 631 Robertson, L., 23, 36 Robertson, P.B., 636, 655 Robin, G. de Q., 67, 114 Robinson, G.G.C., 337, 359 Robinson, G.R., 680, 686 Robinson, J., 456, 464 Robinson, M., 57, 58, 113, 114 Robinson, S.K., 474, 486 Robinson, T.W., 428, 431, 451 Roboson, A.D., 527, 541 Robson, A.D., 381, 383 Rocha Osorio, C.M., 60, 105 Roche, S., 680, 685 Rochefort, L., 560, 567 Rodgers, K.A., 63, 114 Rodin, S., 368, 384 Rodriguez, M.A., 555, 557, 568 Rodriguez, M.J., 64, 95 Roebuck, B.D., 390, 391, 395 Rogers, G.F., 317, 328 Rogers, K.H., 590, 606, 714, 715, 717, 719, 721 R¨ohrig, E., 453, 466 Roland, J., 259, 268 Rollo, C.D., 750, 766 Rolstad, E., 172, 185 Rolstad, J., 172, 185 Roman, C.T., 195, 220, 354, 361 Romans, J.C.C., 23, 36 R¨omermann, H., 403, 411

AUTHOR INDEX Romheld, V., 524, 525, 540, 541 Romme, W.H., 198, 220, 458, 459, 466, 716, 718, 722, 764, 767 Romney, E.M., 562, 568 Romo, J.T., 555, 569 Romoser, W.S., 254, 268 Rondeau, R.J., 310, 326 Rongstad, O.J., 387, 395 Ronkens, M.J.M., 123, 124, 129, 133 Roodman, D.M., 701, 705 Roose, E., 126, 134 Root, T.L., 321, 329 Roovers, L.M., 193, 196, 220 Rorabaugh, J.C., 432, 447 Rosales, J., 477, 485 Rosario, M., 477, 482 Røsberg, I., 21, 35 Rose, A.B., 188, 193, 198, 204, 208, 221, 681, 686 Rose, C.R., 462, 466 Rose, F.L., 390, 396 Rose, J., 21, 23, 24, 36 Rose, S.A., 478, 485 Ros´en, E., 618, 631 Rosen, K., 550, 554, 568 Rosen, S., 429, 434, 450 Rosenau, A.U., 561, 567 Rosenau, R.C., 507, 517 Rosenberg, R., 40, 96 Rosenfeld, G.A., 51, 114 Roser, D.J., 65, 114 Ross, C.W., 510, 519 Ross-Todd, B.M., 546, 552, 554, 566 Rossenaar, A.J.B.A., 555, 567 Rossnes, G., 58, 114 Rosswall, T., 525, 542 Roth, E., 230, 251 Roth, E.R., 239, 251 Roth, L.C., 235, 251, 333, 336, 362 Rothery, P., 26, 33, 59, 64, 69, 89, 91, 99, 100, 108, 110, 591, 604 Rothman, E.D., 602, 603 Rothwell, R.L., 336, 362 Roundy, B.A., 423, 451 Rounsevell, D., 70, 114 Rounsevell, D.E., 91, 114 Rousch, M.L., 615, 630 Rouse, W.R., 43, 114 Roush, M.L., 495, 501 Rousteau, A., 236, 249, 476, 484 Rovira, A.D., 525, 542 Rowe, B., 56, 109 Rowe, J.S., 52, 99, 164, 167, 175, 185 Rowe, W.E., 467, 483 Rowell, T.A., 682, 686 Rowland, E.L., 590, 606 Rowland, N.W., 645, 656 Rowlands, P.G., 388, 394

809 Rowley, J., 90, 114 Rowntree, R.A., 397, 411 Roxburgh, S.H., 78, 114, 589, 608 Roy, C.R., 66, 114 Roy, J., 274, 285 Roy, S., 93, 119 Royalty, A.C., 416, 449 Rozas, L.P., 343, 360 Rozema, J., 337, 361 Rozowski, J.R., 535, 539 Rubilis, S., 25, 37 Rubulis, S., 17, 27, 36 Rudd, N.T., 421, 422, 449, 613, 616, 630 Rudkin, R.A., 63, 115 Rudnicky, J.L., 189, 197, 198, 206, 218 Rudolph, E.D., 51, 53, 115, 141, 157 Rudolph, T.D., 164, 185 Ruesink, A.E., 262, 267 Ruess, R., 43, 83, 88, 106 Ruess, R.W., 288, 305, 555–558, 561, 568 Rufﬁnoni, C., 682, 687 Ruge, U., 403, 411 Ruhland, C.T., 88, 103 Ruiz Reyes, J., 255, 268 Rundel, P.W., 271–273, 278, 279, 283, 285, 316, 317, 328, 329, 575, 576, 583, 589, 608 Runkle, J.R., 187, 189–191, 193, 197, 202–204, 206, 208, 215, 218, 220, 221, 224, 234, 242, 243, 247, 251, 642, 655, 709, 719 Rusch, D., 67, 116 Rusch, G., 645, 655 Rusch, G.M., 497, 501 Rushforth, R.S., 536, 539 Rushforth, S.R., 563, 564, 565 Russell, A.E., 592, 608 Russell, E.J., 508, 518 Russell, E.W., 505, 518 Russell, M.B., 508, 518 Russell-Hunter, W.D., 340, 349, 361, 362 Russell-Smith, J., 438, 442, 443, 451 Russo, M., 496, 501 Russo, R., 480, 485 Rutherford, M.C., 303, 306 Ruyle, G.B., 418, 423, 445, 451 Ryan, M.G., 461, 462, 466 Ryan, P.G., 44, 53, 64, 85, 115 Rybczyk, J.M., 352, 362 Rychert, R., 562, 568, 591, 608 Ryder, J., 21, 33 Rydin, H., 590, 608 Rykiel Jr, E.J., 223, 224, 251, 710, 711, 721 Ryszkowski, L., 682, 686 Ryvarden, L., 26, 28, 34, 36, 591, 608 Sabin, T.E., 261, 262, 269

Sackett, S., 457, 464 Sader, S.A., 589, 608 Sadiq, M., 388, 395 Sadovnikov, Y.S., 532, 544 Sadul, H.A., 346, 359 Safaya, N.M., 377, 384 Safﬁr, H.S., 635, 655 Saffouri, L., 123, 125, 126, 128, 129, 131, 132, 134 Saffouri, R., 729, 746 Sahani, F.M., 143, 159 Sainz, M., 525, 539 Sakai, A., 43, 44, 115 Sala, O.E., 287, 288, 291, 294, 298, 302, 303, 305, 306, 417, 418, 450, 451, 497, 502, 557, 568 Salati, E., 89, 110, 260, 262, 268 Salcedo, I.H., 473, 485 Saldarriaga, J.G., 229, 230, 251 Sale, P.F., 154, 159 Sale, P.J.M., 506, 518 Salinger, M.J., 66, 113 Salisbury, F.B., 41, 115 Salmah, S., 471, 475, 484 Salo, J., 590, 592–594, 606 Salomons, W., 661, 671 Salsac, L., 524, 542 Samenus Jr, R.J., 639, 640, 653 Sampaio, E.V.S.B., 473, 485 Sample, V.A., 675, 684 Samsonowicz, A., 661, 671 Samuel, M.J., 535, 541 San Jos´e, J.J., 477, 485 S´anchez, F., 64, 95 S´anchez, M.J., 245, 249, 752, 753, 766 Sanchez, P.G., 324, 327 Sandberg, C.J., 211, 218 Sanden, J.J. v.d., 478, 485 Sanderson, F.R., 590, 607 Sandhaug, A., 43, 55, 112 Sandi, C.L., 480, 485 Sandoval, F.M., 511, 518 Sandquist, D.R., 312, 326 Sands, W.A., 474, 483 Sanford Jr, R.L., 225, 243, 245, 251, 471, 484 Sanford, W.W., 576, 583 Sanger, C.W., 58, 115 Sannikov, S.N., 164, 167, 185 Santanatoglia, O.J., 496, 501 Santee, M.L., 65, 115 Santianni, D., 52, 100 Santos, A.A., 467, 475, 477, 482 Santruckova, H., 533, 542 Sarig, S., 507, 518 Sarrazin, F., 681, 686 Sarukh´an, J.K., 226, 249, 482, 485 Sasser, C.E., 590, 591, 599, 606, 608

810 Sasser, C.L., 455, 466 Sastre, C., 142, 143, 159 Sato, H., 66, 94 Saunders, D.A., 282, 284, 681, 685, 686, 714, 721 Saunders, S.C., 202, 219 Savage, M., 419, 420, 451 Savelle, J.M., 59, 110 Sawaf, H.M., 536, 542 Sawhney, B.L., 524, 539 Sawtell, N.L., 264, 266, 268 Sawyer, S.A., 229, 250 Sayer, J., 700, 705 Scatena, F.N., 223–226, 228, 230–232, 234, 236, 238, 240–245, 246, 247, 249, 251, 590, 608, 716, 721, 752, 753, 755, 766 Schaetzl, R.J., 187, 221, 244, 251 Schaller, G.B., 44, 58, 115 Schaller, N., 674, 687 Sch¨appi, B., 87, 115 Scharron, R., 642, 643, 657 Schatz, G.F., 60, 121 Schauer, T., 69, 115 Scheffer, V.B., 44, 69, 115 Scheibe, R., 43, 115 Scheid, G.A., 423, 425, 452 Scheiner, S.M., 203, 215, 236, 237, 247, 641, 652 Schell, D.M., 41, 115 Schelling, G.C., 525, 541 Schemske, D.W., 415, 416, 444, 450, 642, 643, 655 Schenck, H., 39, 115 Scherer, S., 42, 45, 115 Scheromm, P., 524, 542 Schertz, D.L., 128, 134 Schiechtl, H.M., 62, 115 Schimel, D.S., 288, 291, 292, 294, 302, 304–306, 322, 325, 435, 450 Schimel, J.P., 525, 533, 541, 542, 594, 608 Schimel, S.S., 557, 560, 561, 566 Schimmel, J., 166, 185 Schimpf, D.J., 314, 328 Schindler, J.F., 64, 85, 115 Schlegel, F.M., 147, 160 Schlenther, L., 401, 411 Schlentner, R.L., 594, 600, 609 Schleshinger, R.C., 202, 219 Schlesinger, W.H., 313, 314, 322, 323, 328, 616, 617, 631 Schmid, M.K., 317, 328 Schmidt, A.M., 90, 115 Schmidt, G.R.B., 661, 671 Schmidt, J.O., 324, 328 Schmidt, S., 43, 115 Schmidt, T.L., 460, 466 Schneider, J.P., 433, 434, 447

AUTHOR INDEX Schneider, R., 64, 115 Schneider, S.H., 321, 329 Schnell, R.C., 64, 95 Schnoor, J.L., 369, 384 Schoch, P., 457, 466 Schoener, T.W., 598, 608 Schoenike, R.E., 17, 34, 600, 601, 606, 608 Schoenly, K., 593, 600, 608, 647, 655 Schonewald, C., 290, 306 Schønning, P., 510, 518 Schoonmaker, D., 47, 115 Schoonmaker, P., 189, 193, 217 Schoonmaker, P.K., 191, 217 Schot, P.P., 682, 684 Schowalter, T.D., 253, 255, 257–263, 268, 269, 318, 329, 643, 655 Schramm Jr, H.L., 334, 357 Schramm, J.R., 594, 608 Schreck, J., 129, 134 Schreiber, L.R., 530, 543 Schroder, D., 510, 517 Schroeder, F.-G., 405, 411 Schr¨oter, C., 75, 115 Schroth, G., 480, 485 Schubert, G.H., 574, 583 Schubiger-Bossard, C.M., 30, 36 Schuhmacher, H., 403, 411 Schuler, R., 524, 543 Schuler, R.T., 561, 568 Schulin, R., 661, 671 Schulte, K.H., 510, 518 Schulte-Karring, H., 510, 518 Schultz, A.M., 55, 115 Schultz, J.C., 716, 721 Schulze, E.-D., 43, 115, 287, 306, 488, 502 Schumacher, T.E., 513, 517 Schuman, G.C., 127, 134 Schuman, G.E., 556, 568 Schumm, S.A., 514, 518 Schupp, E.W., 213, 221, 240, 241, 251 Schuster, J.L., 298, 306 Schuster, W.S., 378, 384 Schuster, W.S.F., 312, 326 Schwackh¨ofer, W., 59, 103 Schwartz, M.W., 239, 250 Schwartz, S.S., 311, 329 Schwartzman, S., 665, 670 Schweickart, R.L., 69, 115 Schweinsburg, R.E., 64, 111 Schwerdtfeger, W., 224, 251 Scifres, C., 288, 303 Scott, A., 525, 540 Scott, A.J., 30, 36 Scott, G.A.M., 590, 607 Scott, H.D., 512, 518 Scott, J.J., 56, 68, 69, 75, 88, 115, 116

Scott, J.M., 390, 395 Scott, L., 88, 115 Scott, M.L., 189, 203, 215, 238, 246, 336, 346, 362, 428, 429, 431, 451 Scott, P.A., 174, 185, 198, 221 Scott, R.W., 17, 36, 75, 116 Scott, T.A., 69, 108 Scott, V., 65, 96 Scotter, G.W., 164, 185 Scrifes, C.J., 638–641, 656 Seaburg, K.G., 41, 45, 112 Seagle, S.W., 288, 294, 305, 306, 555–558, 561, 568 Sear, C.B., 66, 106 Seastedt, T.R., 67, 118, 121, 262, 269, 288, 293–295, 304, 306, 436, 449, 640, 641, 645, 654, 655 Sebacher, D.I., 89, 116 Sebacher, S.M., 89, 116 Secrest, M.F., 637, 643, 655, 657 Sedell, J.R., 140, 142, 151, 152, 158, 173, 184, 211, 218, 711, 720 See, Lee Su, 472, 482 Seely, M.K., 324, 329 Segeren, A.G., 507, 518 Segerstrom, U., 352, 362 Seigler, D.S., 55, 106 Seischab, F.K., 190, 193, 194, 221 Sekarev, A.V., 57, 62, 63, 65, 100 Sekyra, J., 44, 59, 98 Selkirk, D.R., 44, 69, 116 Selkirk, J.M., 52, 94 Selkirk, P., 42, 66, 94 Selkirk, P.M., 44, 56, 66, 69, 91, 94, 116 Semmartin, M., 295, 306 Semmel, A., 188, 209, 221 Sengupta, M., 369, 384 Senst, C.G., 510, 519 Seppelt, P.M., 88, 116 Seppelt, R., 42, 65, 66, 94 Seppelt, R.D., 31, 36, 42, 44, 52, 56, 65, 69, 78, 87, 94, 114, 116 Serey, I., 279, 283 Serrano, M., 477, 482 Serrao, E., 467, 485 Setala, H., 524–526, 543 Seth, K.K., 233, 251 Setterﬁeld, S.A., 681, 684 Severinghaus, M.C., 389, 390, 396 Severinghaus, W.D., 385–390, 394–396 Sevick, S.H., 647, 656 Sexson, T.N., 336, 360 Sexstone, A., 87, 105 Sexstone, A.J., 89, 116 Sexton, J., 235, 244, 245, 248 Sexton, J.C., 524, 538 Sexton, J.M., 262, 269 Sexton, P., 507, 516

AUTHOR INDEX Sferra, P.R., 646, 653 Shachak, M., 131, 134, 295, 306, 313, 325, 327, 329, 680, 685, 708, 713, 716–718, 719–721 Shacklette, H.T., 90, 116 Shaefer, B., 524, 539 Shaffer, G.P., 590, 591, 608 Shafroth, P.B., 428, 429, 431, 451 Shainberg, I., 506, 514, 517, 518 Shainsky, L.J., 622, 624, 631 Shakesby, R.A., 20, 21, 35 Shanan, L., 323, 326, 536, 540 Shanholtzer, G.G., 337, 362 Shanklin, J.D., 63, 65, 101, 114 Shankman, D., 174, 185, 334, 362 Shapley, D., 58, 116 Shariﬁ, M.R., 316, 329 Sharitz, R., 336, 346, 361 Sharitz, R.R., 190–192, 194, 195, 199–201, 203, 210, 215, 220, 221, 224, 230, 239, 250, 333, 334, 337, 349, 358, 360, 362, 590, 608 Sharma, S., 480 Sharman, L.C., 2, 15, 17, 18, 33, 139, 157, 588, 591, 592, 598–602, 604, 613, 628 Sharp, M., 22, 36 Sharpe, D.M., 398, 411 Sharpe, P.J.H., 258, 267 Sharpley, A.N., 436, 443, 451 Shaver, G.R., 78–80, 98, 102, 116 Shaw, G., 507, 516, 524, 543 Shaw, R.B., 385, 386, 389, 396, 416, 451 Shaw, R.J., 28, 36 Shaw, W.B., 224, 228, 239, 251 Shay, J.M., 343, 347, 362, 619, 631 Sheets, P.D., 1, 16 Shefﬁeld, R.M., 195, 221 Shein, G.N., 532, 541 Sheldon, A.L., 427, 450 Sheldon, F., 432, 451 Shepard, R.F., 261, 266 Sheppard, M.I., 64, 66, 116 Sherman, E., 704, 705 Sherrod, C.L., 336, 362 Shi, S.H., 84, 120 Shi, Y.H., 44, 82, 116 Shibata, M., 188, 199, 208, 219 Shields, L.M., 530, 540 Shigesada, N., 758, 766 Shigo, A.L., 239, 251 Shimmel, S., 336, 361 Shipley, B., 617, 619, 620, 630 Shisler, J., 337, 345, 357 Shisler, J.K., 346, 362 Shiver, B.D., 618, 630 Shiyatov, S.G., 189, 193, 221 Shoemaker, V.H., 646, 656

811 Sholes, O.D.V., 190, 210, 215 Showers, W., 66, 97 Shubert, L.E., 562, 568 Shugart, H.H., 161, 182, 321, 329, 437, 451, 712, 722 Shumacher, T.E., 514, 516 Shumway, S.W., 333, 344, 346, 351, 357, 362 Shunula, J.P., 337, 361 Shure, D.J., 259, 269, 590, 594, 608, 614, 616, 618, 631 Shvidenko, A., 460, 466 Siccama, T.G., 243, 244, 251, 752, 766 Siddiqi, M.Y., 545, 567 Siddoway, F.H., 513, 519 Sidle, R.C., 17, 22, 25, 26, 31, 33, 36, 590, 601, 603 Sieburth, J.M., 53, 116 Siegel, B.Z., 64, 116 Siegel, S.M., 64, 116 Siegfried, W.R., 53, 116, 281, 283 Siemer, E.G., 508, 516 Siepel, H., 528, 543 Sierra Nevada Ecosystem Project, 459, 466 Sietz, M.A., 126, 134 Sievering, H., 52, 67, 99, 116 Sigafoos, R.S., 29, 36, 80, 105 Sig¨uenza, C., 532, 543 Sigurdsson, H., 155, 159 Sigurdsson, O., 20, 36 Silander Jr, J.A., 615, 630 Silen, R.R., 199, 221 Silva, J.F., 295, 305 Silva, J.N.M., 467, 472, 485 Silver, W.L., 2, 16, 212, 222, 224, 243–245, 251, 252, 471, 476, 486, 635, 650, 651, 656, 657, 752, 754, 763, 766, 767 Silvertown, J., 612, 631 Simard, A.J., 456, 466 Simberloff, D., 435, 451, 598, 602, 604 Simchenko, Yu.B., 56, 58, 116 Sime, F., 44, 112 Simenstad, C.A., 337, 357, 362 Simkin, T., 2, 16, 587, 608 Simmons, F.W., 510, 518 Simmons, G.A., 171, 175, 184 Simmons Jr, G.M., 41, 45, 112, 120 Simon, E., 376, 384 Simon, J.L., 724, 746 Simon, J.P., 165, 184 Simon, M., 64, 111 Simonetti, J., 279, 283 Simons, M., 123, 134 Simpson, B.B., 622, 624, 631 Simpson, M.A.J., 30, 33, 591, 593, 603 Simpson, R.H., 195, 221

Simpson, R.L., 212, 219 Sims, P.L., 555, 565 Sinclair, A., 287, 295, 306 Sinclair, A.R.E., 346, 359 Sinclair, J., 524, 539 Sinclair, K., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Sinding, K., 63, 116 Singer, F.J., 58, 111, 438, 442, 451 Singer, M.J., 506, 507, 513, 517, 518 Singh, J.S., 296, 305, 555, 568 Singh, R.S., 295, 306 Singh, S.P., 82, 83, 111, 114 Singh, T.V., 124, 134 Singh, V.P., 512, 519 Sinha, S.K., 89, 110 Sinsabaugh, R., 645, 657 Sinun, W., 475, 483 Sionit, N., 87, 111 Sipe, T.W., 189, 198, 206, 221 Siple, P.A., 45, 50, 116 Siren, S., 84, 119 Sir´en, S., 55, 112 Sirois, L., 169, 175, 185 Siron, R., 89, 100 Sivakumar, M.V.K., 702, 705 Sj¨oberg, K., 161, 163, 166, 167, 169, 173, 174, 183, 185 Skar, H.-J., 43, 55, 112 Skartveit, A., 43, 55, 112 Skerbek, W., 288, 302, 305, 311, 327 Skidmore, E.L., 514, 516 Skinner, J.D., 69, 119 Skira, I.J., 69, 97 Sklar, F.H., 354, 358 Skogland, T., 43, 55, 112 Skorupa, J.P., 471, 473, 484 Skre, O., 43, 55, 112 Skuja, H., 90, 116 Skujins, J., 562, 568, 569, 591, 608 Skvarca, P., 66, 94 Skye, E., 88, 116 Sladen, W.J.L., 65, 70, 91, 116 Slater, F.M., 432, 451 Slatyer, R.O., 153, 157, 236, 238, 240, 242, 250, 596–598, 600, 602, 604, 607, 611, 613, 621, 628, 678, 680, 686 Slaughter, C.W., 62, 116 Slavin, P., 346, 362 Slingsby, P., 281, 283 Slocombe, D.S., 701, 705 Slye, R.E., 263, 267 Small, T.W., 187, 221, 244, 251 Smalley, I.J., 47, 116 Smallwood, K.S., 290, 306 Smathers, G.A., 152, 154, 159 Smeins, F.E., 343, 346, 353, 361, 560, 565 Smettem, K.R.J., 509, 517

812 Smil, V., 60, 116 Smirnov, A.B., 168, 186 Smith, A.A., 663, 671 Smith, A.P., 235, 248 Smith, C.J., 345, 358 Smith, C.T., 545, 547, 552, 554, 566–568 Smith, C.W., 423, 439, 440, 442, 443, 451 Smith, D.D., 513, 519 Smith, D.M., 460, 466, 469, 481, 485 Smith, E.F., 640, 655 Smith, E.H., 153, 158 Smith, F., 462, 465, 523, 525, 543 Smith, F.A., 525, 541 Smith, F.W., 189, 204, 215 Smith, G., 316, 329 Smith, G.A., 42, 112 Smith, G.R.T., 432, 451 Smith, H.G., 75 Smith, J.G., 423, 440, 449 Smith, J.L., 143, 153, 158, 159, 294, 304, 488, 502, 557, 560, 563, 566 Smith, J.M.B., 323, 329 Smith, L.M., 334, 342, 346, 348, 357, 359, 362 Smith, M., 89, 116 Smith, M.E., 646, 656 Smith, M.S., 525, 540 Smith, M.W., 47, 121 Smith, R.A., 88, 97 Smith, R.A.H., 374–376, 384 Smith, R.E., 430, 451 Smith, R.I.L., 17, 20, 27, 36, 52, 54, 55, 68, 69, 75, 79, 83, 88, 91, 107, 108, 116, 117, 119, 120 Smith, R.L., 386, 396 Smith, S., 66, 112, 523, 525, 543 Smith, S.D., 307, 310, 317, 329, 428, 431, 446, 590, 604 Smith, S.E., 525, 527, 530, 540, 542 Smith, T., 593, 596–599, 602, 606, 611, 621, 625, 630 Smith, T.G., 56, 64, 90, 102, 117 Smith III, T.J., 195, 220, 233, 235, 244, 245, 251, 333, 334, 336, 345–348, 352, 354, 355, 358, 361, 362 Smith, V.R., 53, 54, 88, 89, 96, 99, 116, 117 Smith, W.H., 462, 466 Smith, W.T., 62, 117 Smock, L.A., 195, 200, 201, 217 Smoliak, R., 555, 557, 569 Smoliak, S., 644, 645, 654 Smuts, J.C., 663, 671 Smythe, N., 255, 267 Snaydon, R.W., 348, 362 Snedaker, S.C., 70, 109 Snook, L.K., 235, 240, 251, 478, 485 Snyder, J.R., 336, 359

AUTHOR INDEX Snyder, M.C., 554, 566 Snyder, W.D., 429, 431, 451 Soane, B.D., 505, 518 Sobaski, S.T., 55, 95 Society of American Foresters, 470, 485 Soderstrom, B., 533, 539 S¨oderstr¨om, L., 173, 185 Soil Conservation Service, 131, 134 Soil Survey Staff, 504, 505, 518 Sojka, R.E., 506, 507, 509–514, 515–517, 519 Solano, R., 227, 228, 231, 232, 234, 238, 249 Solantie, R., 172, 186 Solbrig, O.T., 622, 624, 631, 756, 766 Solhøy, T., 43, 55, 112 Sollins, P., 173, 184, 211, 218, 472, 483, 524, 540, 711, 720 Solomina, O.N., 17, 28, 36 Solomon, A.M., 321, 329 Solomon, S., 65, 105 Sombroeck, W., 129, 134 Sommerville, P., 17, 28, 36 Sondheim, M.W., 17, 36 Soni´e, L., 274, 285 Soo Hoo, J.B., 42, 112, 117 Soo Hoo, S.L., 42, 117 Sorensen, D., 562, 568, 591, 608 Sørensen, T., 43, 117 Soriano, A., 287, 288, 306, 311, 314, 329, 418, 451, 487, 489, 491, 493, 497, 502 Sosa, V., 467, 483 Soto, A., 196, 221 Soule, J.A., 495, 502 Sousa, W.P., 223, 251, 333, 334, 337, 340, 350, 362, 397, 411, 414, 440, 451, 572, 583, 585, 586, 608, 708, 721, 755, 764, 766, 767 South, D.B., 618, 628 South African Mining, 366, 384 Southgate, D., 125, 126, 134, 691, 701, 702, 705 Souza Jr, C., 478, 486 Souza-Pinto, I., 213, 222 S¨oyrinki, N., 79, 117 Spaeti, F., 62, 117 Spaltenstein, H., 26, 31, 33 Sparkes, K.E., 535, 543 Sparks, P.R., 635, 656 Sparks, R.E., 717, 721 Sparks, S.R., 417, 418, 423, 451 Spatt, P.D., 65, 117 Spatz, G., 435, 438, 443, 450, 451 Specht, R.L., 271, 283, 573, 583 Speed, R.J., 682, 685 Speir, T.W., 53, 104 Spellerberg, I., 589, 608 Spellerberg, I.F., 660, 671

Spence, J.R., 17, 28, 30, 36 Spencer Jr, J.S., 460, 466 Spencer, T., 332, 334, 337, 343, 363, 472, 475, 483, 485 Sperry, P.D., 64, 96 Speth, J.G., 125, 129, 135 Spicer, K.W., 434, 446 Spies, T., 190, 216 Spies, T.A., 189–191, 198, 202, 203, 206, 215–217, 221, 229, 234, 243, 247, 251, 462, 463, 466, 641, 656, 709, 715, 716, 719, 720 Spiller, M., 433, 449 Spindler, M., 50, 89, 105, 117 Spingarn, A., 415, 416, 446 Spivey Jr, L.D., 507, 519 Spomer, G.G., 41, 115 Spomer, R.G., 127, 134, 135 Sprenger, B.S., 281, 283 Sprent, J.I., 593, 608 Spritz, L., 123, 125, 126, 128, 129, 131, 132, 134, 729, 746 Sprugel, D.G., 40, 117, 161, 162, 172, 174, 186, 198, 199, 221, 572, 583, 675, 687 Spurr, S.H., 189, 195, 205, 221, 238, 239, 251, 469, 485 Squiers, E.R., 613, 630 Srivastava, D.S., 335, 347, 351, 362 St. Clair, L.L., 536, 541, 543, 563, 564, 567, 569 St. Hilaire, L.R., 211, 221 Staaf, H., 551, 554, 569 Stachow, U., 125, 130, 134, 500, 502, 729, 746 Stadel, C., 57, 94 Stafford, S.G., 262, 267, 269 Stafford Smith, D.M., 429, 431, 448, 675, 686 Staiger, R., 64, 97 Staley, J.T., 43, 117 St˚alfelt, F., 170, 183 Stallard, R.F., 590, 603 Stam, A.C., 60, 121 Standish, J.T., 17, 36 Stanford Biology Study Group, 385, 396 Stanganelli, Z.B., 89, 119 Stanley, E.H., 427, 451 Stanton, G.R., 42, 112 Stanton, N., 255, 269 Stanturf, J.A., 467, 472, 485 Starﬁnger, U., 398, 409, 411 Stark, J.M., 209, 221, 545, 569 Stark, P., 88, 117 Starks, T.L., 562, 568 Stasiak, R.H., 639, 640, 653 States, J., 529, 543 Steadman, D., 426, 451

AUTHOR INDEX Stearns, F., 398, 411 Stearns, F.W., 190, 193, 196, 221 Stebbings, R.E., 337, 362 Steenbergh, W.F., 310, 317, 319, 329 Steenkamp, M., 88, 89, 117 Steiger, J.W., 319, 324, 329 Steijlen, I., 166, 168, 186 Steindl, H., 506, 515 Steinhagen-Schneider, G., 64, 115 Steinitz Kannan, M., 475, 483 Steinke, T.D., 333, 362 Steiof, K., 407, 411 Stelfox, J.B., 167, 186 Stenseth, N.C., 43, 55, 112 Stephens, E.P., 189, 193, 209, 219, 221 Stephens, F.C., 66, 87, 100 Stephens, F.R., 27, 36 Stepniewski, W., 512, 516 Stern, R., 59, 117 Sterner, R.W., 60, 121 Steudler, P.A., 244, 245, 251, 435, 451 Steuter, A.A., 298, 306 Stevens, G.C., 228, 251 Stevens, P.A., 548, 550, 551, 554, 568, 569 Stevenson, J.K., 376, 383 Stewart, A.J.A., 616, 617, 631 Stewart, B.A., 124–126, 128, 129, 131, 133 Stewart, B.S., 91, 96 Stewart, G., 439, 451 Stewart, G.D., 188, 204, 221 Stewart, G.G., 44, 104 Stewart, G.H., 188, 193, 198, 204, 206, 208, 211, 221 Stewart, J.A., 392, 396 Stewart, L.G., 255, 268 Stewart, M.M., 636, 638, 656, 752, 767 Stewart, W., 523, 543 Stewart, W.D.P., 53, 101 Steyer, G.D., 336, 339, 345, 351, 355, 356, 359 Steyn, P., 523, 543 Stigliani, W.M., 661, 671 Stiles, E., 599, 607 Stiles, E.W., 150, 159, 213, 219, 614, 616, 630 Stillwell, M.A., 288, 294, 306 Stinson, D.W., 636, 656 Stirling, I., 65, 70, 109, 117 Stock, W.D., 281, 285 St¨ocklin, J., 17, 36 Stocks, B.J., 164, 186, 456, 465, 466 Stockton, J.R., 510, 516 Stoddard, H.L., 639, 656 Stoffolano Jr, J.G., 254, 268 Stohlgren, T.J., 423, 425, 450 Stokes, B.J., 467, 472, 485

813 Stoll, E., 400, 404, 410 Stolzy, L.H., 388, 394, 507, 508, 512–514, 515, 517, 519 Stone, C.P., 438, 442, 443, 445, 451 Stone, E.L., 209, 210, 215, 221, 713, 719 Stonehouse, B., 70, 89, 117 Stones, G.A., 392, 396 Storck, H.J., 407, 410 Stork, A., 28, 36, 75, 117 Stortelder, A.H.F., 276, 285 Stowe, L.G., 493, 502 Strahm, W., 419, 420, 450 Strahm, W.A., 415, 416, 436, 440, 449 Strahmand, W., 154, 159 Strain, B., 87, 89, 111 Strain, B.R., 317, 325, 594, 605 Strain, I.R., 578, 582 Strasberg, D., 154, 159, 415, 416, 436, 440, 449 Stratil, H., 403, 411 Strauss, R.B., 293, 294, 304 Strayer, D.L., 432, 451 Streibig, J.C., 491, 501 Strelkov, S.A., 65, 117 Streng, D.R., 193, 203, 217 Streudler, P., 546, 565 Streveler, D., 27, 36 Streveler, G.P., 22, 30, 31, 35, 600, 607 Stromberg, J.C., 428, 431, 451 Strong, D.M., 190, 220 Strong, D.R., 336, 358 Strong Jr, D.R., 200, 221, 228, 234, 251 Stroosnijder, L., 324, 328 Strouse, J.J., 229, 248 Struhsaker, T.T., 475, 485 Struik, G.J., 25, 33 Struthers, P., 508, 518 Stryker, J.D., 662, 670 Stuckey, R.L., 428, 431, 451 Studier, E.H., 647, 656 Sturges, D.L., 557, 558, 569 ˇ Stursa, J., 43, 117 Stuth, J.W., 337, 356 Stvan, J., 62, 117 Suarez, J., 17, 25, 27, 36, 37 Su´arez, S.A., 491, 492, 494, 496, 501, 502 Suero, A., 491–494, 501 Sugden, A.M., 129, 135 Sugden, D., 18, 36 Sugg, P.M., 151, 153, 158, 593, 604, 605 Suijdendorp, H., 574, 583 Sukopp, H., 397–399, 401–409, 409–412 Sulkinoja, M., 84, 97 Sullivan, C.W., 42, 45, 50, 105, 112, 117 Sullivan, M.J., 343, 362 Sullivan, R.G., 392, 396 Sullivan, S., 309, 329 Sullivan, T.E., 478, 485

S¨umegi, P., 175, 186 Sumner, M.E., 509, 519 Sun, S.Z., 59, 121 Suriadarma, A., 140, 142, 154, 160, 599, 608 Sussman, R., 426, 449 Sutherland, G.D., 202, 219 Suwa, H., 141, 159 Suzuki, E., 140, 142, 154, 160, 599, 608 Suzuki, W., 142, 145, 159 Svejcar, T., 417–419, 422, 423, 450, 451, 555, 557, 564, 568, 569 Svoboda, J., 20, 33, 39, 43, 55, 62, 64, 83, 87, 88, 105, 106, 113, 117, 118, 120 Swaby, J.A., 640, 641, 656 Swain, A.M., 194, 221 Swain, R.W., 510, 519 Swaine, M.D., 225, 233–237, 240, 242, 249, 251 Swan, J.M.A., 189, 193, 205, 209, 210, 218 Swan, L.W., 53, 117 Swank, W.T., 190, 197, 203, 206, 208, 216, 262, 269, 438, 442, 451, 455, 466, 545, 547, 552, 554, 567–569 Swanson, F.J., 140, 142, 151, 152, 158, 173, 184, 211, 218, 435, 451, 459, 464, 466, 575, 583, 711, 713, 716, 719–721 Swanson, G.A., 332–334, 360 Swanson, J.C., 53–55, 110 Swanson, L.E., 336, 362 Swanson, S.R., 555, 568 Sweazy, R.M., 390, 396 Swetnam, T.W., 164, 174, 186 Swift, D.J.P., 337, 357 Swift, M.J., 487–489, 497, 500, 502, 593, 608 Swift, R.S., 128, 132, 505, 506, 516 Swithinbank, C., 59, 64, 69, 70, 118 Switzer, G.L., 462, 466 Swoboda, A.R., 506, 519 Syers, J.K., 436, 443, 451, 594, 609 Sykes, D.J., 167, 185 Sykes, M.T., 154, 160, 175, 185 S´ykora, T., 43, 117 Syron, W., 636, 652 Syskoowski, B.J., 646, 656 Systematics Agenda 2000, 522, 543 Szacki, J., 529, 543 Szaniszlo, P.J., 524, 540 Taagholt, J., 58, 109 Tabacchi, E., 427, 429, 430, 450, 452 Tackaberry, R., 225, 249 Tadmor, N., 323, 326, 536, 540 Tagawa, H., 140–142, 144, 146–149, 154, 159, 160, 599, 608 Tahvanainen, J., 84, 97

814 Tainter, J.A., 741, 746 Taisacan, E.M., 636, 656 Takacs, D.A., 125, 130, 134, 500, 502, 729, 746 Takahashi, M., 599, 602 Takahaski, F., 198, 221 Talling, J.F., 154, 160 Tallis, J.H., 336, 362 Tamblin, J.F., 43, 95 Tanaka, D.L., 556, 557, 566 Tanaka, H., 188, 199, 208, 219 Tanimoto, T., 142, 145, 159 Tanner, E.V.J., 223, 224, 235, 238–240, 242–244, 247, 251 Tanner, J.E., 762, 767 Tansley, A.G., 595, 608 Tappeiner, J.C., 462, 465 Tarafdar, J., 524, 543 Targulyan, V.O., 209, 222 Tariche, S., 129, 134 Tarifa, R., 472, 486 Tarr, P.W., 324, 327 Tast, J., 55, 88, 106, 118 Tatarko, J., 504, 517 Tateno, M., 143, 153, 159 Tatur, A., 41, 53, 113, 118 Tausch, R., 417, 418, 422, 423, 451 Tavakol, R.K., 66, 106 Tavernetti, J.R., 510, 516 Tay, J., 478, 485 Taylor, A.H., 188, 207, 221 Taylor, B.R., 532, 543 Taylor, B.W., 142, 160 Taylor Jr, C.A., 561, 569 Taylor, C.M., 235, 236, 238, 252 Taylor, D.R., 615, 631 Taylor, G.S., 124, 132 Taylor, H.W., 64, 87, 105, 117, 118 Taylor, K.L., 333, 347, 353, 362 Taylor, R.H., 55, 62, 70, 91, 118, 121 Taylor, R.V., 67, 118 Taylor, S.A., 507, 519 Taylor, T.N., 521, 543 Taylor, W.P., 645, 656 Taynton, K.M., 635, 654 Tazik, D.J., 386, 387, 389, 390, 395, 396 Teas, H., 337, 362 Tecchi, R.A., 489, 502 Teel, P.D., 638–641, 656 Teeri, J.A., 51, 118, 493, 502 Telewski, F.W., 229, 251 Tellier, S., 274, 284 Templet, P.H., 334, 362 Tennesen, M., 390, 396 Tenow, O., 169, 172, 183, 186 Teotia, J.P., 130, 135 ter Braak, C.J.F., 149, 160 ter Heerdt, G.N.J., 352, 362

AUTHOR INDEX ter Steege, H., 478, 485 Terashima, I., 599, 602 Terborgh, J., 229, 251, 473, 474, 486 Tester, J.R., 640, 656 Tevis Jr, L.P., 645, 654 Tharp, M.L., 229, 230, 251 Thatcher, F.M., 367, 384 The Ecologist, 667, 671 Th´ebaud, C., 154, 159, 415, 416, 427, 430, 436, 440, 449, 452 Theobald-Ley, S., 188, 213, 221 Theodorou, C., 525, 539 Therivel, R., 668, 671 Thiault, M., 573, 582 Thiele, K.R., 418, 450 Thieret, J.W., 758, 766 Thiollay, J.-M., 471, 473, 486 Thirgood, J.V., 367, 384, 665, 671 Thomas, B.W., 55, 91, 118 Thomas, D.S.G., 665, 671 Thomas, J.R., 255, 268 Thomas, J.W., 713, 719 Thomas, M.E., 57 Thomas, T., 62, 70, 118 Thomas, W., 386, 396 Thomas, W.L., 1, 2, 16 Thomas, W.P., 128, 133 Thompson, C.H., 594, 609 Thompson, D.B., 319, 327 Thompson, D.J., 343, 347, 362 Thompson, D.Q., 428, 431, 434, 452 Thompson, I.D., 169, 186 Thompson, J.N., 189, 208, 210, 221, 546, 556, 568, 707, 721 Thompson, K., 27, 33, 598, 608, 616, 617, 631 Thompson, L.M., 125, 135 Thompson, M.T., 195, 221 Thompson, R.B., 70, 118 Thompson, S., 668, 671 Thompson, T.M., 65, 103 Thomsen, C.D., 278, 284 Thomson, A.D., 353, 362 Thomson, G.M., 368, 384 Thomson, J.K., 416, 448 Thorarinsson, S., 47, 118 Thordardottir, T., 42, 66, 93, 98 Thorhallsdottir, T.E., 62, 118 Thorn, C.E., 47, 51, 55, 118 Thornburgh, D.A., 462, 465 Thornton, I., 402, 411 Thorsteinsson, I., 52, 95 Thorsteinsson, S., 92, 118 Threadgill, E.D., 510, 519 Thurow, T.L., 387–390, 396, 680, 687 Thurrow, T.L., 555, 560–562, 569 Tiedje, J.M., 522, 543

Tieszen, L.L., 52, 55, 83, 89, 94, 98, 110, 118 Tiffen, M., 662, 668, 671 Tiffney Jr, W.N., 599, 607 Tikhomirov, B.A., 39, 54, 118 Tilbrook, P.J., 43, 95 Tilman, D., 75, 118, 130, 135, 488, 489, 502, 592, 594, 596, 598, 599, 602, 608, 611, 612, 614–617, 619–624, 627, 629–632, 716, 718, 721 Tilmant, J.T., 195, 220, 354, 361 Times Atlas of the World, 6, 16 Timmins, S., 142, 160 Timmins, S.M., 336, 363 Tiner, R.W., 337, 363 Ting, I.P., 261, 269 Tinker, P., 524, 543 Tir´en, L., 168, 186 Tisdale, E.W., 564, 566 Tisdall, J.M., 506, 519 Tisdell, C.A., 442, 445 Tissue, D., 87, 89, 111 Tissue, D.T., 87, 118 Titova, Ye.V., 167, 182 Titus, J.H., 142, 147, 149, 151–153, 158, 160 Tivy, J., 488, 502 Tizon, A., 137, 160 Tjaden, P., 497, 501 Tobiessen, P., 261, 267 Todd, A.W., 524, 540 Todd, C.D., 294, 306 Todd, D.E., 545, 547, 552, 567 Todd, R.L., 262, 269, 524, 540, 546, 547, 552–554, 569 Todesco, M., 155, 158 Tolley, M.D., 345, 358 Tolonen, K., 161, 166, 182, 186 Tolonen, M., 161, 182 Tomanek, G.W., 557, 569 Tomaselli, R., 279, 283 Tomlinson, D.W., 66, 114 Tomlinson, P.B., 348, 363 Tongway, D.J., 561–563, 567, 680, 686, 687 Torgerson, O., 390, 395 Torres, J.A., 257, 259, 269, 593, 608, 636, 637, 656 Torres-S´anchez, A.J., 590, 606 Torri, D., 514, 519 Torsvik, V., 522, 543 Tosi, J.A., 228, 229, 234, 235, 249 T´oth, A., 175, 186 Toth, L.A., 682, 687 Totsuka, T., 172, 184, 198, 218 Touliatos, P., 230, 251 Toumey, J.W., 212, 221 Towne, G., 641, 656

AUTHOR INDEX Towns, D.R., 681, 687 Townsend, C.R., 593, 603 Townson, M.A., 336, 339, 345, 351, 355, 356, 359 Toyoda, T.S., 65, 118 Toyooka, H., 196, 221 Trabaud, L., 272, 277, 285, 422, 423, 452, 573, 582 Trabaud, L.V., 456, 465 Tracey, W.H., 381, 384 Tracy, J.C., 369, 384 Trail, P., 636, 653 Trainer, M., 64, 101, 112 Trajano, E., 647, 656 Tranquillini, W., 57, 118 Trappe, J., 210, 211, 219 Trappe, J.M., 522, 524, 529, 537, 538–540, 543, 591, 604 Traxler, R.W., 89, 90, 100 Traynor, M.M., 715, 719 Tr´ehen, P., 69, 118 Tremblay, J., 68, 112 Tremmel, D.C., 311, 328, 619, 631 Tr´emont, R.M., 344, 361, 418, 452 Trenchard, M.H., 196, 221 Trent, J., 555, 568 Trepl, L., 398, 403, 404, 411 Trexler, J.C., 683, 687 Trimble, S.W., 124, 135 Tripathi, R.S., 188, 202, 206, 215 Triplehorn, C.A., 645, 652 Tritton, L.M., 545, 547, 552, 554, 566–568 Trivelpiece, S.G., 91, 101 Trivelpiece, W., 70, 91, 101, 118 Trlica Jr, M.J., 298, 306 Troeh, F.R., 125–127, 129, 135 Trofymow, J.A., 530, 534, 541 Trojan, P., 408, 411 Troll, C., 18, 36, 39, 118 Trollope, W., 291, 306, 456, 465 Trout, T.J., 507, 513, 518, 519 Trudell, J., 55, 120 Truelove, B., 523, 530, 540 Truett, J.C., 78, 118 Trumble, J.T., 261, 269 Trumbull, V.L., 386–390, 396 Tscharntke, T., 259, 267 Tsering, T., 56, 58, 60, 65, 118 Tsitas, S.R., 65, 118 Tsuda, S., 167, 186 Tsuyuzaki, S., 69, 118, 141, 151, 152, 160 Tucker, C.J., 312, 324, 329 Tueller, P.T., 557, 569 Tufty, B., 194, 195, 221 Tuholske, J., 697, 698, 705 Tulloch, R.N., 126, 133 Tuma, C.A., 213, 222

815 Tunison, T., 423, 425, 439, 440, 442, 443, 448, 451 Tunstall, B., 392, 396 Tunstall, B.R., 594, 609 Tuomi, J., 84, 97, 119 Turchin, P., 259, 269 Turk, J., 67, 121 Turkington, R., 614, 616–620, 622, 625, 626, 627, 630, 631 Turmanina, V.I., 17, 28, 36 Turnau, K., 524, 543 Turnbull, J.W., 467, 479, 485 Turner, B.L., 673, 687, 717, 721 Turner, C.L., 293, 295, 306 Turner, I., 467, 471, 486 Turner, J., 462, 466 Turner, M.G., 337, 342, 347, 363, 708, 716, 718, 721, 722, 764, 767 Turner, R.E., 353, 354, 363 Turner, R.G., 374, 382 Turner, R.M., 310–312, 318, 319, 324, 326, 327, 329, 415, 416, 446, 599, 608 Turton, S., 476, 486 Turton, S.M., 263, 267 T¨uxen, R., 408, 411 Tvorogov, V.A., 90, 119 Twilley, R.R., 354, 363 Tyler, C.M., 278, 283, 423, 425, 426, 445, 447, 452 Tyser, R.W., 415, 416, 421, 452 Tyurina, M.M., 43, 119

U.S. Department of Agriculture, 262, 269 Uemura, S., 167, 186 Uggla, E., 164, 166, 168, 177, 186 Ugolini, F.C., 26, 31, 33, 36, 41, 53, 119, 151, 160, 600, 601, 608 Uhl, C., 223, 225, 227–231, 233, 235, 240, 242, 244, 251, 467, 472, 473, 476–478, 484–486, 614, 631 Ullmann, I., 399, 411 Ulrich, B., 453, 466 Umbanhowar Jr, C.E., 347, 350, 363 Underwood, A.J., 624, 631 Underwood, E.J., 646, 656 Ungar, I.A., 349, 361 United Kingdom Department of the Environment, 374, 384 United Nations Conference on Environment and Development, 480, 486 United States Bureau of Land Management, 388, 396 United States Bureau of the Census, 734, 736, 746 United States Department of Agriculture, 125, 128, 131, 135, 444, 452, 757, 767

United States Department of Agriculture Forest Service, 453, 460, 466 United States Forest Service, 123, 135 United States National Research Council, 367, 384 United States Salinity Laboratory Staff, 507, 519 Uno, G.E., 594, 608 Unterholzner, R., 43, 113 Upper, C.D., 528, 539 Urban, D.L., 712, 722 Urbanek, R.P., 593, 608 Urbanska, K.M., 348, 363 Uren, N.C., 506, 519 Uresk, D.W., 423, 452 Urquhart, U.H., 78, 96 Usachev, V.L., 532, 541 Usher, M.B., 371, 372, 384, 535, 543, 762, 767 Usher, P.J., 56, 119 Utomo, W.H., 507, 519

Vaitkus, M.R., 190, 195, 221 Vald´es, J., 279, 283 Valdes-Cogliano, S., 131, 134 Valdivia, S., 479, 486 Valett, H.M., 682, 684, 715, 722 Valiela, I., 336, 338, 363, 646, 656 Valiente-Banuet, A., 599, 609 Valinger, E., 200, 222 Valone, T.J., 319, 327 van Aarde, R.J., 69, 119, 379, 380, 383, 384 van Amburgh, G.L., 640, 641, 656 van Andel, J., 223, 224, 252, 497, 502 van Assche, J.A., 524, 539 van Auken, O.W., 598, 609 van Buuren, M., 682, 687 Van Cleve, K., 52, 79, 98, 119, 533, 542, 543, 577, 583, 594, 600, 608, 609, 615, 618, 631 van den Bergh, J.P., 223, 224, 252 van der Laan, D., 589, 609 van der Maarel, E., 153, 154, 159, 160, 161, 175, 186, 488, 501, 585, 593, 605, 609, 611, 629 van der Meer, P.J., 224, 231, 232, 252, 471, 486 Van der Pluijm, A.M., 340, 352, 358 van der Valk, A.G., 243, 252, 332, 333, 336, 340, 348, 352, 363 van der Zee, S.E.A.T.M., 661, 671 van Dyk, P.J., 379, 380, 384 Van Ek, R., 715, 720 Van Epps, G.A., 536, 542 van Hensbergen, H.J., 272, 281, 282, 285 Van Hook, R.I., 255, 268

816 van Hulst, R., 611, 617, 619–621, 630, 631 van Kekem, A.J., 478, 485 van Klinken, R.D., 88, 119 van Kraagenoord, C.W.S., 441, 452 Van Lear, D.H., 457, 465, 545, 547, 552, 554, 567, 568 van Meurs, L.H., 366, 384 Van Miegroet, H., 547, 551, 552, 554, 567, 569 van Pelt, R., 198, 221 Van Poolen, H.W., 578, 582 van Rensburg, P.J.J., 69, 119 Van Rhee, J.A., 131, 135 van Rompaey, R.S.A.R., 225–229, 231–234, 238, 242, 249, 250 Van Ruymbeke, M., 510, 517 van Tooren, B.F., 589, 596, 599, 607 Van Veen, J.A., 525, 541 Van Vliet-Lano¨e, B., 23, 34, 83, 102 Van Vuren, D., 443, 452 van Wagner, C.E., 164, 166, 167, 186 van Wijk, W.R., 508, 519 van Wilgen, B.W., 1, 5, 15, 272, 281, 284, 285, 435, 436, 439–442, 449, 452, 592, 603, 694, 695, 705 Van Zee, J.W., 314, 326 van Zinderen Bakker, E.M., 88, 115 Vanclay, J.K., 477, 486 Vandaele, K., 513, 516 Vandereerden, L.J., 530, 540 Vandermeer, J.H., 235, 236, 238–242, 247, 252, 488, 489, 497, 502, 636, 657 Vankat, J.L., 208, 219, 271, 272, 278, 285, 613, 616, 629 V¨are, H., 169, 170, 186 Varnamkhasti, A.S., 294, 305 Vasek, F.C., 307, 310, 314, 321, 329, 591, 592, 599, 609 Vasenev, I.I., 209, 222 Vasilevskaya, A.I., 532, 544 Vasquez, R., 229, 250 V´asquez-Yanes, C., 212, 222, 475, 483 Vass, W.P., 65, 96 Vaughan, D.G., 66, 100 Vaughan, R.W., 43, 95 Vaughn, C.E., 294, 304 Vavra, M., 417, 419, 452 Vayda, A.P., 471, 486 V´azquez de Aldana, B.R., 294, 305 Veblen, A.T., 147, 160 Veblen, T.T., 17, 27, 31, 36, 147, 160, 169, 171, 186, 187–189, 193, 196, 198, 199, 204, 206–208, 211, 217, 220–222, 253, 259, 260, 267, 269, 349, 350, 363, 455, 466, 585, 591, 605, 609, 611, 613, 616, 629, 630, 641, 642, 656, 693, 705, 715–717, 720, 722, 751, 767

AUTHOR INDEX Vega, A.J., 225, 252 Vega Condori, L., 469, 486 Velasco-Molina, H.A., 506, 519 Vel´azquez, A., 147, 160 Velazquez, I., 642, 643, 657 Velner, H., 683, 685 Veneklass, E., 224, 228, 242, 250 Veresoglou, D.S., 294, 305 Vergani, D.F., 89, 119 Ver´ıssimo, A., 472, 478, 486 Verkholat, V.P., 137, 140, 148, 149, 158 Vernon, P., 69, 99, 118 Verrecchia, E., 563, 569 Verrecchia, K., 563, 569 Verweij, P.A., 90, 115, 119 Verwijst, T., 172, 184, 188, 199, 218 Vesk, M., 78, 94 Vessel, M.F., 279, 285 Vestal, A.G., 591, 610 Vestal, J.R., 68, 90, 101, 591, 592, 609 Vetaas, O., 17, 30, 36 Vibe, C., 44, 119 Vicu˜na, F.O., 58, 119 Vidal, E., 478, 486 Vieira, I.C.G., 472, 477, 486, 614, 631 Viereck, L.A., 29, 37, 52, 75, 83, 90, 106, 113, 119, 167, 168, 186, 577, 583, 594, 599, 600, 609, 615, 618, 631 Viktorov, A.G., 532, 541 Vilarino, A., 525, 539 Viles, H., 332, 334, 337, 343, 363 Villalba, R., 25, 37, 253, 260, 269 Villase˜nor, R., 279, 283 Villeneuve, M., 614, 616, 622, 624, 628 Vimmerstedt, J.P., 600, 606 Vincent, W.F., 42, 43, 45, 93, 104, 105, 115, 119 Vint, M.K., 317, 328 Vinton, M.A., 579, 583 Vinzant, B., 131, 134 Virginia, R.A., 311, 322, 328, 522, 536, 537, 539, 540, 543 Virtala, M., 62, 92, 119 Visser, J.H., 259, 269 Visser, P.C., 47, 119 Visser, S., 522, 543 Vitebsky, P., 57, 62–64, 119 Vitousek, P.M., 1, 2, 16, 70, 73, 119, 140, 142, 147, 149, 153, 154, 157, 159, 160, 243, 244, 252, 370, 384, 413, 423, 425, 426, 435, 439, 440, 442, 443, 447, 448, 452, 471, 477, 486, 487, 502, 525, 527, 531, 533, 535, 542, 543, 546, 547, 549, 550, 553, 554, 569, 586, 587, 592, 594, 595, 599, 603, 608, 609, 611, 615, 628, 631, 673, 681, 684, 687, 694, 705, 723, 746 Vlamis, J., 575, 583

Vlassak, K., 523, 543 Voellmy, A., 51, 119 Vogel, H., 56, 119 Vogel, M., 69, 119 Vogl, R.J., 288, 291, 293, 301, 306, 307, 329, 336, 363, 574, 583 Vogt, D., 524, 543 Vogt, D.J., 211, 212, 222 Vogt, K.A., 198, 211, 212, 221, 222, 524, 543 Vogt, M., 379, 380, 384 Voigt, G., 524, 539, 543 Volkman, N.J., 70, 118 Volney, W.J.A., 169, 186 Volodina, E.R., 17, 28, 36 von St¨ulpnagel, A., 399, 401, 410, 411 von Willert, D.J., 311, 329 Vooren, A.P., 224, 231–234, 252 Voorhees, W.B., 510, 519 Voorhies, M.R., 645, 656 Vos, C.C., 681, 687 Vos, S., 276, 285 Vought, L.B.-M., 682, 687 Vourlitis, G., 66, 87, 89, 111 Vroom, M.J., 277, 284 Vtorov, I.P., 438, 442, 452 Vulinck, T.J.G., 510, 517 Waaland, M.C., 533, 543 Wace, N.M., 64, 69, 105, 119 Wackernagel, M., 1, 16 Wade, G.L., 681, 687 Wade, M., 758, 766 Wadge, G., 139, 160 Wadsworth, F.H., 229–232, 252, 469, 478, 481, 486 Waelbroeck, C., 89, 119 Wagener, S., 27, 34 Wagenet, R.J., 530, 540 Waggoner, P.E., 509, 519 Wagner, F.H., 311, 327 Wagner, J., 43, 108 Wagner, M.R., 257, 262, 267, 268 Wagner, W.L., 378, 384 Wagstaff, F.J., 320, 329, 417, 418, 452 Waheed Khan, M.A., 233, 251 Waide, J.B., 4, 16, 262, 269, 552, 554, 569, 597, 607, 712, 721, 725, 746 Waide, R.B., 2, 16, 211, 222, 224, 235, 236, 238, 252, 471, 486, 633, 635, 636, 651, 656, 657, 713, 722, 752, 763, 767 Wainwright, J., 313, 328 Waite, B., 474, 483 Waldren, R.P., 126, 135 Wali, M.K., 377, 378, 383, 384, 588, 591, 599, 604, 606, 680, 686, 760, 767 Walker, B., 310, 329, 675, 678, 687 Walker, B.H., 287, 306, 417, 419, 452

AUTHOR INDEX Walker, D.A., 62, 65, 68, 107, 116, 119, 589, 603 Walker, G.P.L., 588, 590, 609 Walker, J., 57, 119, 333, 363, 594, 609 Walker, K.F., 432, 451 Walker, L.R., 2, 15, 16, 17, 18, 33, 124, 135, 139, 153, 154, 157, 160, 169, 170, 186, 199, 211, 215, 222, 224, 235, 238–245, 246–248, 252, 370, 384, 471, 476, 486, 487, 502, 585, 586, 588, 590–602, 604, 605, 609, 611, 613–615, 626, 628, 631, 635, 650, 651, 656, 657, 713, 714, 717, 722, 752–754, 763, 766, 767 Walker, M.D., 589, 603 Walker, P.H., 515, 519 Walker, R., 467, 485 Walker, S., 472, 483 Walker, T.W., 594, 609 Wall, C.J., 530, 543 Walla, M.D., 65, 96 Wallace, A., 562, 568, 611, 629 Wallace, J.B., 427, 450 Wallace, J.R., 594, 603 Wallace, L., 287, 305 Wallace, L.L., 713, 719, 720 Wallace, M.M.H., 639, 656 Wallach, L.S., 574, 583 Waller, D.M., 596, 599, 607 Wallin, D.O., 464, 466 Walsh, J.E., 66, 98, 334, 337, 363 Walsh, J.S., 611, 629 Walsh, M.E., 390, 391, 395 Walsh, R.P.D., 471, 475, 486 Walstad, J.D., 461, 466 Walter, H., 43, 120, 130, 135, 163, 167, 186, 287, 303, 306, 309, 312, 317, 329 Walters, C., 683, 687, 740, 746 Walters, M., 271, 272, 284 Walters, N.M., 39, 109 Walton, D.W.H., 17, 37, 68, 69, 108, 120, 591, 599, 600, 607, 609, 611, 630, 713, 721 Wang, D., 554, 566 Wang, G.M., 578, 582 Wang, Q.J., 84, 120 Wang, Z.J., 471, 486 Wang, Z.M., 60, 122 Wanless, H.R., 233, 235, 244, 245, 251, 333, 334, 336, 347, 348, 352, 355, 362 Warburg, M., 276, 285 Warburg, O., 508, 519 Ward, C.J., 333, 362 Ward, L.K., 130, 135 Ward, R.T., 47, 105 Ward, S.C., 380, 381, 382, 674, 687 Ward, T.J., 313, 326, 389, 395 Wardle, P., 17, 37

817 Waren, S.D., 563, 564, 566 Waring, M.R., 386, 388, 389, 395 Warming, E., 595, 609 Warneke, A., 130, 134 Warner, A., 528, 543 Warner, N.J., 528, 529, 543 Warren, A., 309, 314, 316, 323, 325 Warren, S.D., 386–390, 396, 561, 569, 638–641, 656 Warren Wilson, J., 50–52, 120 Wassen, M.J., 682, 684 Wassenaar, T.D., 379, 380, 384 Wasserman, S.S., 259, 267 Watanabe, R., 188, 195, 222 Waterhouse, D.F., 646, 656 Waterhouse, J.C., 678, 685 Waters, J.W., 65, 115 Waters, M.R., 56, 120 Watkin, B.R., 646, 654 Watkins, B.P., 53, 69, 115, 120 Watson, D.G., 64, 114 Watt, A.S., 191, 203, 222, 595, 609, 646, 656, 708, 722 Watt, F., 243, 244, 251, 752, 766 Watzin, M.C., 337, 357 Weatherly, N.S., 548, 554, 568 Weaver, H., 574, 583 Weaver, J.E., 574, 583, 645, 656, 714, 722 Weaver, P.L., 239–242, 252 Webb, B.L., 536, 541, 543, 563, 564, 567, 569 Webb, D., 524, 541 Webb, J.W., 255, 257, 262, 269, 337, 343, 346, 347, 353, 361, 363 Webb, K.L., 353, 362 Webb, L.J., 230, 234, 241, 252 Webb, N.R., 532, 543 Webb, R.H., 310, 316, 319, 324, 326, 329, 388, 396, 589, 609 Webb, S.L., 189, 191, 192, 196, 197, 199, 201, 203, 204, 206–211, 213, 222 Webb, T., 493, 501 Webb, W.L., 262, 269 Webber, B., 545, 547, 549, 569 Webber, P.J., 62, 68, 71, 74, 107, 119, 120 Weber, J.C., 199, 221 Weber, M.G., 90, 120, 164, 186, 577, 582 Weber, W.A., 90, 116 Wedin, D., 622, 624, 631 Wedin, D.A., 292, 293, 306 Weed, R., 51, 102 Weeda, W.C., 646, 656 Weeks Jr, H.P., 390, 395 Weesies, G.A., 128, 134 Weetman, G.L., 545, 547, 549, 569 Weidensaul, T.C., 524, 541 Weigmann, G., 403, 411 Weimerskirch, H., 59, 70, 106, 120

Wein, G., 415, 416, 421–423, 444, 452 Wein, R.W., 52, 62, 68, 78, 82, 90, 96, 120, 161–164, 168, 174, 183, 184, 186, 335, 336, 346, 348, 351, 359, 360, 415, 416, 421–423, 444, 452, 577, 582, 583 Weinbaum, B.S., 524, 527, 537, 538, 543 Weinberg, P., 697, 705 Weiner, J., 625, 631 Weinstein, L.H., 257, 258, 267, 525, 540 Weis, I.M., 311, 329, 615, 617, 630 Weiss, K., 456, 465 Weisser, P.J., 379, 382 Welch, D.M., 22, 37 Welden, C.W., 242, 252 Welker, J.M., 89, 121 Weller, M.W., 70, 110, 683, 687 Wellington, A.B., 613, 616, 631 Wellington, E.M.H., 376, 384 Wells, C.G., 550, 569 Wells, J.D., 366, 384 Wells, P.G., 68, 90, 113, 120 Wells, R.D.S., 433, 452 Wells, S., 713, 719 Wells, S.G., 459, 466 Welty, J.C., 645, 656 Wen, D., 123, 135 Wendroth, O., 507, 516 Wendt, R.C., 126, 135 Wenzel, G., 56, 119 Werger, M.J.A., 311, 318, 323, 329 Werner, M., 378, 382 Werner, P., 59, 120 Werner, P.A., 443, 447, 613, 616, 619, 622, 624, 625, 629–631 Weslien, J., 173, 182 Wesseling, J., 508, 519 West, B.A., 128, 133 West, D.C., 229, 230, 251, 437, 451, 545, 547, 552, 554, 568 West, G.C., 57, 120 West, N.E., 309, 311, 329, 417–419, 423, 450, 451, 530, 533, 538, 555–558, 562–564, 566, 568, 569 Westhoff, V., 408, 411 Westman, W.E., 40, 70, 120, 678, 687 Westoby, M., 310, 329, 675, 678, 687 Wetherald, R.F., 66, 109 Wetherald, R.T., 67, 110 Weyer, D., 472, 484 Whalen, S.C., 89, 120 Wharton, R.A., 42, 99 Wharton Jr, R.A., 41, 45, 112, 120 Whelan, C.J., 213, 222, 681, 684, 759, 765 Whelan, R.J., 572, 574, 575, 583, 593, 609 Whetton, P., 66, 91, 94 Whicker, A.D., 295, 306, 580, 583 Whigham, D., 331, 363

818 Whigham, D.F., 226, 228, 232–235, 239, 241, 244–246, 248, 252 Whipps, J.M., 525, 530, 539, 541, 543 Whisenant, S.G., 320, 329, 417, 418, 423, 425, 439, 440, 452, 681, 687 Whitaker, I., 92, 120 Whitaker Jr, J.O., 646, 647, 656 Whitaker, M., 125, 126, 134, 691, 701, 705 Whitby, L.M., 369, 383 White, A.S., 189, 197, 198, 206, 218 White, C.S., 401, 411 White, D., 132, 134 White, D.C., 42, 112 White, E.H., 547, 569 White, G.F., 196, 218 White, H., 472, 484 White, M.G., 59, 120 White, O., 129, 134 White, P.S., V, VI , 2, 16, 161, 185, 187, 189–191, 193, 202, 205, 206, 214, 215, 218, 220, 222, 223, 224, 234, 243, 247, 250, 252, 307, 329, 333, 334, 339, 340, 349, 355, 361, 363, 386, 389, 395, 396, 397, 412, 414, 452, 454, 466, 467, 485, 488, 493, 502, 585, 590, 595, 597, 607, 609, 621, 631, 633, 635, 642, 655, 656, 680, 686, 707, 709, 716, 719–721, 750, 764, 766, 767 White, R.G., 54, 55, 84, 95, 96, 120 White, W.D., 459, 466 Whiteaker, L.D., 153, 154, 160, 419, 420, 439, 441, 447, 599, 609 Whitford, W.G., 312, 314, 318, 322, 326–329, 522, 526, 536, 540, 544, 645, 654 Whitham, T.G., 579, 582 Whitman, A.A., 478, 480, 486 Whitman, W.C., 557, 569 Whitmore, J.L., 469, 478–482, 485, 486 Whitmore, M.C., 224, 233, 249 Whitmore, T.C., 223, 225–229, 233, 234, 236, 237, 240–242, 244, 245, 247, 251, 252 Whitney, G., 454, 456, 459, 460, 466 Whitney, G.G., 189, 193, 194, 199, 200, 208, 222 Whittaker, R.H., 276, 284, 408, 412, 487, 497, 502, 595, 599, 607, 609 Whittaker, R.J., 2, 16, 17, 26, 28, 29, 35, 37, 142, 145, 147–149, 154, 157, 159, 160, 591, 593, 599, 609 Whittingham, W.F., 532, 544 Whitworth, W.R., 386, 396 Whyte, A.V., 661, 671 Wickham, T.H., 512, 519 Wicklow, D.T., 532, 544, 645, 655 Wickman, B.E., 261, 269

AUTHOR INDEX Wiebe, W.J., 39, 105 Wiedemann, A.M., 441, 452 Wielgolaski, F.E., 39, 43, 55, 59, 62, 68, 69, 76, 79, 112, 120 Wienand, K.T., 281, 285 Wiencke, C., 65, 108 Wiens, J.A., 292, 298, 306 Wight, H.M., 211, 216 Wigley, T.M.L., 66, 106, 113 Wijermans, M.P., 277, 284 Wikars, L.O., 166, 186 Wiklander, G., 548, 549, 554, 569 Wilbur, R.B., 353, 363, 459, 466 Wilcher, J.L., 388, 390, 395 Wilcox, D.A., 348, 363, 433, 452 Wilczynski, C.J., 212, 222 Wild, H., 372, 384 Wiles, G.J., 636, 656 Wiley, G.M., 646, 652 Wiley, T.R., 635, 638, 656 Wilkie, D.S., 473, 486 Wilkins, C.W., 762, 766 Wilkins, D., 524, 544 Wilkinson, J., 314, 327 Will, T., 635, 656 Willard, B.E., 69, 79, 110, 120, 121 Willatt, S.T., 561, 569 Willems, J.H., 371, 384 Willemsen, R.M., 618, 630 Williams, A., 369, 384 Williams, A.J., 53–55, 97, 103, 116, 121 Williams, C.E., 425, 428, 452 Williams, D.L., 66, 111, 121 Williams, E.J., 64, 112 Williams, H.F.L., 337, 347, 363 Williams, J.E., 432, 450 Williams, K.J., 289, 290, 295, 305 Williams, M., 453, 456, 460, 465, 466, 675, 687 Williams, M.W., 52, 67, 99, 121 Williams, O.B., 317, 324, 329 Williams, P.A., 90, 121 Williams, P.H., 560, 567 Williams, P.J., 47, 121 Williams, P.L., 52, 121 Williams, R.J., 644, 656 Williams Jr, R.S., 20, 36 Williams, S.E., 713, 719 Williams, S.T., 376, 384 Williams, T.M., 190, 195, 199–201, 217, 454, 465 Williams-Linera, G., 472, 486, 715, 722 Williamson, G.B., 60, 121, 189, 202, 205, 222 Williamson, J.W., 59, 108 Williamson, M., 324, 329, 487, 502 Williamson, M.H., 488, 499, 502 Williamson, N.A., 368, 369, 371, 384

Willig, M.R., 2, 5, 16, 224, 243, 245, 252, 257, 269, 471, 476, 486, 635–638, 643, 650, 651, 652, 653, 655–657, 752, 754, 756, 763, 766, 767 Willis, A.J., 323, 330 Willis, K.J., 175, 186 Willms, W.D., 556, 566 Wills, J.B., 702, 705 Willson, M.F., 213, 222 Wilsey, B.J., 337, 363 Wilshire, H.G., 316, 324, 328, 329, 388, 396, 589, 609 Wilson, A.D., 614, 616, 618, 631 Wilson, A.T., 52, 121 Wilson, B.F., 198, 208, 218, 222 Wilson, D.E., 647, 656 Wilson, E., 668, 671 Wilson, E.O., 729, 746 Wilson, J.B., 17, 28, 36, 78, 114, 589, 608, 613, 615, 616, 625, 627, 629, 681, 684 Wilson, K.-J., 62, 70, 91, 121 Wilson, M., 42, 66, 94 Wilson, M.V., 139, 158 Wilson, P.R., 55, 91, 118 Wilson, R.P., 70, 121 Wilson, S.D., 351, 363, 387, 388, 390, 396, 614, 617, 619–621, 623, 624, 629–632, 681, 687 Windsor, D.M., 255, 267, 475, 484 Wink, M., 493, 500, 502 Wink, R.L., 298, 306 Winkler, D.S., 90, 101 Winkler, S., 20, 37 Winn, A.A., 613, 630 Winn, D.T., 557, 566 Winogradsky, S.N., 523, 544 Wint, G.R.W., 255, 269 Wint, S.H., 471, 486 Winterringer, G.S., 591, 610 Winters, J.K., 337, 347, 363 Winward, A.F., 557, 566 Wirawan, N., 224, 226, 229, 234, 249, 473, 484 Wischmeier, W.H., 513, 519 Wiser, S., 590, 610 Wishart, D.M., 62, 121 Wisheu, I.C., 340, 342, 363, 617, 619, 620, 630 Wissmar, R., 427, 450 Witcosky, J.J., 258, 269 Witham, J.W., 189, 197, 198, 206, 218 Witherick, M.E., 337, 361 Witkowski, E.T.F., 594, 610 Witter, J.A., 171, 175, 184 Wittig, R., 397, 411, 412 Wittman, P.K., 256, 268 Woakes, A.J., 62, 70, 100

AUTHOR INDEX Woerheide, J.D., 202, 205, 219 Woese, C.R., 522, 544 Wojterski, T.W., 436, 441, 452 Wolda, H., 260, 269 Wolden, L.G., 428, 431, 451 Wolf, L.L., 488, 502 Wolf, P.L., 337, 346, 362 Wolf, S.A., 737, 746 Wolfe, A., 66, 112 Wolfe, R.W., 680, 686 Wolff, J.O., 593, 610 Wolfram, C., 64, 96 Wong, H.H., 279, 285 Wong, K.K.Y., 531, 544 Wong, M., 260, 269 Woo, M.-K., 43, 66, 106 Wood, D.M., 52, 121, 140–144, 148, 149, 151–153, 158–160, 591, 592, 594, 599, 610, 613, 629 Wood, J.T., 681, 685 Wood, L.S., 512, 518 Wood, M., 130, 135 Wood, T.G., 318, 328, 474, 483, 646, 654 Wood, T.W.W., 241, 252 Woodhouse Jr, W.W., 646, 655 Woodley, S.J., 577, 582 Woodmansee, R.G., 288, 294, 306, 435, 451, 574, 583, 711, 721 Woodroffe, C.D., 334, 363 Woodruff, N.P., 513, 519 Woods, P., 229, 234, 252, 473, 482, 486 Woodward, F.I., 89, 110, 169, 183, 321, 329 Woodwell, G.M., 408, 412 Wookey, C.W., 67, 111 Wookey, P.A., 89, 121 Woolbright, L.L., 636, 638, 656, 657, 752, 767 Wooldridge, G., 187, 222, 228, 252 Wooldridge, G.E., 528, 530, 538 Woolf, S.W., 701, 704 Worcester, B.K., 515, 517, 518 Worf, T.P., 508, 517 Worland, M.R., 45, 121 World Commission on Environment and Development, 660, 671, 701, 705 World Development Forum, 724, 746 World Meteorological Organization, 314, 315, 329 World Resources Institute, 125, 128, 135, 467, 470, 486, 729, 734, 736, 746 Worley, I.A., 17, 26, 36, 37, 591, 600, 608, 610 Wormald, T.J., 479, 480, 486 Worrall, J.J., 189, 197, 202, 222 Worster, D., 668, 671 Worster, D.E., 2, 16 Wotton, B.M., 164, 166, 175, 183

819 Wright, C., 472, 484 Wright, D.H., 129, 130, 135 Wright, H., 56, 117 Wright, H.A., 298, 306, 455, 459, 466, 564, 569, 572–576, 583, 713, 719 Wright, H.E., 193, 194, 222 Wright Jr, H.E., 20, 28, 37 Wright, R.G., 140, 158, 599, 604 Wright, S.J., 260, 269 Wu, D., 366, 383 Wu, J., 707, 709, 710, 722 Wu, L., 403, 412 Wu, L.S.Y., 555, 556, 566 Wu, X.D., 59, 108 Wunderle Jr, J.M., 241, 242, 246, 467, 479, 485, 635, 636, 638, 642, 643, 656, 657 Wuttunee, W., 58, 114 Wyatt-Smith, J., 239, 241, 252 Wykes, B.J., 381, 383, 674, 684 Wynn-Williams, D.D., 17, 26, 27, 37, 591, 592, 610 Wynne-Edwards, W.C., 43, 95 Xia, W.P., 56, 60, 82, 121 Yadav, J.S.P., 233, 251 Yair, A., 312, 313, 329, 563, 569 Yamamoto, K., 141, 151, 159 Yamamoto, S.-I., 188, 195, 204, 206, 208, 211, 222 Yamashita, T., 318, 328 Yang, D., 67, 121 Yang, F.T., 84, 120 Yarranton, G.A., 591, 610, 616, 617, 632 Yarranton, M., 616, 617, 632 Yarrington, G.A., 151, 152, 160 Yates, C.J., 676, 677, 687 Yates, H.S., 587, 603 Yavitt, J.B., 225, 226, 230, 252, 455, 465 Yeaton, R.I., 311, 320, 329 Yetter, T.C., 190, 203, 204, 206, 221, 224, 251 Yih, K., 235, 238–242, 252, 636, 657 Yiou, F., 590, 603 Yoda, S., 188, 195, 222 Yonker, C.M., 292, 304 Yool, S.R., 324, 330 You, C., 239, 252 Young, A., 126, 127, 135, 514, 519 Young, E.C., 70, 121 Young, J.A., 423, 439, 452 Young, J.Y., 557, 569 Young, K.R., 480, 486, 589, 610 Young, O.R., 57, 112 Young, R.A., 123, 124, 129, 133 Young, S.B., 52, 121 Young, S.S., 471, 486

Young, T.P., 232, 252 Younghans-Haug, C., 65, 114 Younkin, W.E., 79, 121 Yu, X.G., 59, 121 Yudelman, M., 662, 670 Yung, Y.L., 65, 118

Zaady, E., 295, 306 Zach, R., 64, 66, 116 Zachara, J., 151, 160 Zacharias, T.P., 131, 135 Zachvatkin, A.A., 528, 544 Zackrisson, O., 166, 168, 172, 173, 185, 186 Zagt, R.J., 478, 485 Zahran, M.A., 323, 330 Zaidi, B.R., 530, 544 Zak, J.C., 526, 532, 544, 645, 647, 648, 653, 657, 756, 767 Zakharov, V.G., 66, 121 Zamora, B., 644, 654 Zamora, N., 235, 238–242, 247, 252, 636, 657 Zangerl, A.R., 626, 632 Zarin, D.J., 124, 135, 590, 592, 609, 717, 722 Zarling, J.P., 66, 106 Zasada, J.C., 169, 170, 186, 592, 593, 597, 598, 601, 602, 609, 610 Zasoski, R., 457, 465 Zasoski, R.J., 17, 27, 29, 35, 535, 539 Zavitkovski, J., 462, 466 Zebryk, T., 189, 193, 217 Zech, W., 480, 485 Zedler, J.B., 340, 343, 345, 347, 349, 357, 363 Zedler, P.H., 320, 330, 423, 425, 439, 452 Zellmer, I.D., 579, 583 Zhang, B.C., 82, 121 Zhang, H.O., 510, 511, 519 Zhang, J., 555, 569 Zhang, J.X., 60, 122 Zhang, S.L., 84, 92, 121 Zhang, X.W., 352, 362 Zhao, B.L., 60, 122 Zhao, D.H., 92, 121 Zhao, X.Q., 92, 113, 121 Zharkova, Yu.G., 68, 100 Zhdanova, N.N., 532, 544 Zhen, R.D., 82, 121 Zhou, X.M., 56, 84, 92, 109, 121 Zhu, H., 366, 383 Zhu, W., 190, 212, 216 Ziegler, P., 43, 115 Zielinski, G., 66, 112 Ziemann, P.J., 41, 115 Zimmerer, K.S., 59, 121

820 Zimmerman, J.K., 2, 16, 211, 222, 224, 235, 236, 238, 241, 243, 245, 252, 471, 476, 477, 482, 486, 635, 650, 651, 656, 657, 752, 754, 763, 767 Zimmerman, R.J., 343, 363 Zink, T.A., 415, 416, 444, 452, 536, 544 Zipper, C.E., 377, 378, 382 Zisheng, Q., 188, 207, 221

AUTHOR INDEX Zlattner, L., 530, 544 Zobel, D.B., 140, 142, 144, 148, 157, 160, 587, 610 Zobel, K., 616, 618, 632 Zobel, M., 616, 618, 632 Zoller, W.H., 123, 134 Zollitsch, B., 28, 30, 37 Z¨ottl, H., 73, 121

Zou, X., 471, 486 Zucca, C.P., 471, 486 Zugasty Towle, P., 226, 234, 252 Zuo, K.C., 60, 122 Zurek, J., 661, 671 Zutter, B.R., 618, 628 Zwally, H.J., 66, 67, 98, 122 Zwerman, P.J., 127, 133

SYSTEMATIC INDEX Abies spp. (ﬁrs), 162, 163, 170, 188–190, 200, 206, 211, 261 A. amabilis (Dougl. ex Loud) Dougl. ex Forbes, 29, 204, 211, 454 A. balsamea (L.) Mill. (balsam ﬁr), 163, 167, 168, 170, 172, 175, 201, 207, 209, 455, 548, 552, 580, 750 A. concolor (Gord. & Glend.) Lindl. ex Hildebr. (white ﬁr), 261 A. faxoniana Rehd. & Wils., 207 A. fraseri (Pursh) Poir., 455 A. lasiocarpa (Hook.) Nutt., 204, 458 A. mariesii Mast., 204 A. sibirica Ledeb. (Siberian ﬁr), 163, 167, 168 Abronia maritima Nutt. ex S. Watson (sand verbena), 532 Acacia, 170, 274, 425 A. bonariensis Gill. ex Hook. et Arn. (˜napinday), 498 A. caven (Molina) Molina, 279 A. cyclops G. Don, 436, 439 A. karoo Hayne, 379, 380 A. longifolia (Andr.) Willd., 436, 439 A. mearnsii De. Wild., 430, 436, 441 A. saligna (Labill) H. Wendl, 436, 439–441 A. senegal (L.) Willd., 665 A. seyal Delile, 665 Acanthopanax sciadophylloides Franch. & Sav., 207 Acari, 529, 647 Acer, 192, 205 A. japonicum c. P. Thunberg ex A. Murray, 207 A. mono Maxim., 207 A. pensylvanicum L., 206, 208 A. rubrum L. (red maple), 206, 210, 212, 350, 552 A. saccharinum L., 192, 552 A. saccharum Marsh. (sugar maple), 201, 203–210, 212–214, 378, 403, 548, 622, 623 A. spicatum Lamb., 178 Aceras anthropophorum (L.) Ait. f., 371 Acomastylis rossii ssp. turbinata (Rydb.) W.A. Weber, 86 Aconitum septentrionale Koelle., 180 Acrididae, 639 Adelges tsugae Annand, 260 Adenostoma fasciculatum Hook. & Arn. (chaparral), 573 Adesma bicolor (Poir.) DC., 491 Aegopodium, 398 Aextoxicon punctatum Ruiz & Pav., 204 Agamemnon iphimedeia Moxey (walkingstick), 257, 637 Agave deserti Engelm., 311 Agropyron, 694 A. cristatum (L.) Gaertn., 623, 624 A. desertorum (Fischer) Schultes, 420, 442 A. elongatum (Host.) Beauv. (grama alargada), 497 A. spicatum (Pursh) Scrib. & Smith, 141, 442 Agrostis, 374 A. capillaris L., 374, 375

A. gigantea Roth, 376 A. montevidensis Doell, 491 A. stolonifera L., 89 Ailanthus altissima (Mill.) Swingle (tree of heaven), 405, 614 Ajania tenuifolia (J. Jacq.) Tzvelev, 82 Alauda arvensis L. (ﬁeld lark), 398 Alcadia striata Lamarck, 638 Alces alces L. ( = A. americanus) (moose), 169, 170, 529, 579, 593, 750 Alligator, 332, 751 A. mississippiensis (Daudin) (American alligator), 340 Allophylus natalensis De Winter, 380 Alnus, 26, 622 A. ﬁrma S. & Z., 142, 143 A. fruticosa Rupr., 168 A. incana (L.) Moench, 180 A. kamtschatica (Regel) Kom. ( = A. fruticosa ssp. kamtschatica), 146 A. rubra Bong. (red alder), 145, 152, 153, 551–554, 622 A. sinuata (Reg.) Rydb., 29, 600 A. tenuifolia Nutt. ( = A. incana ssp. tenuifolia), 533 Aloysia gratissima Gill. & Hook. (azahar del campo), 498 Alsophila pometaria Harris (cankerworm), 259, 552 Alternaria, 532 Amaranthus quitensis H.B.K. (yuyo colorado), 494 Ambrosia, 310, 321 A. artemisiifolia L., 613, 617, 618 A. dumosa (Gray) Payne, 324 A. triﬁda L., 621 Ammophila, 528 A. arenaria (L.) Link, 153, 278, 436, 441 Amphibia, 404 Amphipoda, 332 Amphispiza belli (Cassin), 390 A. bilineata (Cassin), 390 Anaphalis margaritacea (L.) Benth & Hook. F. ex C.B. Clarke, 151 Anas, 390 Andropogon gerardi Vitman, 619 A. virginicus L., 439 Anemone patens L., 166 A. vernalis L., 166 Angophora, 264, 265 A. ﬂoribunda (Smith) Sweet, 264 Anoda cristata (L.) Schltdl. (malva cimarrona), 494 Anolis, 636 Anoplognathes, 264 Anser caerulescens L. (lesser snow goose), 335, 337, 348, 353, 579, 751 Anthoxanthum odoratum L., 154, 439 Anthracothorax viridis (Audebert & Vieillot), 636

821

822 Antilocapra americana Ord., 529 Antistrophus silphii Gil., 639 Apis mellifera L., 324 Aporrectodea rosea Savigny, 509 Aptenodytes forsteri Gray, 89 A. patagonicus Miller, 70 Aralia nudicaulis L., 178 Archaebacteria, 522 Arctagrostis arundinacea (Trin.) Beal, 82, 86 Arctocephalus gazella (Peters) Monatsb. K. Preuss, 54 Arctophila fulva (Trin.) Rupr. ex Anderss., 83 Arctostaphylos uva-ursi (L.) Spreng., 180 Ardisia, 149 Aristida murina Cav., 491 A. pungens Desf., 436 A. stricta Michx. (wiregrass), 333 Armeria maritima (Mill) Willd., 374 Armillaria, 522 Arphia conspersa Scudder, 649 Arrhenatherum, 406 A. elatius (L.) J. & K. Presl., 372, 407 Artemisia (sagebrush), 287, 417, 573 A. tilesii Ledeb., 86 A. tridentata Nutt., 317, 319 A. vulgaris L., 406, 407 Artibeus jamaicensis Leach, 637, 638 Aster macrophyllus L., 178 A. pilosus Willd., 617, 623 Asteraceae, 274 Astragalus nutzotinensis Rouss., 29 Athyrium ﬁlix-femina (L.) Roth, 178 Atriplex, 325 A. canescens (Pursh) Nutt., 316 A. portulacoides L., 340 Austroselenites alticola H.B. Baker, 643 Avena fatua L., 421 A. sativa L. (oats), 59, 127 Baccharis, 498 B. articulata (Lam.) Pers. (carquejilla), 491, 498 B. coridifolia DC. (m´ıo-m´ıo, romerillo), 491 B. notosergila Griseb., 491, 498 B. trimera (Less.) DC. (carqueja), 491 Balaena mysticetus L. (bowhead whale), 58 Balaenoptera acutorostrata Lac., 91 Banksia, 424, 442 Becium homblei (De Wild.) Duvign. & Plancke, 372, 374 B. obovatum E. Mey, 372 Berroa gnaphalioides (Less.) Beauv., 491 Berteroa incana (L.) DC., 406 Beta vulgaris L. (sugar beet), 130, 515 Betula, 26, 146, 166, 168, 171, 188, 192, 204, 205, 207, 530, 579 B. alleghaniensis Britton (yellow birch), 203, 205, 206, 210, 213, 552, 622 B. ermanii Cham., 146, 150 B. glandulosa Michx. ( = Betula nana), 55 B. lenta L., 205, 552 B. maximowcziana Regel in DC., 151

SYSTEMATIC INDEX B. papyrifera Marsh., 167, 170, 175, 201, 205, 206 B. pendula Roth, 180, 406 B. populifolia Marsh., 205 B. pubescens Ehrh., 20, 29, 30, 162, 166, 171, 173, 180 B. utilis D. Don, 207 Bidens subalternans DC. (amor seco), 494, 496 Bison, 751 B. bison L. (American bison), 350, 579 Blattidae, 639 Bos grunniens L. (yak), 60 B. javanicus d’Alton (banteng), 443 B. taurus L., 129, 438 Bothriochloa laguroides (DC.) Pilger., 490, 491 Bouteloua curtipendula (Michx.) Torr. (sideoats grama), 574 B. gracilis (Willd. ex Kunth) Lang ex Grifﬁths (blue grama grass), 287, 557, 648 Brachiaria, 439 Brachylaena discolor DC., 380 Brachystegia ﬂoribunda Benth., 374 Brachystola magna (Girard), 649 Brachythecium, 178 Branta bernicla L. ( = B. nigricans), 70 Brassica oleracea var. acephala DC., 130 B. tournefortii Govan., 439 Brassicaceae, 274 Briza subaristata Lam., 490, 491 Bromus, 406, 437 B. inermis Leysser, 416, 623 B. japonicus Thunb. ex Murr., 423 B. mollis L. (softchess), 421, 422 B. rubens L. (red brome), 278, 317, 418, 421–423, 439 B. tectorum L. (cheatgrass), 317, 417, 418, 420, 423, 439 B. unioloides Kunth (cebadilla criolla, cebadilla australiana), 491 Broussonetia papyrifera (L.) L’Her ex Vent (mora turca), 498 Bryum argenteum Hedwig, 87 Bubalus bubalis L., 438, 443, 681 Bucculatrix ﬂourensiae Braun, 318 Bufo bufo L. (toad), 407 Bulbostylis cupricola Goetgh., 374 Cakile, 436 Calamagrostis, 144 C. purpurea (Trin.) Trin., 180 Callorhinus ursinus L. (fur seal), 64, 90 Calluna, 575 C. vulgaris (L.) Hull, 165, 170, 180, 375, 573, 575 Caloplaca, 77 Cambaridae, 432 Camelus dromedarius L., 681 Canis familiaris dingo L. (dingo), 417 Capra hircus L., 438, 681 C. ibex L., 58 Caracolus caracolla (L.), 637, 643 Carcinus maenas L., 436 Carduelis tristis L., 642 Carex, 90, 178 C. aquatilis ssp. stans (Drej.) Hulten, 81

SYSTEMATIC INDEX C. bonariensis Desf., 491 C. subspathacea Wormsk. ex Hornem. (Hoppner’s sedge), 83, 353 Carnegiea gigantea (Englem.) Britt & Rose, 310, 317 Carpobrotus edulis (L.) Bolus, 278, 421–423, 425, 426 Carya, 377, 552 C. cordiformis (Wangenh.) K. Koch, 205 Castanea dentata (Marsh.) Borkh., 455 C. sativa P. Mill. (chestnut), 276 Castanopsis cuspidata (Thunb.) Schottky, 211 Castor (beavers), 750, 751 C. canadensis Kuhl (American beaver), 169, 170, 334, 432, 577, 681 C. ﬁber L., 169 Casuarina, 470 C. equisetifolia L., 436 Catharacta, 65 C. antarctica (Lesson), 91 Catharus ustulatus (Nutt.), 642 Ceanothus, 573, 575, 617 C. megacarpus Nutt., 617 Cecropia, 477, 758 C. obtusifolia Bertol., 477 C. schreberiana Miq., 259 Celtis, 310 C. africana Burm., 380 C. laevigata Willd., 377 C. occidentalis L., 205 Cenchrus ciliaris L., 317, 439 C. myosuroides Kunth (cadillo), 497 Centaurea diffusa Lam., 418 C. maculosa Lam., 416, 421 Cepolis squamosa F`erussac, 638, 643 Cerastium fontanum Baumg., 68 Ceratocystis ulmi (Buism.) Moreau, 262, 437, 455, 751 Ceratodon purpureus (Hedw.) Brid., 166 Ceratophyllum demersum L., 433 Cerceopithecus aethiops L., 379 Cercidium, 312 Cervus canadensis L. ( = Cervus elaphus) (elk), 150, 170, 290, 392, 529 Chaerophon jobensis (Miller), 643 Chalinolobus nigrogriseus (Gould), 643 Chamaecyparis, 434 C. thyoides (L.) B.S.P. (Atlantic red cedar), 340 Chamaenerion angustifolium (L.) Scop. ( = Epilobium angustifolium), 146 Chaptalia, 491 Chelidonium, 406 Chen caerulescens (L.) (lesser snow goose), 43 Chenopodium album L. (quinoa blanca), 494 C. botrys L., 405, 406 C. ﬁcifolium SM., 407 C. glaucum L., 407 C. rubrum L., 407 C. strictum Roth, 406 Chevreulia sarmentosa (Pers.) Blake, 491 Chimaphila umbellata (L.) Barton, 166

823 Chionis, 65 Chionochloa, 90 Chlamydomonas nivalis (Bau.) Wille (snow algae), 42 Chlorostilbon maugaeus (Audebert & Vieillot), 636, 642 Choristoneura fumiferana (Clemens) (spruce budworm), 169, 170, 580, 750 Chrysothamnus nauseosus (Pall.) Britt., 141 Chusquea, 207 Cinna latifolia (Trevir.) Griseb., 178 Circaea alpina L., 178 Cirsium arvense (L.) Scop., 407 C. helenoides (L.) Hill, 180 C. vulgare (Savi) Tenore, 154 Cladina, 166, 177 C. arbuscula (Wallr.) Burgaz, 170 C. rangiferina (L.) Nyl., 170 C. stellaris (Opiz) Brodo, 170 Cladium jamaicense Crantz, 331, 351 Cladonia, 165, 166, 173, 177 Clethrionomys gaperi Vigors, 213 Cliftonia monophylla (Lam.) Britt. ex Sarg. (titi), 355 Clintonia borealis (Ait.) Raf., 178 Coccinoidea, 639 Cochlearia pyrenaica (Babington) D.H. Dalby, 374 Cochliobolus sativus (S. Ito & Kurib.) Drechsler ex Dastur, 530 Coereba ﬂaveola L., 636, 642 Coffea arabica L., 691 Coleogyne ramosissimma Torrey, 562 Colinus virginianus L., 639 Collema curtisporum Degel., 167 C. fragrans (Sm.) Ach., 167 Collembola, 372, 523, 647 Colobanthus, 88 C. quitensis (Kunth.) Bartl., 45, 77, 88 Coloradia pandora Blake, 257 Columba livia f. domestica Gmelin, 398, 407 C. squamosa Bonnaterre, 636 Conyza, 491 C. canadensis (L.) Cronquist, 406 Copepoda, 332 Corbicula ﬂuminea (Mueller), 432 Cordia multispicata Cham., 477 Coriaria arborea Lindsay, 153 Corispermum, 406 Cornus canadensis L., 178 C. drummondi C.A. Mey., 213 C. ﬂorida L., 208, 212 C. suecica L., 180 Coronopus didymus (L.) Smith (mastuerzo), 494 Corophium volutator Pallas (amphipod), 343 Cortaderia jubata (Lem.) Stapf, 426 C. selloana (Schult.) Aschers & Graebn. (cortadera), 498 Corvus, 407 C. monedula L. (jackdaw), 407 Corylus cornuta Marsh., 178, 207 Coturnix coturnix (L.) (partridge), 398 Crassula helmsii (T. Kirk), 433, 434 Crataegus, 617

824 Crematogaster punctulata Emery, 647–649 Crepis, 497 Crocodylus, 332 Cronartium ribicola Fischer, 436 Crotalaria cobalticola Durigneaud & Plancke, 374 C. juncea L., 379 Cryphoonectria [Endothiella] parasitica (Murr.) Barr., 437 Cryptocarya alba (Molina) Looser, 279 Cryptomeria japonica (L.) D. Don, 436 Cryptostigmata, 372 Cucumis, 406 Cucurbita, 406 C. pepo (squash), 59 Cunonia capensis L., 430 Cyanocompsa cyanoides (Lafresnaye), 642 Cyathea smithii Hook f., 209 Cyathus, 528 Cygnus, 390 Cynodon, 428 C. dactylon (L.) Pers., 428, 432, 441 Cynomys (prairie dogs), 645 C. ludovicianus (Ord) (prairie dogs), 580 Cyperaceae, 52, 282 Cyphorhinus phaeocephalus Sclater, 642 Cyprinodon nevadensis Eigen. & Eigen., 324 Cyrilla racemiﬂora L., 331, 355 Cytisus scoparius L. (Link), 421–423 Dacrydium cupressinum Soland. ex Forst. f., 209 Dacryodes excelsa Vahl., 257, 638, 749 Dactylis glomerata L., 372 Dactylotum bicolor Scudder, 649 Daedaleopsis septentrionalis (P. Karst.) Niemel¨a, 167 Danthonia montevidensis Hack. & Arech., 491 Dasiphora fruticosa Rydb. ( = Potentilla fruticosa L.), 61 Datura ferox L. (charnico), 494 D. innoxia J.S. Miller, 430 Daucus carota L., 614, 619 Dendrocnide excelsa (Wedd.) Chew, 255, 256 Dendroctonus, 201, 259 D. frontalis Zimmerman, 259–261 D. ponderosae Hopkins, 261 D. ruﬁpennis (Kirby) (spruce bark beetle), 171, 260, 455, 751 Dendroica caerulescens (Gmelich), 642 D. fusca (M¨uller), 642 D. magnolia (Wilson), 642 Dendrolimus sibiricus L. (Siberian silkworm), 169 Deschampsia, 45, 53 D. alpina (L.) Roemer & J.A. Schultes, 26 D. antarctica Hack., 45, 53, 77, 80, 91 D. cespitosa (L.) Beauv., 180 D. ﬂexuosa (L.) Trin., 173, 180 Desmanthus, 497 Dichondra microcalyx (Hallier) Fabris. (oreja de rat´on), 494 Dicksonia squarrosa (Forst.) Sw., 209 Dicranum, 170, 173, 178, 180 D. polysetum Sw., 178 Dicrostonyx, 55

SYSTEMATIC INDEX D. torquatus (Pallas), 55 Diervilla lonicera Mill., 178 Digitaria diversinervis Stapf., 380 D. sanguinalis (L.) Scop. (pata de gallina, pasto de cuaresma), 494 Diomedea exulans L., 59, 91 D. melanophris Temminck, 59 Diospyros virginiana L., 377 Diplopoda, 640 Dipodomys, 319, 645 D. ordii Woodhouse, 647 D. spectabilis Merriam, 645 Dipsacus, 416 D. fullonum L., 498 Diptera, 636, 641 Discaria longispina (Hook & Arn.) Miers. (brusqu´ılla), 498 Distichlis spicata (L.) Greene (saltgrass), 351, 356, 574 Doryphora sassafras Endl., 255, 256, 259 Drechslera, 532 Dreissenia (zebra mussels), 432 Drepanocladus, 178 Drimys winteri J.R. Forst & G. Forst., 204, 208 Drosophilidae, 636 Dryas, 82, 601 D. drummondii Richards., 600, 601 Dryopteris disjuncta (Rupr.) Morton, 178 D. expansa (C. Presl.) Fraser-Jenkins & Jermy, 180 D. spinulosa (O.F. Muell.) Watt, 178 Dumetella carolinensis (L.), 642 Dupontia ﬁsheri R. Br., 83 Dysithamnus puncticeps Salvin, 642 Echinochloa colonum (L.) Link. (pasto colorado), 494 Egeria, 433 Ehrharta calycina Sm., 439 Eichhornia crassipes (C. Martius) Solms-Laubach, 432, 433 Elaphidion, 341 Eleagnus, 429 E. angustifolia L., 428, 429, 431 Elephantidae, 337 Elephas (elephants), 751 Eleutherodactylus coqui Thomas, 636, 638 Elymus farctus (Viv.) Runemark ex Melderis, 436 Elyna myosuroides (Vill.) Fiori, 25 Elytrigia repens (L.) Desv. ex B.D. Jackson, 86 Emballonura nigrescens ( = Mosia nigrescens (Gray)), 643 Empetrum, 90, 166 E. hermaphroditum Hagerup, 165, 180 E. nigrum L., 170 Empidonax ﬂaviventris (Baird & Baird), 642 Encelia, 310 E. farinosa Gray ex Torr., 324 Endothiella parasitica (Murrill) H.W. Anderson, 455 Entomophaga maimaiga Humbert, 444 Epilachna vigintioctopunctata (Fabricius), 475 Epilobium angustifolium L. ( = Chamaenerion angustifolium), 150, 152, 166, 170 Epirrita, 171, 172

SYSTEMATIC INDEX E. autumnata (Bkh.), 171 Eptesicus pumilus Gray, 643 E. sagittula KcKean et al., 643 Equisetum sylvaticum L., 180 Equus asinus L., 324 E. caballus L., 438, 681 Eragrostis boehmii Hack., 374 E. curvula (Schrad.) Nees, 439 E. lehmanniana Nees, 418, 423, 648 E. lugens Nees, 490, 491 E. trichodes (Nutt.) Wood, 648 Eria robusta Lindl., 637 Erigeron annuus (L.) Pers., 618 Erinaceus europaeus L. (hedgehog), 407 Eriophorum angustifolium ssp. subarcticum (Vassiljev) Hulten ex Kartesz & Gandhi, 81 E. vaginatum L., 90 ssp. spissum (Fern.) Hulten, 80–82 Erodium, 418, 421 E. cicutarium (L.) L’Her. ex Ait., 278, 749 Erythronium americanum Ker-Gawl., 208 Eschrictius robustus (Lilljeborg) (gray whale), 90 Eubacteria, 522 Eucalyptus, 256, 263–266, 271, 276, 392, 424, 425, 469, 470, 480, 572, 613, 667 E. blakelyi Maiden, 264 E. calophylla R. Br., 381 E. camaldulensis Dehnh. (river red gum), 429 E. globulus Labill., 280, 471 E. incrassata Labill., 573 E. marginata Donn. ex Smith, 263, 281, 380, 381 E. megacarpa F. v. M., 380 E. nova-anglica Deane & Maiden, 263, 264 E. patens Benth., 381 E. salmonophloia F. Muell. (salmon gum), 676 Euchordrus, 131 Eudyptes chrysolophus (Brandt), 91 Eupatorium subhastatum Hook. & Arn., 491 Euphausia superba Dana, 59 Euphonia musica (Gmelin), 636 Euphorbia esula L., 392 E. fendleri T. & G., 649 E. lasiocarpa Klotz., 494 E. ovalifolia (Klotz. et Garek.) Boiss., 494 E. paralias L., 436 E. peplus L., 398 Euphydryas editha bayensis (Boisduval) (bay checkerspot butterﬂy), 718 Eutamias minimus Bachman ( = Tamias minimus), 213 Fabaceae, 274 Facelis retusa (Lam.) Sch.-Bip., 491 Fagus, 188, 192, 195, 205, 211, 213, 262 F. crenata Blume, 195, 204, 207 F. grandifolia Ehrh., 192, 203, 204, 206, 210, 378, 442, 548, 552, 622 Fallopia japonica (Houtt.) Ronse ( = Reynoutria japonica), 427, 429, 431

825 Felis catus L. (cat), 69, 417 F. concolor Jardine ( = Puma concolor) (cougar), 699 Ferocactus, 311 Festuca, 374 F. arundinacea Schreb., 497 F. ovina L., 375 F. pallescens (St. Yves) Parodi, 287 F. pratensis Hudson, 614 F. rubra L., 154 F. trachyphylla (Hack.) Krajina, 406 F. vivipara (L.) Sm., 26 Ficus, 310 F. erecta Thunb., 142, 143 Filipendula ulmaria (L.) Maxim., 180 Flourensia cernua DC., 318 Frankenia salina (Molina) I.M. Johnston [syn: Frankenia grandifolia] (alkali seaheath), 356 Frankia, 523 Fraxinus, 205 F. americana L., 201, 205, 206 F. ornus L., 427, 430 Freycinetia reineckeri Warb., 637 Funaria hygrometrica Hedw., 166 Gaeotis nigrolineata (Shuttleworth), 637, 638 Galium triﬂorum Michx., 178 Gallus domesticus (L.) ( = G. gallus), 129 Geaster, 528 Gentiana lutea L., 55 Geomys (pocket gophers), 645 G. bursarius Shaw (pocket gopher), 580, 647 Geotrygon montana L., 636 Geranium bohemicum L., 166 G. lanuginosum Lam., 166 G. sylvaticum L., 180 Gleditsia triacanthos L. (acacia negra, corona de cristo), 205, 498, 623 Gloiodon strigosus (Sw.: Fr.) P. Karst., 167 Glomus, 528 G. mosseae (Nicol. & Gerd.) Gerdemann & Trappe, 528 Glycine max L. (soybean), 490, 494, 512 Glycyphagus michaeli, 528 Gnaphalium norvegicum Gunn., 180 Goodyera, 178 Gopherus agassizii Stejneger, 318 Gryllidae, 639 Gutierrezia sarothrae (Pursh) Britt & Rusby, 389 Gymnocarpium dryopteris (L.) Newm., 180 Gymnopithys leucaspis (Sclater), 642 Gypaetus barbatus (L.), 58 Hakea, 425 H. sericea Schrad. J. Wendl, 281, 439, 440 Halogeton, 418 Haploporus odorus (Sommerf.: Fr.) Sing., 167 Haumaniastrum katangense (S. Moore) P.A. Duvign. & Plancke, 374 H. robertii (Robyns) P.A. Duvign. & Plancke, 374

826 Hedeoma multiﬂora Benth., 491 Heimia salicifolia Link (quiebra arado), 491 Helianthus annuus L. (sunﬂower), 490 Heliconia, 758 Hemilepistus reaumuri Audouin, 312 Hemiptera, 641 Heracleum mantegazzianum L., 430, 431 Herminium monorchis (L.) R. Br., 371 Heteropsylla cubana Crawford, 436, 443 Hieracium albiﬂorum Hook., 151 H. pratense Tausch, 618 Hipposideros ater Templeton, 643 H. diadema (E. Geoffroy), 643 Holcus lanatus L., 154, 439, 443, 626 Homoptera, 641 Hoplostines viridipennis Blackburn, 255 Hordeum, 398 H. vulgare L. (barley), 59, 515 Humicola, 532 Hydrocharis morsusranae L., 434 Hydrodamalis gigas (Zimmermann) (Steller’s sea cow), 58 Hylocomium splendens (Hedw.) Schimp., 173, 180 Hylophylax naevioides (Lafresnaye), 642 Hylurgopinus ruﬁpes (Eichhoff), 262 Hymenoclea salsola Torr. & Gray ex Gray, 316 Hyparrhenia rufa (Nees) Stapf., 423, 439 Hypochaeris, 491 H. radicata L., 154, 497 Hystrix indica Kerr (Indian crested porcupine), 312, 713 Impatiens glandulifera Royle, 430 Ipomopsis aggregata (Pursh) Grant, 579 Ips, 201, 259, 443 I. interstitialis Eichhoff, 436 Iridomyrmex humilis (Mayr), 281 Isoglossa woodii C. B. Cl., 380 Jatropha, 310 Jaumea carnosa (Less.) Gray (marsh jaumea), 356 Juglans nigra L., 205, 378 Juncus kraussii Hochst., 343 J. roemerianus Scheele (black needlerush), 349, 353 Juniperus, 130 J. communis L. (juniper), 180, 618 Kobresia humilis (C.A. Mey. ex Trautr.) Serg., 61 K. myosuroides (Vill.) Fiori, 80 K. pygmaea C.B. Clarke, 60, 61, 68, 80, 82 Kochia scoparia (L.) Roth, 389 Lactuca, 406 L. alpina (L.) A. Gray, 180 Lagopus, 58 Lama glama L. (llama), 60 L. guanicoe M¨uller (guanicoe), 60 L. pacos L. (alpaca), 60 Lamponius portoricensis (Rehn) (walkingstick), 257, 637 Lantana camara L., 436, 440

SYSTEMATIC INDEX Larix, 144, 162, 163, 200 L. cajanderi Mayr., 146 L. dahurica Lawson (Dahurian larch), 163, 167 L. laricina (Duroi) K. Koch, 175 L. occidentalis Nutt., 457 L. sibirica Ledeb. (Siberian larch), 163, 167 L. sukaczeweii Dylis, 163 Larrea, 321 L. tridentata (Sesse and Moc.) Coville (creosotebush), 287, 310, 312, 316, 317, 389 Larus, 70 Ledum palustre L., 180 Leersia oryzoides (L.) Sw. (rice cutgrass), 342 Lemmus, 55 Lepidoptera, 580, 637, 641 Leptographium engelmannii Davids. (blue stain fungus), 455, 751 Leptospermum laevigatum (J. Gaertn.) F. Muell., 436, 439 Lepus, 58, 593 L. americanus Erxleben, 55, 170 L. californicus Gray, 314, 318, 646 L. europaeus Pallas (hare), 398 Lerchenfeldia ﬂexuosa (L.) Schur, 146 Leucaena, 480 L. leucocephala (Lam.) de Wit, 443, 444 Leucanthemum vulgare Lam. (daisy), 497 Leucauge regnyi Simon, 636 Leycesteria formosa Wall., 436 Leymus interior (Hulten) Tzvelev, 86, 146 Ligustrum, 498 Linnaea borealis L., 178, 180 Linum usitatissimum L. (linseed), 490 Liriodendron tulipifera L., 205, 206, 552, 554 Listera cordata (L.) R. Br., 180 Littorina irrorata Say, 353 Lobodon carcinophagus (Hombron & Jacquinot) (crabeater seal), 59 Locustana, 318 Lolium, 398 L. perenne L., 127, 439, 626 Lonicera canadensis Bartr., 178 L. maackii (Rupr.) Maxim., 416 Lotus corniculatus L., 372 L. tenuis Waldst. & Kit. (tr´ebol pata de p´ajaro), 497 Loudetia simplex (Nees) C.E. Hubb., 374 Loxigilla portoricensis Daudin, 642 Loxodonta (elephants), 751 Lumbricidae, 443 Lupinus, 145, 153 L. arboreus Sims, 154, 419 L. lepidus Dougl. var. lobbii (Gray) Hitchc., 152, 153 Luzula pilosa (L.) Willd., 166, 180 Lycoperdon, 528 Lycopodium annotinum L., 180 Lymantria dispar L. ( = Porthetria dispar), 260, 437, 443, 455, 552 Lythrum salicaria L. (purple loosestrife), 348, 428, 431, 433, 434 Macaranga, 758 Macronectes giganteus (Gmelin), 54, 59

SYSTEMATIC INDEX Macrozamia riedlei (Fisch ex Gaud.) C.A. Gardn., 381 Maesopsis eminii Engl., 436, 441 Magnolia obovata Thunb., 207 Maianthemum bifolium (L.) F. W. Schm., 180 M. canadense Desf., 178 Mammalia, 398, 399, 404, 407 Manihot, 467 Mantidae, 639 Margarops fuscutus (Vieillot), 636 Margyricarpus pinnatus (Lam.) O.K. (yerba de la perd´ız), 491 Marisa cornuarietis (L.), 432, 433 Marmota, 82, 513 Martes foina Erxleben (stone marten), 398 Matteuccia struthiopteris (L.) Tod., 180 Maytenus boaria Molina, 279 Medicago, 497 M. lupulina L., 372 M. sativa L. (alfalfa), 129, 515 Megalomastoma croceum Gmelin, 643 Megascolecidae, 443 Melaleuca, 392, 431, 434 M. quinquenervia (Cav.) Blake, 433, 434 M. uncinata R. Br., 573 Melampus bidentatus Say (eastern melampus), 340, 349 Melampyrum pratense L., 180 M. sylvaticum L., 180 Melanerpes portoricensis (Daudin), 636 Melanoplus, 649 Melastomataceae, 477 Melia azedarach L. (para´ıso), 498 Melica brasiliana Ard., 490, 491 Melilotus alba Med. (tr´ebol de olor blanco), 372, 497 M. ofﬁcinalis (L.) Pallas, 623 Melinis minutiﬂora Beauv., 439 Mesostigmata, 372 Metrosideros, 152 M. polymorpha Gaud., 152, 154 Miconia calvescens DC., 426 Micropsis spathulata (Pers.) Cabr., 491 Microsorum scolopendria (Burm. f.) Copeland, 380 Microstegium vimineum (Trin.) A. Camus, 427, 428 Microtus, 398 M. ochrogaster Wagner (prairie vole), 640 Milium effusum L., 180 Millsonia anomala (Omodeo & Vaillaud), 509 Mimosa pigra L., 429, 431, 439, 441 Mimusops caffra E. Mey. ex A. DC., 380 Minuartia verna (L.) Hiern, 374, 375 Mirounga leonina (L.), 54 Miscanthus sinensis Anders., 143 Mitella nuda L., 178 Modisimus signatus Simon, 636 Monophyllus redmani Leach, 637 Mormopterus beccarii Peters, 643 M. planiceps (Peters), 643 Morus alba L. (mora), 498 Muehlenbeckia axillaris (Hook.) Walp., 143 Murdannia keisak (Hassk.) Hand.-Maz. (Asiatic dayﬂower), 348

827 Mus musculus L. (house mouse), 69, 398 Myocaster coypus (Molina) ( = M. bonariensis) (nutria), 337, 751 Myospalax, 56 Myrica, 154 M. faya Aiton (faya tree), 153, 154, 443 Myriophyllum aquaticum L., 430 Myristica malabarica Lam., 472 Nardus stricta L., 87 Nassauvia glomerulosa (Lag.) Don, 287 Nebria, 26 Nematoda, 523 Nenia tridens (Schweigger), 637, 638, 643 Neospingus speculiferus (Lawrence), 636 Nicotiana glauca Graham, 324, 430 Nostoc commune Vaucher, 42 Nothofagus, 188, 204, 206–208, 211, 287 N. antarctica Macloskie, 204 N. betuloides (Mirb.) Bl., 147, 204 N. dombeyi (Mirb.) Oerst., 204, 206 N. menziesii (Hook. f.) Oerst., 211 N. pumilio Krasser, 204 Nuphar luteum (L.) J.E. Smith (spadderdock), 336 Nycticeius balstoni (Thomas) subgenus scotorepens, 643 Nyctophilus gouldi Tomes, 643 Nyssa, 343 N. aquatica L., 200 N. sylvatica Marsh., 204 Ochna natalitia (Meisn.) Walp., 380 Ochotona, 56 O. princeps (Richardson), 55 Ochroma, 758 Odobenus rosmarus L. (Paciﬁc walrus), 58 Odocoileus, 170 O. hemionus (Raﬁnesque) (black-tailed deer), 150, 529, 697 O. virginianus (Zimmermann) (white-tailed deer), 208, 751 Oenothera, 318, 406 Olea europaea L. (olive), 276 Oleacina playa H.B. Baker, 643 Oligochaeta, 332 Olneya, 310 O. tesota Gray, 311, 317 Ondatra (muskrats), 351, 355, 751 O. zibethicus L. (muskrat), 332, 337 Onomacris, 316 Ononis repens L., 372 Onychomys torridus (Coues), 389 Operophtera brumata (L.) Ochrana Rosthn., 444 Ophrys apifera Huds., 371 O. insectifera L., 371 Opuntia, 325 Oreamnos americanus (de Blainville) (mountain goat), 150 Orgyia pseudotsugata (McDunnough) (Douglas-ﬁr tussock moth), 260, 261 Oribatida, 523 Oryctolagus cuniculus (L.) (rabbit), 55, 68–70, 279, 282, 324, 398, 426, 438, 443, 677, 681

828 Oryza sativa L. (rice), 129, 481, 665, 691 Oryzopsis hymenoides (Roemer & Schultes) Ricker ex Piper, 324 Ostracoda, 332 Ostrya virginiana (Correll) Henrickson, 207, 208 Otariidae (seals), 337 Otomys typus Heuglin ( = O. orestes) (groove-toothed rat), 56 Ovibos moschatus (Zimmermann) (muskoxen), 55, 83, 84, 90 Ovis, 60 O. aries L., 58 O. canadensis Shaw (bighorn sheep), 58 Oxalis, 491 O. acetosella L., 180 O. chrysantha Prog. (vinagrillo), 494

Panicum bergii Arech. (paja voladora), 491 P. coloratum L., 578 P. hemitomon J.A. Schultes (maidencane), 336, 342, 349, 351 P. maximum Jacq., 380, 439 Panthera onca (L.), 473 Parkinsonia aculeata L. (cina-cina), 498 Parmeliopsis, 173 Parthenium incanum (Kunth), 318 Paspalum, 428 P. dilatatum Poir. (pasto miel), 490, 491 P. distichum L. ( = P. paspalodes) (gramilla blanca), 430, 494 P. notatum Flueg. (pasto horqueta), 491 P. quadrifarium Lam. (paja colorada), 498 Passer domesticus L. (house sparrow), 398 Passiﬂora mollisima (Kunth) L.H. Bailey, 443 Paulownia tomentosa Steud., 425, 428 Penicillium, 529 P. brevi-compactum Dierckx, 533 P. spinulosum Thom, 533 Pennisetum ciliare (L.) Link, 324 P. americanum (L.) Leeke, 379 P. ciliare (L.) Link, 317, 418, 423 P. polystachyon (L.) Schult., 439 P. setaceum (Forsk.) Chiov., 439 Perga afﬁnis Kirby, 264 Perognathus ﬂavescens Baird, 647 P. longimembris (Coues), 389 Peromyscus leucopus (Raﬁnesque) (white-footed mouse), 213, 389 P. maniculatus Wagner (deer mouse), 529, 640 Petasites japonicus (Sieb. & Zucc.) Maxim., 151 Peumus boldus Molina, 279 Phaeocystis, 42 Phalaris aquatica L. (mata dulce, falaris), 497 Phaseolus vulgaris L. (bean), 59, 515 Phasianus colchicus L., 639 Pheidole dentata Mayr, 647–649 Pholcidae, 636 Phragmites, 332 P. australis (Cav.) Trin. ex Steud. (common reed), 343 Phyla canescens (H.B.K.) Greene, 491 Physalis viscosa (L.) (camamb´u), 494 Phytophthora, 282 P. cinnamomi Rands, 263, 282, 381, 442

SYSTEMATIC INDEX Picea spp. (spruce), 162, 163, 166, 170, 188–190, 200, 213, 577, 710, 751 P. abies (L.) Karst. (Norway spruce), 162–164, 167, 173, 177, 181, 204, 210, 548, 554 P. engelmannii Parry ex Engelm., 50, 204 P. glauca (Moench) Voss (white spruce), 30, 163, 164, 167, 170, 175, 545 P. mariana (Mill.) B.S.P. (black spruce), 163, 167, 170, 175 P. obovata Ledeb. (Altai spruce), 163, 167 P. rubens Sarg. (red spruce), 548, 552 P. sitchensis (Bong.) Carr., 31, 454, 548, 554, 600, 601 Picoides borealis (Vieillot), 392 Picris hieracioides L., 151 Pinnipedia (seals), 751 Pinus spp. (pines), 154, 162, 163, 166, 189, 190, 194, 201, 203, 206, 207, 209, 210, 260–262, 274, 276, 398, 439, 442, 469, 470, 477, 480, 529, 554, 617 P. banksiana Lamb. (jack pine), 162–165, 175, 189, 454, 572, 617 P. cembra L., 533 P. clausa Vasey ex Sarg., 391 P. contorta Dougl. (lodgepole pine), 162–165, 170, 174, 190, 212, 455 P. elliottii Engelm., 552 P. halepensis Mill., 424, 436, 441 P. mugo ssp. mughus Turra, 49 P. palustris P. Mill. (longleaf pine), 333, 391, 392, 454, 456 P. patula Schiede ex Schltdl. & Cham., 430 P. pinaster Ait., 281, 436, 439 P. ponderosa Laws. (ponderosa pine), 257, 261, 454, 456, 457, 459, 574 P. pumila (Pall.) Regel (dwarf pine), 146, 163 P. radiata D. Don (monterey pine), 280, 281, 424, 461, 525 P. resinosa Ait. (red pine), 163–165, 175, 201, 572 P. rigida P. Mill. (pitch pine), 190, 212, 525, 713 P. serotina Michx. (pond pine), 349, 355 P. sibirica Du Tour (stone pine), 163, 167 P. strobus L. (white pine), 163–165, 175, 194, 201, 378, 613, 760 P. sylvestris L. (Scots pine), 163–165, 168, 170, 177, 180, 181, 204, 549, 554, 579 P. taeda L., 200, 454, 456, 550, 554, 619 P. thunbergii Parl., 142, 143 Pipra mentalis Sclater, 642 Piptochaetium bicolor (Vahl) Desv., 491 P. montevidense (Spreng.) L.R. Parodi, 491 Pisolithus tinctorius Alb. & Schwein, 529 Pitcairnia sulphurea K. Koch, 143 Pithecellobium, 310 Pittosporum undulatum Vent., 429–431 Plantago lanceolata L., 374, 404 Platanus, 428 Platypodidae, 636 Platyrinchus coronatus Sclater, 642 Platysuccinea portoricensis (Shuttleworth), 638 Pleurozium schreberi (Brid.) Mitt., 170, 173, 178, 180 Plutella xylostella L., 89 Poa, 406

SYSTEMATIC INDEX P. alpina var. vivipara (L.) Arcang., 26 P. annua L., 78, 88 P. arctica R. Br., 82 P. bonariensis (Lam.) Kunth., 491 P. compressa L., 154 P. ﬂabellata Hack., 52, 83, 91 P. platyantha Kom., 146 P. pratensis L., 372, 620 Poaceae, 274 Pogonomyrmex, 648 P. barbatus F. Smith, 647–649 P. rugosus Emery, 531 Pohlia gracilis (Hoppe & Hornsch.), 29 Polychaeta, 332 Polydontes acutangula (Burrow), 637, 638 Polygala australia A. W. Benn., 491 Polygonum aviculare L. (sanguinaria), 494 P. japonicum Meissm., 143 P. weyrichii F. Schmidt & Petrop., 144 Polytrichum, 180 P. juniperinum Hedw., 146, 166 P. piliferum Hedw., 166 Pontederia cordata L. (pickerel-weed), 336 Populus, 26, 166, 168, 201, 203, 428, 429, 431, 531 P. balsamifera L. (mountain alder), 533 P. deltoides Marsh., 377, 431 P. suaveolens Fisch, 146 P. tremula L., 166, 170, 204 P. tremuloides Michx., 167, 175, 203, 210, 613 P. trichocarpa Torr. & Gray, 600 Porthetria, see Lymantria Portulaca oleracea L. (verdolaga), 494 Prasiola crispa (Lightfoot) Kutzing, 53, 66 Prionium serratum (L. f.) Drege ex E. Mey., 441 Procavia capensis Pallas ( = P. johnstoni, P. mackinderi) (hyrax), 56 Procellariidae, 54 Prochimys, 529 Prosopis, 429, 430, 470, 479 P. caldenia Burk. (cald´en), 498 P. glandulosa Torr., 318 Protea goetzeana Engl., 374 Proteaceae, 282, 759 Protozoa, 523 Prumnopitys ferruginoides (Comption) de Laub., 209 Prunus, 203, 205, 212 P. pensylvanica L., 206 P. serotina Ehrh., 205, 206, 208, 209, 213, 614, 619, 623 P. virginiana L., 213 Pseudotsuga, 262, 622 P. douglasii (Sabine ex D. Don) Carriere, 170 P. menziesii (Mirb.) Franco, 145, 202, 203, 211, 261, 262, 454, 551, 554, 613, 622, 704 Psidium cattleianum Sabine, 436, 443 Pterepodidae, 636 Pteridium aquilinum (L.) Kuhn, 178 Pteropus samoensis Peale, 636, 637 P. tonganus Quoy & Gaimard, 636–638

829 Ptilium crista-castrensis (Hedw.) De Not., 180 Puccinellia distans (Jacq.) Parl., 403, 407 P. phryganodes (Trin.) Scribn. & Merr. (creeping alkaligrass), 83, 353, 579 Pygoscelis adeliae (Hombron & Jacquinot), 55, 91 P. antarctica (Forster), 91 P. papua (Forster), 53 Pyrola media Sw., 167 Quercus spp. (oaks), 189, 190, 201, 203, 205, 206, 213, 262, 271, 274–276, 278, 279, 377, 398, 400, 403, 455, 530, 552, 589, 591, 614, 710 Q. alba L., 259, 378, 552 Q. coccifera L. (garrigue), 573 Q. coccinea Muench., 259 Q. havardii Rydb., 635, 647 Q. ilex L., 276 Q. ithaburensis Decaisne, 276 Q. muehlenbergii Engelm., 205 Q. prinus L., 552 Q. pubescens Willd., 276, 533 Q. robur L., 401, 403 Q. rubra L. (red oak), 205, 206, 210, 261 Q. suber L., 276 Quillaja saponaria Molina, 279 Racomitrium canescens (Hedw.) Brid. ( = Rhacomitrium), 24, 145 Ramalina sinensis Jatta, 167 Rana berlandieri Baird, 432 R. catesbiana Shaw (common American bullfrog), 429, 432 Randia, 310 Rangifer (caribou, reindeer), 55, 58, 60, 750 R. tarandus (L.) ( = Tarandus rangifer) (reindeer), 23, 44, 54, 55, 58, 62, 64, 69, 83, 84, 92, 168–170 Rattus spp. (rats), 70, 529, 681 R. norvegicus L. (brown rat), 69, 398 Regulus calendula (L.), 642 R. satrapa Lich., 642 Reithrodontomys megalotis (Baird) (harvest mouse), 640 Rendlia cupricola Duvign., 374 Reynoutria japonica ( = Polygonum cuspidatum Sieb. & Zucc.), 427, 430 R. sachalinensis F. Schmidt Nakai, 430 Rhanterium suaveolens Desf., 436 Rhinolopus megaphyllus Gray, 643 Rhizobium, 406, 523 Rhizophora mangle L. (red mangrove), 341 Rhododendron maximum L., 208 Rhus ambigua Lavall., 143 Rhytidiadelphus triquetrus (Hedw.) Warnst., 180 Ribes glandulosum Grauer, 178 R. lacustre (Pers.) Poir., 178 Ricinus communis L., 430 Robinia, 406 R. pseudoacacia L., 378, 406 Rosa acicularis Lindl., 178 R. polyantha Sieb. & Zucc., 143 Rubus, 197, 205, 206, 212

830 R. alceifolius Poir., 436, 440 R. allegheniensis Porter, 205 R. chamaemorus L., 180 R. fruticosus L., 436 R. idaeus L., 178 R. pubescens Raf., 178 R. sachalinensis Levl., 150 R. saxatilis L., 180 R. spectabilis Pursh., 152 Rupicapra rupicapra (L.), 30 Sagina apetala Ard., 89 Sagittaria lancifolia L., 331, 342 S. latifolia Willd. (broadleaf arrowhead), 336 Salicornia, 349 S. europaea L., 343, 349, 350 S. virginica L. (perennial glasswort), 340, 356 Salix, 48, 55, 145, 152, 170, 428, 431, 531 S. hultenii Floderus ( = S. caprea L.), 146, 151, 180 S. nigra Marsh, 377 Salsola iberica Sennen Pav. ( = S. tragus L.), 416 S. kali L., 532 Sambucus, 170 S. nigra L., 406 Sapium, 310 S. sebiferum L. (Roxb.), 420, 425 Sarcorampus papa ( = Vultur gryphus L.), 53 Sarracenia, 340 Sasa nipponica (Makino) Makino & Shibata, 207 Sassafras, 205 Schinus molle L., 429, 432 S. terebinthifolius Raddi, 420, 439, 441 Schismus barbatus L., 423, 439 Schistocerca, 318 S. alutacea Harris, 649 S. gregaria (Forsk.), 257 Schizachyrium condensatum (Kunth) Nees, 439, 442 S. scoparium (Michx.) Nash, 619, 623, 648 S. spicatum (Spreng.) Herter., 491 Scirpus americanus Pers. [syn.: Scirpus olneyi Gray] (Olney’s three-square), 343 Sciurus, 529 Scolytidae, 254, 258, 636 Scolytus, 259 S. multistriatus (Marsham) (bark beetle), 262, 455, 751 Scotorepens, see Nycticeius Seiurus aurocapillus (L.), 642 S. noveboracensis (Gmelin), 642 Senecio burchelii DC., 497 S. jacobaea L. (tansy ragwort), 154, 421, 422 S. sylvaticus L., 618 S. vulgaris L., 421 Sequoia sempervirens (D. Don) Endl., 189 Sesleria albicans Kit. ex Schult., 371 Setaria geniculata (Lam.) Beauv., 494 S. parviﬂora (Poir.) Kerguelen, 491 Setophaga ruticilla (L.), 642 Shorea albida Sym., 225

SYSTEMATIC INDEX S. japonica ( = S. sumatrana Van Slooten), 691 S. parvifolia Dyer, 239 Silene vulgaris ssp. maritima (Moench) Garcke, 375 Sinarundinaria fangiana (A. Camus) Keng ex Keng f., 207 Sirex noctilio Fabricius, 436 Sisymbrium altissimum L., 406 Soja max (L.) Piper (soybean), 129 Solanum, 406 S. mauritianum Scop., 436 S. tuberosum (potato), 59, 129 Solenopsis, 759 Solidago, 625 S. altissima L., 614 S. canadensis L., 613, 618 S. rugosa P. Mill., 178 S. virgaurea L., 180 Sonchus oleraceus L. (cerraja), 494 Sorbus aucuparia L., 180 Sorghum, 379 S. bicolor (L.) Moench (grain sorghum), 490 S. halepense (L.) Pers. (sorgo de alepo), 494, 498 S. vulgare (sorghum), 129 Spartina (cordgrass), 338 S. alterniﬂora Loisel. (smooth cordgrass), 342, 349, 351, 353 S. cynosuroides (L.) Roth (big cordgrass), 353 S. patens (Ait.) M¨uhl. (saltmeadow cordgrass), 343, 349, 351, 355 Speotyto cunicularia Molina, 645 Spermophilus parryii (Richardson), 55, 56 S. tridecemlineatus (Mitchill), 213 Sphaeralcea ambigua A. Gray, 324 S. bonariensis (Cav.) Griseb. (malvavisco), 498 Sphaeroma rugicauda Leach (isopod), 349 Sphagnum, 31, 180, 577 Spiraea media Schm., 168 Spizella breweri Cassin, 390 S. passerina (Bechstein), 390 Sporobolus contractus A.S. Hitchc., 318 S. cryptandrus (Torr.) Gray, 648 S. indicus (L.) R. Br., 490 Stellaria media (L.) Vill. (capiqu´ı), 494 Stenoderma rufum Desmarest, 637, 638 Stereocaulon, 24, 146, 177 Stipa, 287 S. hyalina Nees (ﬂechilla mansa), 491 S. leucotricha Trin. & Rupr. (Texas wintergrass), 574 S. neesiana Trin. et Rupr. (ﬂechilla), 491 S. papposa Nees, 490, 491 Streptomyces, 694 Strix occidentalis (X. de Vesey), 698 Sturnella neglecta Aud., 390 Subulina octana Bruguiere, 643 Sus scrofa L. (pig), 435, 438, 442, 681 Swietenia, 469, 472, 478, 480 S. macrophylla King, 472, 482 S. mahagoni (L.) Jacq., 472 Sylvilagus audubonii (Baird), 318, 647

SYSTEMATIC INDEX Syzygium aromaticum (L.) Merr. & L.M. Perry ( = Eugenia aromatica Baill.), 691 S. gardneri Thw., 472 S. inophylloides (A. Gray) C. Mueller, 637 Taeniatherum asperum (L.) Nevski, 423 T. caput-medusae (L.) Nevski, 439 Tagetes minuta L. (chinchilla), 494, 495 Tamarix spp. (tamarisks), 324, 428, 429, 431, 432, 434, 445 T. aphylla L. Karsten, 429, 431 T. ramosissima Ledeb., 431 Tamias striatus (L.) (see also Eutamias), 213 Tapirus, 473 Tarandus, see Rangifer Taraxacum ofﬁcinale Weber ex Wiggers, 622 Taxidea taxus Schreber, 529, 617, 645 Taxodium, 343 T. distichum (L.) L.C. Rich. (bald cypress), 200, 332, 341 Taxus canadensis Marsh., 178 Teclea gerrardii Verdoorn, 380 Tectona, 480 T. grandis L., 470 Thelypteris novaboracensis (L.) Nieuwl., 208 T. phegopteris (L.) Slosson, 180 Theobroma cacao L., 470 Thlaspi caerulescens J. & C. Presl., 374, 375 Thomomys, 150, 529 T. bottae (Eydoux & Gervais), 422, 529 T. talpoides (Richardson) (pocket gopher), 54, 55, 528 Threnetes ruckeri (Bourcier), 642 Thuja, 262 T. occidentalis L., 164, 175, 189 T. plicata (Donn.), 262 Thysanoptera, 532 Tilia americana L., 210 Todus mexicanus Lesson, 636 Toxostoma lecontei Lawrence, 390 Tragia geraniifolia Klotzsch. (ortiga), 491 Trema, 758 T. orientalis (L.) Bl., 380 Trichechus manatus L. (West Indian manatee), 332 Trichilia emetica Vahl, 380 Trichosurus vulpecula (Kerr.), 438, 441, 681 Trientalis borealis Raf., 178 T. europea L., 180 Trifolium, 154, 497 T. argentinense C. Spegazzini (tr´ebol), 497 T. parryi ssp. parryi Gray, 86 T. polymorphum Poir. (tr´ebol), 497 T. repens L., 154, 626 Tristachya helenae Buscal. & Muschl., 374 Tristaniopsis, 429 T. laurina (Smith) P.G. Wilson & Waterhouse, 431 Triticum aestivum L. (wheat), 59, 126, 129, 490, 515 Tsuga, 189, 190, 192, 193 T. canadensis (L.) Carr, 192, 194, 203, 205, 208, 212, 260

831 T. heterophylla (Raf.) Sarg., 29, 211, 454, 600 T. mertensiana (Bong.) Carr., 454 Turdus plumbeus L., 636 Tussilago, 406 Typha, 332, 336, 355, 428 T. angustifolia L. (narrow-leaved cattail), 348, 434 T. domingensis Pers. (southern cattail), 340 T. orientalis C. Presl. (cumbungi), 343 Uca pugnax (Smith) (ﬁddler crab), 337 Ucides cordatus L. (hairy land crab), 338 Ulmus alata Michx., 377 U. americana L., 377, 455 U. pumila L. (olmo siberiano), 498 Ursinia, 421 Ursus, 70 U. americanus Pallas (American black bear), 150 U. arctos L. ( = Ursus horribilis) (grizzly bear), 30, 695 U. maritimus Phipps (polar bear), 58, 64, 65, 68, 70 Urtica, 398 U. dioica L., 407 Uvularia grandiﬂora Sm., 213 Vaccinium, 166, 180 V. macrocarpon Ait. (cranberry), 334 V. myrtillus L., 166, 173, 180 V. uliginosum L., 180 V. vitis-idaea L., 166, 170, 180 Vaginulus occidentalis Guilding, 643 Valeriana sambucifolia Mikan, 180 Vepris lanceolata (Lam.) G. Don, 380 Verbascum thapsus L., 618 Verbena, 491 Vernonia rubricaulis H. & B. (quiebra-arado), 491 Veronica persica Poir., 495 Vicia, 491 Vicugna vicugna (Molina), 58 Viola, 178 V. calaminaria Lej., 372, 374 Vireo olivaceus (L.), 642 Vulpes vulpes (L.) (fox), 417 Weinmannia racemosa L., 147, 211 Wilsonia canadensis (L.), 642 Xanthium cavanillesii Schoum. (abrojo), 494 X. spinosum L., 494 Zapus hudsonius (Zimmermann), 213 Zea mays L. (corn/maize), 59, 125, 127, 129, 467, 481, 490, 510–512, 515 Zonotrichia albicollis (L.), 642 Z. leucophrys (Forster), 390 Zygogramma tortuosa Rogers, 318

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GENERAL INDEX 1 a’a, 138, 146 Aachen (Germany), 399 abiotic, 769 – amelioration, 151 – conditions, 258 – factors, 157 – vectors, 527 Abisko (Sweden), 74 Aboriginals, 317, 679 above-ground net primary production (ANPP), 287–303, 558, 560 – consumption by ﬁre, 291 – consumption by herbivory, 289–291 – effects of climatic ﬂuctuations, 301, 302 – ﬁre effects, 296–300, 302 – gradient, 291 – grazing effects, 296–300, 302 – interannual variability, 300–303 – prediction, 302 Abraham Plain (Washington, USA), 148 abrasion, 83 – by mineral particles, 25 – by snow, 25 acacia (Acacia spp.), 274, 279, 281, 440, 665 acarinans (see also mites), 647 acclimation, see physiological acclimation accretion, 342, 354 acid mine drainage, 369 acid precipitation, 257 acid rain (see also pollution), 2, 52, 64, 67, 93 acid-sulphate soils, 661 acidiﬁcation, 29, 31, 401, 403 actinorhizae (see also nitrogen ﬁxation), 523 adaptation, 42, 341, 344 – of animals, 340 – polymorphism, 349 – seeds, 349 – stress tolerance, 349 – to disturbance, 26 adaptive adjustment, 495 adaptive management, 662, 668, 669, 690, 695, 696, 703 Ad´elie Land (Antarctica), 91 Ad´elie penguins (Pygoscelis adeliae), 55, 62, 91 Adirondack Mountains (eastern USA), 701 aeolian ( = eolian), 769 aeolian activity, 21, 24, 29 aerial photographs, 193 aerobic respiration, 508 Afghanistan, 324 Africa, 14, 15, 17, 56, 123, 125, 128, 308, 312, 318, 324, 425, 432, 436, 444, 556, 664 1

African desert locusts (Schistocerca gregaria), 257 agave (Agave deserti), 311 age–structure diagrams, 181 aggregate stability, 561 aggregate structure, 561, 562, 565 aggregates, 367 agriculture, 2, 4, 5, 8, 15, 22, 59–62, 263, 264, 272, 274–277, 281, 282, 334, 398, 419, 459, 717 – annual cycle, 487 – clearing, 229 – disturbance, 419, 500 – – species richness, 493, 494 – environmental impact of, 723–726, 729–733 – indigenous, 467 – intensiﬁcation, 488, 489 – lands, 493 – landscape, 492 – policy, 689, 691, 700 – scenarios of agricultural development, 724 – slash-and-burn, 472, 473 – systems, 489 – waste, 536 agroecosystems, 14, 15, 398, 487–501 – biodiversity, 488, 489 agroforestry, 276, 470, 471, 479, 691 Agung (Bali, Indonesia), 587 air-borne pyroclastics, 138 air pollutants, 65 air pollution, 369, 399, 401 Ajo (Arizona, USA), 312 Alabama (USA), 618 Alaska Range (USA), 75 Alaska (USA), 5, 26, 28–31, 45, 52, 58, 63, 72–75, 78, 80–85, 90, 140, 167, 168, 170, 189, 533, 587, 588, 593, 596, 597, 601, 615, 618, 701 albatrosses (Diomedea exulans), 59, 91 albedo, 42, 52, 62, 323, 575 – post-ﬁre, 577 Alberta (Canada), 297 alder (Alnus spp.), 551–553, 600–602 Aleutian Islands (Alaska, USA), 52, 90 alfalfa (Medicago sativa), 129, 515 algae, 26, 27, 88, 130, 331, 332, 343, 347, 521, 562, 563, 591 – Chlamydomonas nivalis, 42 – Prasiola crispa, 66 algal-crust development (see also cryptobiotic crust), 536 algal mats, 563 alien species (see also invasive species and exotic species), 154, 317, 320 allelopathy, 153, 598 alligators (Alligator spp.), 332, 337, 340, 751 allocation, 617, 624

Page numbers in boldface type refer to the Glossary.

833

834 allogenesis (see also autogenesis), 30, 597 – change, 28, 31 – control, 30 – disturbance, 589, 595–598 – factors, 28, 29 – processes, 30, 31, 585, 586 allogenic, 769 alluvial, 125 alpaca (Lama pacos), 60 alpine, 41, 56, 78, 416, 620 alpine environments, 18 alpine pennycress (Thlaspi caerulescens), 374 alpine vegetation, 30 Alps (Austria), 18, 26, 28, 30, 50, 55, 57–59, 66, 75, 88, 533, 591 alternative stable states, 675, 678 alumina, 505 aluminium, 365, 588 amabilis ﬁr (Abies amabilis) (see also conifers), 454 Amarillo (Texas, USA), 298 Amazon, 125, 472, 475–477, 666, 691 – Brazil, 667 – rainforests, 314 – Venezuela, 242 Amazon River, 230 Amazonas (Brazil), 231 Amchitka Island (Alaska, USA), 90 amelioration, 157 American – alligator (Alligator mississippiensis), 340 – beech (Fagus grandifolia), 552 – Cajuns, 334 – chestnut (Castanea dentata), 455 – elm (Ulmus americana), 455 American Samoa, 638 Ames (Iowa, USA), 127 amino acids, 42 ammonia volatilization, 554 ammoniﬁcation, 523 ammonium (see also nitrogen), 545, 551, 553, 565 – ﬁxation, 554 – ﬁxation in clays, 546 – volatilization, 546 amphibians, 43, 407, 636, 638, 645 amphipods, 68, 332, 338, 343–347 amplitude, 678 Anak Krakatau (Indonesia), 147–149 analysis of variance, 623 Andaluc´ıa (Spain), 294 Andean Cordillera, 272 Andes, 44, 53, 56, 58, 60, 279, 471, 480 animals, 26, 27, 153, 751 – agents of disturbance, 307 – distribution, 558 – ecological engineering, 717 – husbandry, 59–62, 291 – infestations, 22 – movement, 529 – site effects, 580, 581 annelid worms (see also earthworms), 87, 130, 332, 337, 347

GENERAL INDEX annual grasses, 418, 425, 440, 443 annual rainfall, 294, 295 annuals, 350, 527 Antarctic – fur seals, 91 – skuas (Catharacta antarctica), 91 Antarctic Convergence, 39 Antarctic International World Park, 69 Antarctic Peninsula, 45, 76, 77, 85, 88 Antarctic Treaty, 58 Antarctica, 14, 15, 17, 18, 26, 27, 30, 31, 39, 41, 44, 45, 48, 50, 52–55, 58, 59, 62, 64–69, 74, 77, 80, 87, 91, 93, 94, 139, 438, 442, 529 antelopes, 60 anthropogenic, 769 antibiotic-producing fungi, 530 antifreeze mechanisms, 43 ants, 188, 213, 281, 350, 531, 593, 639, 645, 647–649, 759 Anvers Island (Antarctica), 68, 88 Anza-Borrego (California, USA), 536 aphids, 257 apical dominance, 579 apiculture, 673 Appalachian Mountains (USA), 51, 190, 205, 206, 208, 262, 377 aquaculture, 334 aquatic ecosystems, 427 aquatic systems, restoration, 682, 683 arable land, 1 Arcadian myth, 663 archeophytes, 405 architecture, 715 Arctic, 26, 39, 45, 50, 52, 54–60, 62, 63, 65–67, 69, 73, 74, 79, 80, 82, 83, 87, 88, 90, 93, 314 – ground squirrels (Spermophilus parryii), 55, 56 – polar front, 39 – salt marshes, 335 Arctic Circle, 57, 166, 177 Arctic Coastal Plain, 46, 72, 78, 80–82, 85 Arctic Ocean, 93 Ardley Island (South Shetland Islands, Antarctica), 48 Ardour River, 427 areal extent of disturbance, 354 Argentina, 188, 198, 204, 207, 208, 291, 311, 418, 420, 734, 756 Argentine ants (Iridomyrmex humilis), 281 Argentinian Primavera Station (Antarctica), 77 arid, 308 arid grasslands, 320 arid regions (see also deserts), 536 Arizona (USA), 2, 310, 319, 321, 416, 418, 423, 425, 428, 429, 432, 441, 574, 579 Arkansas (USA), 196 armadillos, 498, 499 arrested succession, 769 arsenic, 369 arthropods, 130, 153, 523 Arthur Harbor (Antarctica), 45 ascospores, 530 asexual reproduction, 43 ash-fall, 144

GENERAL INDEX ash (see also volcanoes), 137, 138, 142, 143, 575 ashing, 575 Asia, 5, 14, 15, 17, 50, 56, 64, 125, 188, 308, 309, 312, 455 Asiatic dayﬂower (Murdannia keisak), 348 aspen (Populus tremula) (see also cottonwood), 166–168 Assateague Island (Maryland, USA), 351 Assyrians, 2 asymbiotic bacteria, 521 Atacama Desert (Chile, Peru) (see also deserts), 271, 308 Atchafalaya Bay (Louisiana, USA), 356 Atlanta (Georgia, USA), 385 Atlantic Ocean, 128, 137, 164, 224 Atlantic white cedar (Chamaecyparis thyoides), 340, 350 atmospheric deposition, 548, 554 atmospheric pollutants (see also pollution), 257 Atqasuk (Alaska, USA), 46, 72, 74 Auckland Island (Subantarctic), 442 Aurora (New York, USA), 127 Austerdalsbreen (Norway), 28 Australasia, 14, 15 Australia, 14, 188, 225, 255, 256, 263–266, 271–273, 281, 282, 288, 308, 312, 317, 322–324, 334, 342, 369, 416–419, 421, 423–425, 429, 431–433, 436, 438–441, 443, 573, 594, 620, 639, 642, 646, 666, 668, 679, 681, 734, 759 – forest ﬁres, 572 – forest regeneration, 572 – grassland, 574 – human history, 274 – shrublands, 573 – southwestern, 281, 282 – Western Australia, 380, 676, 677, 680 – wet forests, 241 Australian Alps, 701 Australian Army, 393 Austria, 22, 59, 533 autogenesis (see also allogenesis), 30, 597 autogenic, 769 autogenic control, 30 autogenic factors, 29 autogenic processes, 31, 585 avalanches (see also landslides), 22, 27, 29, 30, 49, 51, 59, 75, 137–139, 161, 260 Azores, 137

background canopy disturbance, 223, 225–235, 241, 246 – sources of variation, 226–228 bacteria, 600 bacteria (see also microorganisms), 26, 27, 130, 258, 406, 522, 562, 593, 594 – types of, 523, 525 bacterial-feeding nematodes, 535 bacterial-feeding protozoans, 535 badgers (Taxidea taxus), 617, 645 Baja California (Mexico), 165, 189, 308, 532 Bakersﬁeld (California, USA), 312 bald cypress (Taxodium distichum), 332, 341, 354 baleen whales, 59, 91 Bali (Indonesia), 587

835 balsam ﬁr (Abies spp.) (see also conifers), 163, 167, 168, 170–172, 455, 548, 580, 750 Baltic Sea, 683 bamboo, 207, 208, 214 banana poka (Passiﬂora mollisima), 443 Bangladesh, 123, 334, 470, 667 banner-tailed kangaroo rat (Dipodomys spectabilis), 645 banteng (Bos javanicus), 443 bark, 536 bark beetles, 201, 254, 258–262, 443, 455, 457, 580, 636 – Dendroctonus spp., 259, 261, 751 – Ips spp., 259 – Scolytus spp., 259, 751 barley (Hordeum vulgare), 59, 515 barnacles, 350 barrel cacti (Ferocactus spp.), 311 barrier islands, 391 Barro Colorado Island (Panama), 225, 230, 233, 234, 255 Barrow (Alaska, USA), 74 Barstow (California, USA), 312 basalt, 587 bats, 636–638, 642, 643, 646, 647 – seed dispersal, 143 bauxite, 369, 370 Bavaria (Germany), 213, 392 bay checkerspot butterﬂy (Euphydryas editha bayensis), 718 Bay of Bengal, 334 beans (Phaseolus vulgaris), 59 bearded vulture (Gypaetus barbatus), 58 bears, 30 beaver dams, 169 beavers (Castor spp.), 169, 170, 334, 337, 432, 577, 645, 646, 681, 750, 751 – effects on sites, 580 Beddgelert (Wales, UK), 548–550 bee orchids, 371 beech (Fagus spp.), 205, 207, 262, 548 beef, 129 Beer Sheva (Israel), 295 beetles, 264–266, 318, 455, 532, 593, 646, 647 – dermestid, 647 – detritivorous, 316 – Epilachna vigintioctopunctata, 475 behaviour patterns, 26 Belize, 341, 477, 478, 480 Ben Ohau Range (New Zealand), 30 Benelux countries, 385 benthos, 331, 337, 347, 348 Bergsetbreen (Norway), 21 Bering Sea, 58, 84 Berlin (Germany), 399–401, 403–409 Bermuda grass (Cynodon dactylon), 432, 441 big cordgrass (Spartina cynosuroides), 353 big sagebrush (Artemisia tridentata), 317 bighorn sheep (Ovis canadensis), 58 Bingham (Utah, USA), 368 bio-economic pressure (BEP), 726, 728–730, 736, 738, 739 biochemical defenses of plants, 254 biocides, 757

836 biodiversity (see also diversity), 14, 15, 92, 123, 125, 126, 128–130, 173, 174, 467, 469, 471, 473, 477–479, 481, 489, 500, 529, 698, 703, 713, 748, 756–758, 769 – effect of intensiﬁcation of agricultural activities, 488, 489 – in weed succession, 500 biogeochemical cycling, 261, 262, 266 biological impact, 500 biological invasion, 769 biological legacy, 586, 769 biomass, 333, 336, 341, 343, 345, 351, 352, 419, 453, 457, 461, 462, 487, 545, 547–551, 556–559, 565, 769 – accumulation, 291, 488 – ﬁre, 343, 344, 353 – grazing, 344, 353 – grazing effects on, 337 – nitrogen content, 292–294 – removal, 343, 345, 351–353 biome, 769 biophysical constraints, 725, 726, 734, 735, 738–740 bioregional, 701 bioremediation, 530, 675 Biosphere 2, 660 Biosphere reserve, 675 biota, 14, 343 biotechnology, 665 biotic – control, 30 – interactions, 151 – processes, 18 – stress, 496 – vectors, 527 biotic–abiotic interactions, 487 bioturbation, 312, 335, 337, 338, 343, 345 birch (Betula spp.), 55, 62, 83, 166, 168, 170–173, 205, 552, 579 Bird Island (South Georgia), 91 birds, 42, 43, 51, 54, 68, 83, 89, 130, 262, 264, 276, 282, 319, 399, 407, 473, 474, 498, 593, 599, 636, 639, 642, 645, 714 – dispersal, 680 – seed dispersal, 143 Bisley Watersheds (Puerto Rico, USA) (see also Luquillo Experimental Forest), 637 bison (Bison bison), 350, 579, 751 black-crust phenomenon, 26 Black Death, 2 black locust tree (Robinia pseudoacacia), 378, 406 black needlerush (Juncus roemerianus), 353 blackbrowed albatross (Diomedea melanophris), 59 blackbrush (Coleogyne ramosissima), 562 blackbutt (Eucalyptus patens), 381 blading, 80 blasted, 148 block cutting, 547 block lava, 138 blocks, 138 blood pressure, 44 blow downs, 188, 223, 228, 229, 235 blue grama (Bouteloua gracilis), 557, 648 blue-green algae (see also cyanobacteria), 331, 769 blue moor grass (Sesleria albicans), 371

GENERAL INDEX blue stain fungus (Leptographium engelmannii), 751 Blythe (California, USA), 312 bogs, 175, 331–334, 336, 345, 355, 403, 769 Bohemia (Czechoslovakia), 533 bole snapping, 231 Bolinas Lagoon (California, USA), 356 Bolivia, 56, 63, 468, 470 bomb craters, 385 boreal, 769 boreal forests, 161–163, 386, 416, 421, 423 – circumpolar distribution, 161 – description of, 163, 164 boreal regions, 335 boreal vegetation, 30 borers, 254, 580 Borneo, 225, 226, 229, 230, 232, 233, 473, 758 Boston (USA), 404 bottomland hardwood forests, 336 boulder-cored frost boil, 24 boundaries, 462 Bøverbreen (Jotunheimen, Norway), 19 bowhead whales (Balaena mysticetus), 58, 90 brackish marshes (see also wetlands), 332, 345, 349–353, 355, 356 Bradford (Florida, USA), 552 braided channels, 21 Braunschweig (Niedersachsen, Germany), 405 Brazil, 1, 5, 123, 230, 232, 467, 469, 470, 476, 477, 479, 690 breakdown of traditional resource use, 663, 664, 667 breccia, 138 bristlecone pines, 592 British Columbia (Canada), 25, 30, 189, 204, 418, 454, 618, 625 British heathlands, 573 brittle bush (Encelia farinosa), 324 broadleaf arrowhead (Sagittaria latifolia), 336 broiler fowls, 129 brome (Bromus spp.), 422 broom, Scotch (Cytisus scoparius), 422 browsing, 83, 169, 208, 282 bryophytes, 27, 78, 79, 84, 85, 88, 90, 173, 210, 577, 591, 592 Buenos Aires (Argentina), 294, 295, 489, 495 buffel grass (Cenchrus ciliaris = Pennisetum ciliare), 317 buffer zones, 675, 682 bullich (Eucalyptus megacarpa), 380 burial, 24, 335, 336, 339, 341, 343, 345, 356 Burkit National Park, 691 burros (Equus asinus), 324 burrowing, 320, 324 burrowing animals, 55 burrowing owl (Speotyto cunicularia), 645 burrowing petrels, 91 burrows, 53, 54, 93 Bushmen, 280 butterﬂies, 404, 474, 718 C/N ratios, 550, 552, 557, 558 C3 plants, 322, 493 C4 plants, 82, 322, 493 cacao (Theobroma cacao), 470

GENERAL INDEX cacti, 310, 325 cadmium, 365, 369, 403, 404 calcium, 262, 403, 547, 548, 587 California (USA), 67, 189, 271–274, 277–279, 282, 294, 297, 316, 340, 356, 415–419, 421, 423, 425, 426, 429, 438, 439, 456, 457, 459, 617, 736 – human history, 274 Callahan County (Texas, USA), 298 camels (Camelus dromedarius), 681 Cameroon, 255, 474 Canada, 25, 28, 29, 31, 74, 83, 88, 90, 127, 164, 167–169, 174, 189, 211, 369, 376, 388, 416, 421, 423, 434, 444, 456, 457, 580, 695, 734 Canadian Arctic, 174 Canadian Forces Base Shilo, 392 Canadian Rocky Mountains, 533 Canadian Shield, 576 Canary Islands, 147 Canberra (Australia), 620 cankerworm (Alsophila pometaria), 552 canopy, 461, 462 – disturbance, 224 – gaps, 419, 454 – height, 246 – recovery, 230 – scavenging, 548, 549, 553, 554 – structure, 461 canopy–atmosphere–soil interactions, 261–263, 266 Canterbury (New Zealand), 188 Canyonlands National Park (Utah, USA), 320 cap, 535 Cape Region (South Africa), 271–273, 280, 281 carbohydrate exudation, 553 carbon, 262, 594 – inputs, 558 – labile, 557 carbon allocation, 523 carbon dioxide, 2, 66, 67, 87–89, 132, 307, 321, 322, 401 carbon/nitrogen ratios, 524 Caribbean, 2, 3, 124, 224, 225, 230, 241, 257, 472, 480, 588, 635 caribou (Rangifer spp.), 44, 46, 54, 55, 58, 60 carnivore, 769 Carolina bays (southeastern USA), 333 carotenoids, 42 carousel model, 154 carp (Cyprinus spp.), 432 Carpathian Mountains (central Europe), 57 carrion, 593 carrying capacity, 1, 769 Cascade Mountains (northwestern USA), 454 Cascadia, 701 cassava ( = manioc: Manihot spp.), 124, 467 catastrophe theory, 661 catastrophic disturbance, 224 catastrophic wind, 196, 223–246 catchment, 716 caterpillars, 170, 262 catﬁsh, 432 cation exchange capacity, 83, 506

837 Catlin silt loam, 127 cats, 69, 282, 636, 681 cattails (Typha spp.), 336, 340, 434 cattle, 44, 60, 90, 124, 264, 280, 475, 476, 646, 647, 649 Cedar Creek Natural History Area (Minnesota, USA), 620 Cedar River (Washington, USA), 551 cell walls, 524 Central America, 2, 341 Central Asia, 308, 311, 314 Central Europe, 398, 401, 403–405, 408 centrarchids, 432 chamois (Rupicapra rupicapra), 30 Champaign (Illinois, USA), 127 chance, 149 changes of socio-economic structure, 726–729 channel meandering, 334, 354 chaparral, 271, 274, 278, 279, 416, 425, 426, 429, 439, 573, 575, 576, 617 chaparral–oak woodlands, 386 charcoal, 276, 279 cheatgrass (Bromus tectorum), 417, 425 chemical remediation, 511 chemical “time-bomb”, 661 chemical weathering, 151 Chernobyl (Ukraine), 64, 527, 532, 666 Chesapeake Bay (USA), 683, 695 chestnut blight (Endothiella parasitica), 455 chestnut (Castanea sativa), 276 Chesuncook (Maine, USA), 552 chewers (see also herbivory), 254 Chiba (Japan), 405 Chicago (Illinois, USA), 2 Chihuahua (Mexico), 312 Chihuahuan Desert (Mexico, USA) (see also deserts), 287, 319, 322, 323 Chile, 147, 188, 204, 207, 211, 271–274, 279, 280, 282, 323, 366, 676, 679, 693, 716 – ecosystems, 676 – human history, 274 China, 2, 44, 56, 60, 67, 82, 123, 128, 188, 308, 316, 334, 366, 456, 663, 664 chinchilla (Tagetes minuta), 495 Chinese tallow tree (Sapium sebiferum), 425 chinstrap penguins (Pygoscelis antarctica), 91 chlamydospores, 530 chlorophyll, 563 chromium, 369 chronosequences (see also succession), 17, 175, 600, 769 chrysomelid beetles, 318 Chubut (Argentina), 294 Chukotski Peninsula (Russia), 86 cicadas, 641 cirque glacier, 20 civil unrest, 664 Clarina (Iowa, USA), 127 classical elements, 3–5, 14, 15 classiﬁcation (of wetlands), see wetlands clay, 505 – dispersion, 561

838 clay (cont’d) – mineralogy, 505 – soils, 278 clay-pot irrigation, 536 Clean Water Act (see also environmental policy), 391 clear-cutting, 174, 398, 460–462, 480, 546–550, 591, 614, 618, 622, 769 Clemson (South Carolina, USA), 552 climate, 19, 25, 29, 31, 260, 499, 717 climate change (see also global climate change), 30, 66, 67, 87–89, 174, 175, 307, 309, 317 – effects on vegetation, 321–323 – individualistic responses, 175 climatic ﬂuctuations, 288, 302, 303 – effects on above-ground net primary production (ANPP), 300–302 – grasslands and savannas, 293 climax (see also succession), 28, 152 – collapse, 31 – species, 154 – vegetation, 319 clipping, 83 clonal growth, 406 clonal plants, 307, 310, 344 clonal species, 344, 348, 350 cloud forests, 231 cloves (Syzygium aromaticum), 691 Coachella Valley (California), 278 coal, 63, 365–367, 369, 370 coal strip-mining (see also strip mining), 376–378 – restoration, 378 – – Surface Mining Control and Reclamation Act, 377, 378 – – temporal paradox, 378 coarse woody debris, 208, 210–213, 245, 711 coast silver oak (Brachylaena discolor), 380 coastal – grassland, 421 – habitats, 416 – prairie, 421, 438 – red milkwood (Mimusops caffra), 380 – sage scrub, 272, 416 – shrublands, 421 cobalt, 372 Cockaponset (Connecticut, USA), 552 cockroaches, 639 coffee (Coffea arabica), 470, 474, 691 cold deserts, 45 cold ocean currents, 308 cold region ecosystems, 39–94 cold stress, 40 coleopterans (see also beetles), 647 collard (Brassica oleracea acephala), 130 collared lemmings (Dicrostonyx torquatus), 55 Collembola in limestone quarries, 372 Colombia, 470 colonists, 139, 141 colonization, 26, 79, 140, 142, 143, 147, 153, 526, 587, 589, 593, 680 colonization rates, 147

GENERAL INDEX colonizers, 79, 80, 88 Colorado Desert (California, USA), 431 Colorado Front Range (USA), 51 Colorado River, 445 Colorado (USA), 48, 50, 54, 55, 63, 66, 67, 72, 78, 80, 84, 86, 88, 170, 189, 204, 260, 294, 296, 416, 418, 419, 421, 428, 429, 455, 613 Columbia Glacier (British Columbia, Canada), 25 common bent (Agrostis capillaris), 374 common reed (Phragmites australis), 343 common resources, 667 community – changes, 487, 499 – complexity, 487 – composition, 261 – development, 156 – dynamics, 261, 266, 500 – ecology, 707–710 – effects, 390 – – diversity, 390 – – species replacement, 390 – responses to canopy disturbances, 236–245 – richness, 494, 496 – structure, 301, 497, 499, 500 community-level responses, 236–243 compaction, 472, 511, 561, 588, 589 compensatory growth, 261, 301, 769 – evidence, 578, 579 – theory, 578 competition, 14, 15, 28, 40, 142, 153, 154, 172, 192, 211, 212, 224, 344, 348, 351, 352, 417, 434, 462, 586, 592, 593, 595–600, 602, 611, 616, 626, 748, 751, 752, 769 – competitive advantage, 493 – differences in competitive ability, 493 competition intensity, 615–621, 627 competitive ability, 535, 621–625, 627 competitive effects, 627 competitive inhibition, 595, 601 competitive responses, 627 complementarity in competitive ability, 625 complete tree harvest (CTH), 551 composition, 675 compost, 536 condors (Sarcorampus papa), 53 Confucianism, 2 Congo, 469, 470 conidia, 530 coniferous forests, 211, 577 conifers (see also pines and spruce), 188–190, 532 connectivity, 715 Connell’s intermediate-disturbance hypothesis (see also theory under disturbance), 390 conservation, 2, 32, 673, 674 Conservation Reserve Program, 666 conservation values, 675 constancy and environmental change, 40, 494, 495 consumerism, 664 consumers, 14, 15, 289 consumption (see also herbivory)

GENERAL INDEX – ﬁre, 291 – herbivores, 291 – herbivory, 289, 290, 301, 302 – invertebrates, 289–291 – livestock, 290, 291 – vertebrates, 289–291 contaminated land, 666 contamination of groundwater, 369 continental climate, 308 contingent factors, 149 controlled colonization, 679 controlled ecosystems, 659, 660 controlled species performance, 680 Convention on International Trade in Endangered Species, 482 convergence, 320, 769 – during succession, 30 copepods, 90, 332 copper, 279, 365, 366, 369, 372, 588 copper ﬂower (Becium homblei), 372 coppice, 471 cordgrass (Spartina), 338 C´ordoba (Argentina), 489 core–buffer–matrix models, 675 cork, 276 Corn Belt (USA), 127 corn (Zea mays), 59, 126, 127, 129, 481, 510–512, 515 cornucopians, 724, 738, 740, 769 Cornwall (UK), 366 corridors, 415–417, 444, 497, 498, 682, 769 Costa Rica, 225–227, 230–233, 255, 474, 476, 479, 482, 642 cottontail rabbits (Sylvilagus audubonii), 647 cottonwood (Populus spp.), 531 – Populus balsamifera, 533 – Populus deltoides, 431 – Populus trichocarpa, 600 cougars (Felis concolor), 699 Coweeta (North Carolina, USA), 552 crabeater seals (Lobodon carcinophagus), 59 crabs, 337, 338 cranberry (Vaccinium macrocarpon), 334 Crater Lake (Oregon, USA), 142, 144 Craters of the Moon (Idaho, USA), 141 crayﬁsh, 334, 432 creaming, 470 creation (of wetlands), 333, 334 creosotebush (Larrea tridentata), 310, 312, 317 crickets, 593, 639, 646 critical limit, 661 critical thresholds, 660, 661 crocodiles, 332 crop climatic environment, 499 crop production, 673 crop–weed community, 488 croplands, 499 cropped lands, 491–496 – maize, 491 cropping activities, 492 cropping system, 493 crown burls, 573

839 crown ﬁres, 165 crown size, 229, 234 crows, 407 crustaceans (see also copepods, decapods, isopods, ostracods), 87 cryogenic processes, 21, 25 cryoturbation, 23, 26, 28, 29, 769 cryptobiotic crusts, 320, 535, 536, 555, 562–565, 769 cryptogamic crusts (see also cryptobiotic crusts), 320 cryptogams, 140, 142, 173 cryptostigmatic mites, 532 Cuba, 470 cultivated landscape, 492, 495 cultural activities, 493 cultural resistance to change, 740–742, 744 cumulative effects, 388, 660 cut-bank, 46 cyanobacteria, 27, 42, 140, 331, 353, 521, 562, 563, 565, 769 cycling rate, 554 cyclone scrub, 241 cyclones (see also hurricanes), 225, 469, 635–638 cynipid gall wasps, 640 cypress, see bald cypress cypress swamp, 433 cyprinodontids, 432 Czech Republic, 49, 430 Dahurian larch (Larix dahurica), 163, 167 damage to trees by windstorms, 199–201 dammar, 691 damping, 678 damselﬂies, 641 Dari´en (Panama), 467 DDT (dichloro-diphenyl-trichloroethane), 65 Death Valley National Park (USA) (see also deserts), 324 debris avalanche (see also avalanches and landslides), 145, 154 debris dam, 714 debris ﬂows, 21, 22, 27, 51, 137–140, 145 decapods, 68 Deception Island (Antarctica), 64 deciduous forests, 203, 693 decision making, 725, 735, 740, 741, 743, 744 decomposers, 14, 15, 54, 769 – activity, 496 – as drivers of succession, 488 – organisms, 488 decomposition, 41, 52, 54, 55, 67, 173, 291–293, 302, 343, 352, 353, 523, 556, 559, 560, 593, 594 – soil organic matter, 488 deep-pipe irrigation, 536 deep plowing, 511, 536 deer, 60, 214, 426, 697 – white-tailed deer (Odocoileus virginianus), 208, 751 deer mouse (Peromyscus maniculatus), 640 deﬂation, 21, 23, 29 defoliation, 84, 240, 318, 385, 455, 555, 557, 559, 560, 564 – artiﬁcial, 572, 578, 579 – factors inﬂuencing response, 579, 580 – insects, 579, 580 – measuring response, 579

840 defoliators, 260, 261 deforestation, 2, 123, 169, 272, 275, 276, 281, 282, 321, 469–471, 475, 481, 690 deglaciated substrates, 18, 19 degradation, 674 degraded production lands, 674 degrees of degradation, 500 dehesas, 276 Delta Marsh (Manitoba, Canada), 336 dematerialization of developed economies, 736, 737 demographic pressure, 725, 729–734, 743 demographic transition, 738, 739, 769 dendrochronology, 401, 403 denitriﬁcation, 512, 523, 546, 549, 552–554, 562, 565 density – plants, 613 – soil invertebrates, 527 deposition, 528 – dry, 548 – nitrogen, 551 depositional areas, 514 derecho event, 196 dermestid beetles, 647 desertiﬁcation, 5, 29, 276, 277, 279, 288, 323, 324, 665, 770 deserts, 5, 14, 15, 287, 307–325, 416, 418, 419, 565, 591, 592, 708, 713, 716, 769 – animals, 318, 319 – Atacama Desert (Chili, Peru), 308 – cause of, 308 – Chihuahuan Desert (Mexico, USA), 319, 322, 323 – cold, 309 – deﬁnition of, 307, 308 – environment, 530 – grassland, 418 – Great Basin Desert (USA), 317 – hot, 309 – Kalahari Desert (southern Africa), 308 – Karakum Desert (Turkmenistan), 323 – management, 310 – Mojave Desert (USA), 308, 311, 313, 316, 317 – Namib Desert (Namibia), 308, 311, 316, 324 – Negev Desert (Israel), 312, 313, 319, 323 – pavements, 316 – plants, 307 – polar, 308 – pupﬁsh (Cyprinodon nevadensis), 324 – Sahara Desert (northern Africa), 308, 314, 323, 324 – Sonoran Desert (Mexico, USA), 308, 310, 311, 317, 320, 321, 323, 324 – streams, 708 – succession, 319–321 – Takla Makan Desert (China), 316 – tobacco (Nicotiana glauca), 324 – tortoises (Gopherus agassizii), 318 desiccation, 25 designed disturbance, 679, 680 destabilization during succession, 31 detritivores, 640, 647, 770 detritivorous beetles, 316

GENERAL INDEX Devon Island (Canada), 74 Devon (UK), 366 Dhaulagiri (Nepal), 49 diamicton, 18 diamonds, 63 diatoms, 331 dichloro-diphenyl-trichloroethane (DDT), 65 dieback, 336, 340, 342, 770 differential dispersal, 149 differential impact, 149 diffuse disturbance, 714 digestion, 555, 556, 564 digging, 318 dinoﬂagellates, 331 dipterans, 647 dipterocarps, 229 direct glacial disturbance, 19, 20 direct impacts of munitions, 385 directed blasts, 140 disease outbreaks, 716 diseases, 4, 455 disharmony, 148, 149, 156 Disko Island (Greenland), 74 dispersal, 148, 149, 151, 527, 591–593, 598–602, 680, 681 – ability, 152 – barriers, 152 – birds, 680 – rates, 535 disturbance, 545–549, 552, 553, 770 – abiogenic, 635 – adaptations to, 26 – anthropogenic, 1, 2, 4–6, 8, 10, 12, 22, 56–70, 92–94, 173, 174, 307, 310, 316–318, 323, 324, 333–335, 337, 339–341, 343–345, 350, 352, 354–356, 415–419, 432, 435, 456, 459–464, 467, 469, 471, 475, 481, 496, 497, 620, 659, 673, 693, 758, 759 – aquatic, 3 – areal extent, 338, 354 – atmospheric, 3 – biogenic, 635, 643–650, 750, 751, 754 – biological effects, 640, 641 – biotic, 2–5 – by animals, 710 – causes of, 3 – characterization, 572 – concept of, 707, 708, 714 – corridors, 415–417, 422, 444 – deﬁnitions of, 2, 3, 307, 413–415, 571, 585, 633, 707 – diffuse, 714 – direct effects, 634, 644, 650 – direct glacial, 19, 20 – during succession, 27–31 – ecosystem response to, 752–754 – effects, 307, 471, 478–480 – – above-ground productivity, 295–303 – – cycles, 493 – – grasslands and savannas, 301 – – of animals, 644–647 – – on animals, 26, 27, 389, 390, 473–475, 481

GENERAL INDEX – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – – –

– on ecosystems, 18, 475, 481 – on micro-organisms, 26, 27 – on substrate, 18 – plants, 18, 389, 472, 473, 480 – recovery and restoration, 475–477, 481 – soils, 471, 472 endogenous, 3, 22, 224 exogenous, 3, 22, 224, 708, 711, 712 extent, 3, 4, 633, 634 ﬁre, 753 frequency, 3, 4, 71, 75, 191, 193, 194, 196, 197, 260, 266, 454, 633, 634 glacier-conditioned, 21, 22 glacier-dependent, 20, 21 glacier-independent, 22 hierarchy, 710 history, 229, 571 incorporation of, 712 indirect effects, 26, 634, 644, 650 intensity, 3, 15, 191, 214, 260, 266, 571, 633, 634 interactions, 3, 4, 15, 342, 343, 350, 586, 749 intermediate, 276 intuitive connotations, 709 invasion, 415–445 invertebrate responses, 635–641, 643 magnitude, 3, 191, 214, 340–342, 586 microbial responses, 526, 527 natural, 1–4, 15, 309–323, 332, 333, 335, 356, 419–426, 692 – ﬁre, 316–318 – temperatures, 310, 311 – water, 311–314 – wind, 314–316 of substrate, 23, 24 paraglacial, 18, 21, 22 patterns of, 586–591 physical effects, 640 predictability of, 755, 756 primary (see also primary succession), 4 regime, 22–27, 137, 139, 435, 445, 492, 633, 634, 716–718, 749, 750, 760, 764 – grasslands and savannas, 289 – on glacier forelands, 22–27 – windstorms, 192–199 regularity, 260, 266 response to, 591–595 return interval, 454 scale, 260, 266 – spatial, 39 – temporal, 39 secondary, 32 severity, 3, 4, 15, 191, 260, 266, 571, 586 shrub-steppe, 526 spatial heterogeneity, 750, 751 spread of, 715 stand initiation, 29 theory, 644, 648, 749, 760–764 timing, 571 types of, 3–5, 18, 19, 260, 309, 335–337 vertebrate responses, 636–640, 642, 643

841 – wind, 50, 51, 187–214, 635–638, 753, 757 disturbed lands, 521 divergence, 770 divergence, during succession, 30, 31 diversity and disturbance, 192, 200, 206, 208–211 diversity (see also biodiversity), 28, 162, 258, 275–278, 280, 281, 391, 462, 532, 612, 618, 625 dogs, 636 domes, 138 domestic animals, 59, 636 dominant species, 497 dormancy, 26, 274 Dorrigo National Park (New South Wales, Australia), 256 Douglas-ﬁr (Pseudotsuga menziesii), 198, 203, 261, 454, 457, 551, 552, 704 Douglas-ﬁr tussock moth (Orgyia pseudotsugata), 260, 261 down-bursts, 196, 197 dragonﬂies, 641 dredging, 335 dropseed (Sporobolus spp.), 318, 648, 649 drought (see also deserts), 3–5, 29, 41, 45, 89, 224, 229, 254, 262, 272, 295, 309, 311, 315, 324, 333, 335, 336, 344, 417, 427, 432, 445, 558, 588, 714, 749 dry beans (Phaseolus vulgaris), 515 dry ﬂush, 26 dry forests, 416 dry meadow, 79 dry season, 235 dry tropical forests (see also forests), 226, 227 dryland rice (Oryza sativa), 691 dune false currant (Allophylus natalensis), 380 dunes (see also sand dunes), 41, 46, 51, 73, 75, 278, 421, 436, 587, 589, 591, 593–596, 599, 600 dung, 578 dung beetles, 646, 647 dust, 65, 314 Dust Bowl, 2, 667, 668, 691 dust storms, 316 Dutch elm disease (Ceratocystis ulmi), 262, 455 dwarf forests (see also forests), 241 early-seral species, 532 early-successional, 531 earthquakes, 3, 4, 137, 272, 334, 588, 589, 716 earthstars, 528 earthworks, 512 earthworms (see also annelid worms), 130, 131, 443, 509, 523, 593, 641, 645, 754 East Africa, 137, 138, 287, 289, 290 eastern deciduous forests, 386 eastern hemlock (Tsuga canadensis), 260 eastern white pine (Pinus strobus) (see also pines), 378 ecesis, 28, 586 eco-refugees, 667 ecological footprint, 1 ecological restoration, 683 ecological “time-bomb”, 661 ecology, 467 – history of, 17

842 economic growth, 723–745 ecosystem, 690, 695, 700, 703, 770 – development, 139 – disturbance, 659–661, 663, 665–668, 710 – dynamics, 689, 692, 693, 696 – – models, 693 – effects, 390 – equilibrium, 659 – functions, 488, 675 – health, 676 – management, 386, 393, 394, 690, 692–699, 770 – processes, 254 – recovery, 139 – resilience, 694, 696 – responses, 70–91, 243–245 – – to canopy disturbances, 236–245 – roles, 521 – spatial and temporal patterns, 692 – stability, 488, 696 – valued components (VEC), 694 ectomycorrhizal fungi (see also mycorrhizae), 522 ectothermy, 43 ectozoochory, 593 Ecuador, 125, 126, 473, 691 edge effects, 415–417, 475 Edgeøya (Spitsbergen), 74 edges, 463, 715 Eglin Air Force Base (Florida, USA), 391, 392 Egypt, 323, 365, 663 EIA, see impact assessment under environmental El Chich´on (Mexico), 137, 142, 145, 587 El Ni˜no, 89, 473, 475, 477, 717, 770 El Par´ıcutin (Mexico), 142, 144, 148 El Verde (Puerto Rico, USA) (see also Luquillo Experimental Forest), 637 elasticity, 40, 678 elemental cycling, 523–525 elephant seals (Mirounga leonina), 54, 59, 83 elephants, 337, 475, 751 elevation and wind damage, 200 elk – Cervus canadensis, 153, 290 – Cervus elaphus, 392 elm bark beetles – Hylurgopinus ruﬁpes, 262 – Scolytus multistriatus, 262 eluviation, 504 emigration, 56 emissions, 398, 401 – NOx , 401 – sulfur dioxide, 401 emperor penguins (Aptenodytes forsteri), 89 enchytraeids, 130 Endangered Species Act, 391, 392, 699 endemic species, 140, 467 endogenous causes of disturbance (see also endogenous under disturbance), 22, 224 endosomatic energy, 728 endotherms, 43, 770

GENERAL INDEX endozoochory, 593 energy, 258 – consumption, 1 – efﬁciency, 735, 736 – ﬂow, 560, 710, 718 England (see also Great Britain), 371, 406 enrichment planting, 469, 470, 478 entrainment, 528, 770 Entre R´ıos (Argentina), 489 environmental – change, 257–260 – compliance, 390, 391 – conditions, 257, 258 – degradation, 662, 665 – ﬂuctuations, 41 – gradients, 73, 625 – impact assessment (EIA), 659, 661, 665, 668, 697 – impact of food production, 734 – impact statement (EIS), 696 – loading, 726, 739 – management, 662, 668 – policy, 760 – politics, 2 – processes – – cultural, 18 – – natural, 18 – stress, 146 Environmental Conservation Program, 391 ephemeral ponds, 313 epiphytes, 200, 201, 227, 233–235, 770 epistemological predicament of sustainability, 725 equilibrium, 309, 678 ericoid mycorrhizae (see also mycorrhizae), 524 ericoid roots, 524 eroded areas, 514 erosion, 2, 4, 15, 24, 44, 47, 51, 52, 59, 60, 67, 69, 75, 141, 151, 155, 262, 275–278, 281, 282, 307, 309, 312, 314, 315, 334–337, 339, 342, 345, 354, 392, 462, 472, 475, 480, 511, 528, 561, 565, 588–592, 753, 757 – river-banks, 441 – wind, 591 eruption characteristics, 140 espinal, 279, 676 establishment, 29, 529–531, 591, 592, 601, 602, 617, 624, 770 estuary, 716 ethics, 663 Ethiopia, 56 eucalypts (Eucalyptus spp.), 256, 281, 469, 470, 480 – forests, 281 – woodlands, 421, 424 Euphrates River (Turkey, Syria, Iraq), 2, 663 Eurasia, 62, 163, 164, 170 Europe, 2, 14, 15, 17, 51, 59, 62, 64, 67, 125, 175, 187, 188, 194, 255, 262, 271, 274, 278–280, 308, 350, 352, 372, 374, 388, 397, 398, 403–405, 408, 434, 456, 499, 500, 663, 666, 679, 683, 695, 738 European annuals, 274 European beach grass (Ammophila arenaria), 441 European gypsy moth (Lymantria dispar), 443

GENERAL INDEX European Union (EU), 664 eutrophication, 65, 91, 131, 399, 403, 682, 770 evapotranspiration, 39, 262 – potential, 308 evening primrose (Oenothera spp.), 318 Everglades (Florida, USA), 335, 336, 340, 352, 355, 434, 683 evolutionary processes (see also phenotypic plasticity), 500 excavations, 4, 387 exclosure studies, 318 excreta [faeces], 555, 556, 558–560, 565, 579, 581 excretion, 546, 555 exhumed plant communities, 20 Exit Glacier (Alaska, USA), 26, 28 exogenous causes of disturbance (see also exogenous under disturbance), 224 exosomatic/endosomatic energy, 728, 736 exotherm, 770 exotic species (see also alien species and invasive species), 154, 272, 274, 278, 279, 281, 282, 333, 340, 343, 477, 479–481, 487, 498, 689, 694, 695, 749, 758 – ﬁsh, 282 – mammals, 282 – plants, 282 experiments, 612–614, 618–620, 622–626, 711 extinction, 58, 147, 148, 392, 699 extreme environments, 309 exudates, 562 facilitation, 770 facilitation (see also succession), 79, 143, 152–154, 586, 595–602, 613–616, 620, 626, 627, 752 factor complex, 28 faeces, see excreta Falealupo Rain Forest Reserve (Samoa), 637 fall cankerworm (Alsophila pometaria), 259 feather mosses, 577 Federal Forestry Department (Germany), 392 Federal Land Development Authority (FELDA) (Malaysia), 667 fellﬁelds, 50 Fennoscandia, 66, 162, 166, 167, 169, 171, 172 fens (see also marshes and wetlands), 331, 332 feral, 770 – animals, 443, 445 – herbivores, 681 – predators, 681 ferns, 140, 404–406, 531, 591, 592 ferrets, 282 fertility, 558 fertilization, 401, 499, 549, 554 – experiments, 573, 575 ﬁber nets, 536 ﬁbrous roots, 25 ﬁddler crab (Uca pugnax), 337 ﬁeld age, 617, 619, 624 ﬁeld lark (Alauda arvensis), 398 ﬁeld mouse (Microtus spp.), 398 Fiji, 469, 470 ﬁlter-feeders, 90 Finke River (Australia), 431

843 Finland, 65, 166, 171, 188, 200 Fiordland (New Zealand), 188 ﬁr (Abies spp.) (see also Douglas-ﬁr), 162, 168, 170–172, 205, 261, 580 ﬁr waves, 172, 198, 455 ﬁre, 1, 2, 4, 5, 52, 90, 161, 162, 164–169, 190, 194, 229, 235, 240, 257–262, 266, 271, 272, 275–281, 288, 299, 301, 302, 307, 309, 316–318, 333–338, 340–345, 347, 350–356, 417, 419, 422–426, 434, 435, 439, 440, 442, 445, 455–459, 461, 462, 472, 473, 475, 477, 478, 480, 481, 532, 562, 564, 572–577, 581, 592, 615, 617, 619, 625, 638–750 – effect on nutrient cycles, 574, 575 – effects of absence of ﬁre, 577 – effects on above-ground net primary production (ANPP), 296–300, 302 – erosion, 459 – frequency, 291, 302, 303, 317, 322, 393, 572, 573 – – mechanisms, 291–293 – fuel load, 316 – grasslands, 316, 317 – history, 641 – human use, 571, 573, 574, 576 – intensity, 456 – microclimate after, 573 – nutrient cycles, 577 – nutrient losses, 575 – pest control, 640 – recovery from, 317 – regime, 31, 459 – – in grasslands and savannas, 291 – return interval, 317, 456, 457 – severity, 166 – size distribution, 457 – stand-replacing ﬁres, 456–459 – suppression, 425, 441, 456, 458 – temperatures, 458, 575, 576 – timing, 572, 574 ﬁreweed (Epilobium angustifolium), 170 ﬁrewood, 276 ﬁsh, 53, 68 ﬁshing, 58 Flambeau Tract (Wisconsin, USA), 197 ﬂatwoods, 391 ﬂight missions, 391 ﬂood control, 432 ﬂooded prairie, 420 ﬂooding, 25, 26, 28, 161, 229, 335, 340–342, 351, 511–513, 588 – effects on animals, 336 ﬂooding regime, 331, 333–335, 338, 340, 353 ﬂoodplains, 2, 25, 190, 355, 586, 587, 590–594, 597–601, 770 ﬂoods, 1, 4, 5, 51, 307, 309, 311, 313, 314, 427, 431, 434, 441, 444, 692, 693, 754 ﬂora, 397, 399 Florida Keys (USA), 701 Florida (USA), 123, 128, 189, 195, 196, 335, 336, 340, 341, 351, 352, 354, 355, 391, 420, 433, 434, 439, 441, 618, 682, 683 ﬂoristic composition, 491, 497 ﬂoristic richness, 492 ﬂoristic stability, 487

844 ﬂowering, 26 ﬂowering plants, 404–406 ﬂush, 26 ﬂy orchids (Ophrys insectifera), 371 ﬂycatchers, 642 ﬂying foxes, 636 foliar cover, 561 folivore, 770 folivory, 254, 257, 260, 261, 266 food webs, 390, 522, 711 foot slopes, 514 forage production, 574 forage quality, 292, 293 forbs, 144 foredunes, 532 forest cover, 463 forest damage by wind, 199–201 forest dynamics and wind storm disturbance, 200 forest ﬁre, 572 forest ﬂoor, 456 forest-tundra, 577 forested wetlands, 225 forestry, 4, 8, 22, 174, 462, 469, 673, 690 – policy, 689, 698, 699 forests, 14, 15, 287, 311, 398, 400, 401, 404, 405, 407, 408, 438, 534, 546, 565, 613, 615, 617, 618, 622, 624, 625 – coniferous, 5 – evergreen, 386 – fragmentation, 254, 476 – management, 468–470, 477–482 – nitrate losses, 554 – old-growth, 187, 453, 459, 462 – paleotropical, 226 – post-ﬁre productivity, 572, 573 – regeneration, 586 – structure, 228–230, 253 – subalpine, 204 – temperate, 453–464 – – harvesting, 459–464 – tropical, 467–482 – – disturbance effects, 471–477 – tropical hardwood, 354 – types, 192, 193, 226, 255–259, 261, 262, 266, 453, 615, 621 – vegetation, 398 – windstorm disturbance, 187–214 Formicidae, see ants Fort Carson (Colorado, USA), 385, 389 Fort Knox (Kentucky, USA), 389 Fort Richardson (Alaska, USA), 390 foundations, 511, 512 foxes, 44, 282, 499, 513, 681 fragility, 476 fragmentation, 259, 260, 264, 266, 474–476, 481, 675, 682, 770 France, 189, 194, 276, 423, 427, 430, 573 Franz Josef Glacier (New Zealand), 31 Fraser ﬁr (Abies fraseri) (see also ﬁr), 455 French Guiana, 230, 232, 233 frequency of disturbance, 191, 193, 194, 196, 197 freshwater bass, 432

GENERAL INDEX freshwater marshes (see also wetlands), 332, 336, 342, 345, 349–352, 354, 356 frog bit (Hydrocharis morsusranae), 434 frogs, 432, 636, 638 Front Range, Colorado Rocky Mountains (USA), 48 frost, 21, 23–25, 28, 29, 51, 271, 307, 310, 335, 336, 340 frugivores, 473, 475, 477, 636–638, 642 fruit-eaters, 254 fruit ﬂies, 636 fuelwood, 123, 470 fugitive plant, 714 Fuji-san (Japan), 141–144, 147 function (of ecosystems), 675 functional diversity, 756 functional redundancy, 14, 15, 770 fungal communities, 522 fungal pathogens, 266, 282 – Phytophthora cinnamomi, 263 fungi, 130, 166, 167, 173, 258, 262, 444, 522, 562, 564, 580, 591, 593, 594, 600, 710 – and disturbance, 194 – and wind throw, 199, 201 – and windstorms, 192 – bird-nest fungi, 528 – blue-stain (Leptographium engelmannii), 455 – melanized, 532 – non-melanized, 532 – on coarse woody debris, 211 – root-rot, 455 – saprophytic, 525 – sugar, 532 fungivores, 647 fur seals, 79, 91, 93, 94 – Arctocephalus gazella, 54 – Callorhinus ursinus, 90 fynbos, 271, 272, 280, 281, 424, 425, 430, 439–442, 770 gabbro, 52 Gal´apagos Islands (Ecuador), 142 gall-formers, 254 gallery forests, 271 Gamage Point (Antarctica), 88 gamma diversity, 500 gaps, 188–190, 195, 197, 198, 202–208, 214, 615, 714, 716 – background, 223–246 – belowground, 189, 190, 211, 212 – canopy, 454 – deﬁnition, 202 – dynamics, 585 – experimental, 189, 198, 202, 208, 211 – formation rate, 225, 227, 233–235 – light, 190, 191 – old-ﬁeld, 710 – partitioning, 206 – roots, 212 – size, 202–206, 225, 227, 233, 234, 242, 246 – species, 472, 477 garden, 617, 624 garrigue, 271, 279, 573

GENERAL INDEX geese (see also lesser snow geese), 83, 337, 751 geliﬂuction, 24 gene ﬂow, 495 genetic diversity, 756 genetic erosion, 472 genetic hardpan, 510 genetic variability, 199 gentoo penguins (Pygoscelis papua), 53 geoecological landscape, 17 geoecology, 18 geographical information systems (GIS), 662 geographical landscape, 18 geomorphological processes, 23, 29, 309 geophytes, 140 Georgia (USA), 126, 334, 336, 385, 618, 700 geothermal, 43 Germany, 59, 188, 194, 200, 365, 392, 399, 401, 404–407, 433 germination, 26, 274, 311, 594, 602, 617, 624 germination inhibitors, 573 Ghana, 469 giant petrels (Macronectes giganteus), 54 giant tortoises, 593 Gila Bend (Arizona, USA), 312 girdling, 552–554 glacial disturbance, 19 glacial forelands, 587–589, 591, 592, 594, 596–601, 770 glacial meltwater, 22 glacial moraines, 586, 592, 593, 598, 770 Glacier Bay (Alaska, USA), 30, 31, 596, 600–602 Glacier National Park/Waterton Biosphere Reserves (USA, Canada), 701 glaciers, 4, 18, 47, 49, 50, 53, 66, 67, 71, 75, 83, 93, 333, 354, 588 – advance, 19, 21 – – effects of, 25 – bulldozing, 21 – climate, 19, 20 – disturbance, 19–22 – foreland, 18, 20 – ice, buried, 20 – moraines, 19, 20, 25, 28, 29, 588 – propagule bank, 27 – retreat, 18, 22 – wind, 19, 20 glacio-ﬂuvial processes, 20, 23, 25, 28, 29 Glenamoy (Ireland), 74 global climate change, 2, 174, 321–323, 475, 481, 588, 662, 668 goats (Capra spp.), 60, 279, 681 – feral, 443 gold, 277, 365, 366, 369, 370 Golmud (Tibet), 61 gopher mounds, 645 gophers (Thomomys spp.), 153, 422, 442, 593 gradient, 716 gradual ecosystem disturbance, 661 grain (ecological), 708 grain sorghum (Sorghum bicolor), 490 graminoids, 355 grass–scrub, 694

845 grasses, 25, 26, 82, 86, 140, 144, 264, 592, 597, 619, 623, 626, 760 grasshoppers, 55, 318, 639–641, 645, 649 grasslands, 14, 15, 264, 278, 287–303, 308, 311, 386, 392, 406–408, 421, 424, 438, 440, 442, 443, 445, 489, 493, 534, 556–559, 613, 615, 617, 619, 624, 625, 714 – Africa, 556–558 – annual, 297 – climate, 287, 288 – community, 500 – conversion, 417 – deﬁnition, 287 – distribution, 287 – disturbance, 288 – disturbance effects on above-ground net primary production (ANPP), 295–303 – East African, 290 – ecotones, 287 – Mediterranean, 294, 298 – model of disturbance effects, 302 – North America, 556–558 – post-ﬁre productivity, 573–575 – precipitation gradient, 287–303 – semiarid, 294 – short, 558 – soil, 288 – species richness and extintion, 492, 497 – structure, 287, 288 – subhumid, 294, 295 – tall, 558 – temperate, 5, 289 – tropical, 5, 289 Gratton Dale (Derbyshire, UK), 375 gravel, 536 gray beech (Fagus grandifolia), 442 gray whales (Eschrictius robustus), 90 Graz (Austria), 407 grazers, 92, 353 grazing, 29, 30, 60, 83, 84, 92, 169, 272, 275–279, 281, 282, 288, 291, 294, 298, 299, 301–303, 320, 333–338, 340–345, 350–355, 414, 417–419, 432, 442, 444, 445, 480, 497, 499, 522, 546, 547, 555–565, 578–580, 617, 644, 677, 753 – aggregate stability, 560 – ecosystem-level processes, 556 – effects, 546, 547 – effects on above-ground net primary production (ANPP), 296–300, 302 – inﬁltration, 560 – intensity, 578 – interception of precipitation, 560 – land, 418 – regime – – grasslands and savannas, 289 – severity, 578 – sheep, 574 – surface detention, 560 Great Basin Desert (USA), 311, 317, 322, 416–418, 420–423, 428, 429, 528

846 Great Britain, 130, 189, 365, 403, 421, 427, 429, 430, 433, 434, 460, 675 Great Dismal Swamp (Virginia/North Carolina, USA), 336 Great Lakes (North America), 164, 683 Great Plains (North America), 292, 295, 556, 560, 576 – Texas high plains, 298 Greece, 1, 137, 272, 275, 365, 663 Greenland, 18, 63, 64, 67, 74, 92 grizzly bears (Ursus horribilis), 695 groove-toothed rat (Otomys orestes), 56 ground-troop maneuvers, 391 groundwater, 369, 390, 401–403, 770 groundwater recharge, 21 growth, 261, 617, 619, 623 – stimulation, 25 – suppression, 25 growth rates, 461, 592, 593, 598 Guadeloupe, 476 Guam (USA), 636 guanicoe (Lama guanicoe), 60 guano, 54, 646 Guaymas (Sonora, Mexico), 312 Guelph loam soil, 127 Guelph (Ontario, Canada), 127 Gulf of Mexico, 343, 349 gullies, 514 gulls (Larus spp.), 70 gully erosion, 22 Gyantse (Tibet), 47 gypsy moths, 444, 455, 552 – Lymantria dispar, 260, 262 habitat, 682 habitat fragmentation, 390 habitat heterogeneity, 141 Hadley Cells, 308 Hardangervidda (Norway), 79 hardpans, 510 hardwood forests, 438 hares, 55, 58, 398, 593 harvest mouse (Reithrodontomys megalotis), 640 harvester ants, 531 harvesting, 58, 59, 257, 546, 547, 551, 553, 565 – complete tree harvest (CTH), 551 – forests, 459–464, 545–553 – low-impact, 480 – sawlog (SAW), 547–549 – whole-tree (WTH), 545, 547, 549–552 Hawaii (USA), 1, 123, 128, 137, 138, 140, 142, 147, 149, 152, 154, 264, 423, 435, 436, 438, 439, 442, 443, 587, 588, 611 Hawaii Volcanoes National Park (Hawaii, USA), 440 Heard Island (Antarctica), 30, 66, 88 heath vegetation, 25 heather (Calluna vulgaris), 375 heathland, 271, 421, 423, 575, 576 heavy metals, 335, 337, 388, 390, 401, 404, 407, 524 hedgehogs (Erinaceus europaeus), 407 helicopters, 70 Helm Glacier (British Columbia, Canada), 30

GENERAL INDEX hemeroby, 5, 408, 409 hemicryptophytes, 140 hemipterans, 647 hemlock, 454, 456, 600 hemlock woolly adelgid (Adelges tsugae), 260, 262 herbaceous culms, 557 herbaceous vegetation, 401, 403 herbicides, 337, 399, 546, 550, 554 herbivores, 43, 54, 55, 88, 289–291, 443, 444, 546, 547, 560, 770 – biomass, 289–291 – consumption, 289 – effects of exclusion, 578 – outbreaks, 253, 254, 259–264, 266 herbivory, 1, 3, 4, 88, 161, 162, 169–172, 174, 192, 208, 224, 253–266, 311, 318, 335, 355, 422, 426, 557, 577–580, 593, 597–599, 602, 681, 749 – below-ground, 579 – measurement, 253–257 – patterns, 254–257 – responses to environmental change, 257–260 – types, 254, 257, 260, 262, 266 – variability, 257 herbs, 618, 625 heterogeneity, 70, 332–334, 336, 339, 349–351, 356, 675, 708, 711, 713, 714, 716, 719, 748 heteromyid rodents, 645 heterotrophic succession, 596, 600 heterotrophs, 545, 554, 565 hibernation, 43, 44, 93 hickory (Carya spp.), 377 hierarchy, 709, 710, 712, 714, 718 high-alpine vegetation, 32 High Arctic, 26 high-grading, 470 high-pressure zones, 308 high temperatures, 310 highly invasible, 492 Himalaya, 40, 44, 49, 61, 83, 480 Hohenfels Combat Maneuver Training Center (Germany), 392 Hohokam, 2 holism, 663 Holocene, 161, 174, 273, 280 homopterans, 641 honey mesquite (Prosopis glandulosa), 318 honeybees (Apis mellifera), 324 Hong Kong (China), 341 Honshu (Japan), 145 Hopen Island (Svalbard Archipelago), 88 Horn Island (Mississippi, USA), 443 horses (Equus caballus), 83, 681 house mouse (Mus musculus), 398 house sparrow (Passer domesticus), 398 Hubbard Brook (New Hampshire, USA), 546–548, 552, 711 Hudson Bay (Canada), 348, 350, 577, 579 human appropriation of net primary productivity, 723 humans (see also anthropogenic under disturbance), 1, 593, 651, 708, 749, 754, 755, 758–760 – health, 123 – history, 273

GENERAL INDEX – in wetlands, 331, 332, 334, 337 – modiﬁcation, 674 – population growth, 1, 700 – responses to disturbance, 659–669 humid prairie, 420 hummingbirds, 474 hummock grasses, 574 humus, 574, 770 – ﬁre effects, 573, 575–577 Hungary, 130 hunter–gatherers, 277, 279, 280, 323 hunting, 58, 90, 473 Hurakan, 1 hurricane scrubs, 230 hurricanes, 1–5, 15, 188–190, 194, 195, 200, 201, 203, 205, 210, 223–226, 230, 232, 235, 257, 334–336, 338–341, 348, 352, 354–356, 435, 444, 469, 476, 478, 480, 481, 635, 636, 693, 713, 716, 717, 749, 770 – Hurricane Andrew, 336, 339, 341, 349, 352, 354, 356 – Hurricane Camille, 426 – Hurricane Gilbert, 239, 241, 244, 245, 636 – Hurricane Hugo, 2, 200, 201, 236, 239, 243–245, 257, 259, 343, 435, 454, 635–642, 762, 763 – Hurricane Joan, 636 – mangrove wetlands, 333 – mangroves, 348 – marshes, 356 – salinity, 335, 340, 356 – salt water intrusion, 201 – sediment deposition, 336 – tropical cyclones, 225 hydraulic seeding, 373 hydric, 770 hydrocarbons, 401 hydrogen, 401 hydrologic cycle, 555, 560 – grazing, 560–562 hydrological changes, 682 hydrological imbalances, 682 hydrological regime, 683 hydrology, 331–333, 335–337, 343, 348, 352, 427, 434, 441, 560, 675, 678 – effects, 389 hydrophobic soil, 459 hydrophobic substances, 576 hydroseeding, 155 hygrophilous, 770 hyper-arid, 308 hyphae, 522 hyporhoeic zone, 715 hypothermia, 43 hypothesis-testing, 31 hypoxia, 44, 770 hyrax (Procavia spp.), 56 Hyrnebreen (Svalbard, Norway), 28 hysteresis, 678 Iberian Peninsula, 276 ibex (Capra ibex), 58

847 ice, 5, 45–50 Ice Ages, 58 ice cap, 18 ice-cored moraines, 28 ice crystals, 43, 45 ice lenses, 23 ice sheet, 18 Iceland, 52, 67, 83, 87, 92, 137, 142, 148, 588 ictalurids, 432 Idaho (USA), 140, 418, 423, 425, 440 Iles Crozet (Antarctica), 69 Illinois (USA), 189, 205, 208, 297, 298, 614, 615, 617, 619, 621, 640, 642 illuviation, 504 immobilization, 525, 558, 560, 770 – of nutrients, 31 impact – spatial, 387 – temporal, 387, 388 impact assessment, 770 imprinting, 536 Inca people, 663 incentives, 690, 691, 699, 700 incorporation of disturbance, 712 increasing crop, 489 India, 124, 188, 202, 206, 296, 334, 470, 472, 663 Indian Deccan plateau, 137 Indian Ocean, 137, 636 Indiana (USA), 126, 189, 196, 415, 416, 459 indicator species, 474 indigenous people, 277, 456, 467 indirect effects, 599 individual-level experiments, 612 Indonesia, 1, 137, 142, 148, 263, 470, 474, 475, 479, 587, 588, 667 industrial development, 673 industrial wood, 470 inﬁltration, 536, 558, 560–562, 565 inﬁltration rates, 562 inﬂux variable, 29 ingestion, 555 ingrowth, 242 inhibition, 79, 152, 153, 597 inland pampas, 420 inorganic nitrogen, 535, 550 – soil solution, 550 insectivores, 636, 642 insects, 54, 173, 253–266, 282, 335, 341, 355, 407, 455, 523, 585, 636, 637, 639, 645, 750, 751 – defoliation, 553, 554, 579, 580 – functional groups, 254 – herbivory, 336, 341 – outbreaks, 164, 318, 552, 716 – population eruptions, 580 – responses to environmental change, 257–260 institutional changes, 683 insular tropical forests, 420 intensity of disturbance, 191, 214 intensity of plant interactions, 620

848 interactions of disturbances, 14, 748 interfacial ice, 30 intermediate disturbance hypothesis (IDH), 276, 488, 757 Intermountain West (USA), 418, 556, 557 internal stand disturbances, 172 International Model Forests, 695 introduced grasses, 611 introduced livestock, 674 introduced species, see invasive species invasion, 4, 68, 69, 415–445, 477, 487, 488, 497–500, 611, 625, 677, 713 – mechanisms, 426 – undisturbed vegetation, 415 – wetlands, 427–434 invasive species (see also alien species, exotic species and weeds), 14, 15, 94, 674, 748, 758, 759 invertebrates, 42, 43, 45, 54, 68, 87, 166, 173, 593 Inyo (California, USA), 312 Iowa loess soil, 127 Iowa (USA), 125, 297, 298, 351, 617, 625 Iran, 324 Iraq, 334 Ireland (see also Great Britain), 74, 188, 194, 200, 664 Irian Jaya, 17 iron, 63, 369, 370, 588 iron-pan development, 31 ironwood (Olneya tesota), 311, 317 irrigation, 128, 129, 536 island ecosystems, 426 islands, 14, 15, 143 islands of fertility (see also nucleation and deserts), 314, 770 isolation, 143, 148, 149 isomorphic substitutions, 505 isopods, 68, 88, 276, 312, 349 isostatic rebound, 66, 589 Israel, 276 Italy, 1, 140, 142, 276 Ithaca (New York, USA), 130 ivory, 58 Ivory Coast, 232, 472 Ixtocewatl (Mexico), 1 jack pine (Pinus banksiana) (see also pines), 454, 459, 572 jackdaws (Corvus monedula), 407 jackrabbit (Lepus californicus), 314, 319, 646 jaguars (Panthera onca), 473 Jamaica, 124, 636 James Bay, 175 Janaina, 1 Japan, 140, 142, 148, 155, 172, 188, 195, 197, 198, 204, 207, 208, 211, 405, 407, 455, 588 Japanese knotweed (Fallopia japonica = Reynoutria japonica), 427 jarrah (Eucalyptus marginata), 263, 281, 380 Jengka Forest Reserve (Malaysia), 226 jet stream, 50 Jevons’s paradox, 725, 735, 736, 770 Johannesburg (South Africa), 366, 367 Jostedalsbreen (Norway), 20, 21, 30

GENERAL INDEX Jotunheimen (Norway), 19, 23–25, 28, 30 Jujuy (Argentina), 495 juniper (Juniperus spp.), 130 – community, 298 K-selection, 592, 596 Kabetogama Peninsula (Minnesota, USA), 580 kakadu (monsoonal forest), 416 Kalahari Desert (southern Afrika) (see also deserts), 308 Kalimantan (Borneo, Indonesia), 234 Kamchatka (Russia), 140, 146, 156 kangaroo rats (Dipodomys spp.), 319, 645 Kansas (USA), 196, 294, 295, 297, 298, 340, 575, 639, 640 kaolinite, 506 Karakum Desert (Turkmenistan) (see also deserts), 323 Karo La Pass (Tibet), 47 Karoo (South Africa), 318, 320 K˚arsa Glacier (Sweden), 26 karst, 590, 770 Kashmir (India), 57 Katmai (Alaska, USA), 140, 144, 156, 587 Kautz Creek lahar, 139, 145, 148 Kautz Creek (Mount Rainier, USA), 146 Kekla (Iceland), 142 Keller Peninsula (Antarctica), 77 Kentucky (USA), 189, 195, 201, 205, 415, 416 Kenya, 470 Kerguelen Islands, 69, 91 Kershop Forest (Cumbria, UK), 548 keystone species, 473 Kilauea (Hawaii, USA), 152, 587, 588 King George Island (Antarctica), 53, 77 king penguins (Aptenodytes patagonicus), 70, 91 kipukas, 152 Kissimmee River (Florida, USA), 682, 683 Klutlan Glacier (Yukon Territory, Canada), 28–31 kochia (Kochia scoparia), 389 Kola Peninsula (Russia), 57, 62, 63, 65 Kolombangara (Solomon Islands, Papua New Guinea), 240 Konza (Kansas, USA), 297 Krakatau (Indonesia), 1, 137, 138, 142, 149, 154, 587, 591 krill, 53, 59, 91, 94 Krkonose Mountains (Czech Republik, Poland), 49 krummholz, 48–51, 198, 770 Ksudach (Kamchatka, Russia), 140, 141, 149, 150 kwongan, 271, 281, 282 Kyushu (Japan), 142 La Pampa (Argentina), 489 La Paz (Baja California, Mexico), 312 La R´eunion (Mascarene Islands), 137, 138, 142, 154, 415, 416, 436, 440 La Selva (Costa Rica), 225, 227, 228, 231, 233 labile nutrients, 564 labile organic carbon, 559 labile substrate, 564 Labrador (Canada), 55 lag periods, 23, 444 lagomorphs, 318

GENERAL INDEX lahars, 51, 137–140, 142, 144, 145, 148, 149, 152, 155, 771 Lake Agassiz plain (North America), 577 Lake Duparquet (Quebec, Canada), 175, 176 lakes, 91, 432, 682 Laki (Iceland), 137 land degradation, 665, 666, 668 land grants, 58 land managers, 15, 521 land tenure, 690, 691 land-use, 2, 690, 691, 700, 717 land-use history, 493 landscape, 3, 18, 331–333, 335, 336, 344, 356, 711, 715, 716, 718 – boundary, 715 – context, 714, 715 – deﬁnition of, 708 – design, 682 – disturbance, 711 – diversity, 494 – dynamics, 492 – ecology, 595 – geometry, 682 – heterogeneity, 454, 641 – human effects, 354, 355 – position, 514 – – of wetlands in, 331, 335 landscape-scale processes, 675, 682 landscape-scale restoration, 675, 682 Landschaft, 18 landslides (see also avalanches), 3, 4, 27, 41, 44, 51, 59, 66, 124, 161, 229, 334, 426, 435, 586–590, 592–594, 597, 599, 749 Langdon Head (County Durham, Great Britain), 375 lapilli, 138, 142–144 Lappland (Scandinavia), 63 larch (Larix spp.), 162, 163, 167, 457 Las Vegas (Nevada, USA), 312 later-seral species, 532 later-succession (see also succession), 532 lateral blasts, 141, 142 lateral eruption, 140 lateral moraines (see also glaciers), 20, 22 Latin America, 690 lava domes, 142 lava ﬂows, 139, 586 lava (see also volcanoes), 137, 138, 140–144, 146–149, 151, 152, 154, 156, 587, 588, 591, 771 leaching, 771 leaching (see also nutrient cycling), 26, 29, 401, 512, 547–549, 551, 554, 558, 575 – base cations, 551 – nitrate, 545, 549–553 – nitrogen, 549–551, 553 – – inorganic, 550 lead, 64, 65, 365, 369, 375, 402–404 lead tolerance, 403 leaf area, 257, 462, 477, 571, 615, 621 leaf-fall, 531 leaf hoppers, 639 leafy spurge (Euphorbia esula), 392 legends, 1

849 legumes, 523, 760 Lehman’s lovegrass (Eragrostis lehmanniana), 648 Lekkerkerk (The Netherlands), 666 lemmings, 55 Lepidoptera (see also butterﬂies, caterpillars, moths), 259, 318 lesser snow geese, 43, 53, 55, 335, 348, 353, 579 Lhasa (Tibet), 61 liana, 637 lichens, 24, 27, 31, 45, 55, 62, 64, 68, 73, 77–80, 84, 87, 88, 90, 93, 131, 140, 141, 143, 149, 166, 167, 170, 173, 177, 180, 521, 562–564, 591, 592 Liebig’s Law of the Minimum, 661 life-history characteristics, 598, 602, 771 life-history strategies, 333, 334, 344–348, 499 – animal, 344–348 – plant, 344, 347, 348 life span, 592 light, 257, 262, 263, 599, 621, 623 light competition, 615 light gaps, 190, 191, 195, 197, 198, 202, 212, 771 lightning, 260, 262, 317, 335, 336 lignin, 557, 771 lignotubers, 573 limestone quarries, 371–373 limiting resources, 709 Lincoln (Nebraska, USA), 127 lingonberry, 170 linseed (Linum usitatissimum), 490 litter [also: litterfall], 531, 532, 554–557, 559–561, 564, 594 – accumulation, 496 – communities, 261 – cover, 561 – decomposition, 292, 293 – deposition, 441 – quality, 292, 293, 302 – removal, 574 – succession, 531 – types, 523 – windthrow pits, 210 litterfall, 771 little bluestem (Schizachyrium scoparium), 648, 649 Little Ice Age, 18, 19, 32, 174 little pocket mouse (Perognathus longimembris), 389 liverworts, 140, 144, 153 livestock, 44, 59, 290, 291, 303, 323, 432, 677 livestock and rabbit grazing, 677 livestock trails, 561 – nutrient cycling, 561 lizards, 282, 316, 636, 645 llama (Lama glama), 60 llanos, 423 loblolly pine (Pinus taeda) (see also pines), 454, 457, 550, 552 lodgepole pine (Pinus contorta) (see also pines), 455, 458, 459 loess, 21, 25, 47 logging, 1, 67, 198, 205, 206, 213, 229, 335, 338, 353, 354, 463, 470, 472–475, 478, 481 – damage, 472 – salvage, 214 logs, 202

850 logs of windthrow trees, 210, 211 Loki, 1 London (England), 405 long-distance dispersal, 144, 156, 591 long-term studies, 194, 196 longevity, 592, 593, 602, 771 – plants, 310, 320 longleaf pine (Pinus palustris) (see also pines), 454 – old-growth, 392 loons (Gavia spp.), 43 Los Angeles (California, USA), 278, 749 Louisiana (USA), 195, 334, 336, 338, 339, 342, 349–353, 355, 356 Love Canal (USA), 666 lovegrass (Eragrostis spp.), 648 low-fertility soils, 536 low temperatures, 310, 311 lucerne, see alfalfa Lucerne Valley (California, USA), 312 Luquillo, 1 Luquillo Experimental Forest (Puerto Rico, USA), 230, 232, 235, 243, 245, 636, 637 Luquillo Mountains (Puerto Rico, USA), 259, 763 Macaroni penguins (Eudyptes chrysolophus), 91 Mackenzie Valley, 164 Macquarie Island (Antarctica), 52, 55, 66, 68–70, 75, 78, 88, 91 macro-climate, 19 macroorganisms, 522 Madagascar, 702 magnesium, 548, 587 magnitude of disturbance, 140, 191, 214, 348, 349, 355, 356 mahogany (Swietenia spp.), 472, 478, 480, 482 – large-leaf mahogany (S. macrophylla), 472 – little-leaf mahogany (S. mahagoni), 472 maidencane (Panicum hemitomon), 336, 351 Maine (USA), 189, 193, 194, 200, 206, 459 maize (Zea mays), 467, 490, 492, 494 Malawi, 470 Malaysia, 226, 227, 230, 232, 469, 667 Maldives, 137 malleability, 678 mallee, 271, 281, 573, 771 mammal disturbance, 435, 442, 443 mammals, 43, 52, 89, 130, 262, 351, 398, 407, 642, 643, 645, 750 man orchids, 371 management, 14, 717, 718, 748, 759, 760, 765 management myths, 683 management of wetlands, 334, 335, 354, 356 manatees (Trichechus manatus), 332 mangal, see mangrove wetlands manganese, 369, 588 manganese uptake, 524 mangold, 130 mangrove swamps, 347 mangrove wetlands (see also wetlands), 332, 338, 345, 352 – ﬂooding, 341 – frost damage, 336 – herbivores, 341, 355

GENERAL INDEX – hurricanes, 336, 341, 352, 354, 355 – sedimentation, 334, 352 – seedling establishment, 348 mangroves, 235, 331, 332, 341, 392, 665 Manhattan (Kansas, USA), 297 manioc ( = cassava: Manihot spp.), 124, 467 Manitoba (Canada), 343, 392, 619 mantids, 639 mantle, 525 manures, green, 536 Maori, 676 MAPET (mean annual potential evapotranspiration), 295 maple (Acer spp.) (see also sugar maple), 205, 403, 406, 548 mapping, 17 maquis, 271, 276, 279 marginalization, 666, 667, 771 mariculture, 334 marine algae, 64 mariola (Parthenium incanum), 318 Marion Island (Subantarctica), 54, 69, 70, 88 maritime chaparral (see also chaparral), 423 marmots (Marmota spp.), 82, 513 marram (Ammophila arenaria), 441 marri (Eucalyptus calophylla), 381 Marsh Arabs, 334 marsh buggies, 339 Marshall silt-loam, 127 marshes (see also wetlands), 52, 73, 80, 331–334, 336, 337, 339, 343, 353, 433, 771 Martinique, 142, 143 Maryland (USA), 351 Marzellferner (Austria), 22 Mascarene Islands, 142 mass movement, 21, 23–25, 51, 52 Massachusetts (USA), 189, 195, 198, 350 matorral, 271, 274, 279, 280, 676 – coastal, 272 matrix, 675 matter ﬂow, 710, 718 mature communities, 32 Mauna Loa (Hawaii, USA), 147 Mauritius, 420, 636 Maya, 1, 2, 467, 663 meadow, 534 meadow cordgrass (Spartina patens), 343 meadowlarks, 390 mean annual potential evapotranspiration (MAPET) (see also evapotranspiration), 295 mechanisms of invasion, 426 mechanized maneuvers, 387 Mediterranean, 2, 271, 272, 274–279, 325, 430, 441, 576, 594, 661, 665, 679 – desertiﬁcation, 276, 277 – human history, 273 – shrublands, 5 – species, 417 Mediterranean climate, 771 Mediterranean-type ecosystems (MTE), 271–282, 390, 425, 676 – climate, 271

GENERAL INDEX – conservation of, 282 – ﬁre, 271, 272, 278 – land-use history, 273–275, 277–282 – vegetation, 271, 272 megafauna, 289 meiofauna, 332, 337 melanin, 42 melanized fungi, 532 meltwater, 22 meltwater channel, 29 Mendoza (Argentina), 495 mercury, 64, 369 Mersing (Sarawak, Malaysia), 233 mesic, 771 mesic forests, 438, 439 meso-climatic, 19 mesohaline marshes, 332 Mesopotamia, 2 mesquite grasslands, 386 metalliferous mining wastes, 371–376 – copper and cobalt ﬂora, 372, 374 – metal tolerance, 376 – metal-tolerant plants, 374 – metallophytes, 372–375 – plant–soil factors, 374–376 – restoration, 376 metallophytes, 372–375, 771 metapopulation, 718 metastability, 718 methane, 89, 401 methodology, wind storm studies, 201 Mexican volcanoes, 147 Mexico, 1, 137, 142, 144, 148, 189, 195, 226, 232, 240, 323, 324, 418, 423, 425, 470, 477, 587, 636, 695 – grassland, 574 Mexico City, 147 mice, 69, 499 Michigan (USA), 128, 170, 189, 193, 200, 204, 205, 212, 213, 596, 613, 614, 618, 619, 622, 625 Micky Mouse bush (Ochna natalitia), 380 microbes, 130, 545, 556, 565 – activity, 561 – biomass, 523, 555, 557–564 – breakdown, 524 – carbon, 533, 534 – decomposition, 546, 547, 555–559, 561, 562 – dynamics, 558 – immobilization, 550, 551, 553, 554 – mass, 531 – mineralization, 553, 558 – on recently-deglaciated terrain, 27 – re-invasion, 521 – richness, 527 microbial-mediated processes, 526 microcatchments, 536, 680 microclimate, 243 – post-ﬁre, 575, 577 microfauna, 496, 523 microlichens, 27

851 Micronesia, 14 microorganisms (see also bacteria), 26, 27, 43, 521–538 – establishment, 529–531 – responses to disturbance, 526, 527 – succession, 531–535 microphyllous leaves, 310 microrelief, 24 microsites, 29, 140, 151, 188–190, 202, 208–211, 214 microsymbionts, 677 microtines, 55 microtopography, 146 mid-alpine vegetation, 19 mid-Atlantic region, 696 Middle East, 275, 713 migration, 27, 41, 51, 84, 88, 93, 528, 586, 717 Migratory Bird Treaty Act, 391 military, 4, 316 – exercises, 385–394 – lands, 386 – wastes, 65 millipedes, 523, 640 Mima mound, 529 mine sites, 674, 680 mineral resources, 365 mineral soils, 332 mineralization, 527, 549, 558, 561, 771 – nitrogen, 553 – rates, 560 minerals, 331 miners, 254 mining, 2, 4, 63, 86, 173, 279, 281, 282, 586, 673 – disturbance, 366 – ecosystems, 370 – extraction, 366 – geographic centres, 366 – history of, 365, 366 – land disturbance, 366–368, 381 – – mineral processing, 367 – – stockpiles, 367 – – subsidence, 367 – – surface spoil heaps, 367 – – waste-disposal facilities, 367 – methods, 366–368 – – deep underground mining, 366, 367 – – dredge mining, 367 – – open pit, 367 – – open-pit mining, 367 – – quarrying, 367 – – shallow underground mining, 366, 367 – – strip mining, 367 – milling, 366 – processing, 366 – quarrying, 366 – restoration, 370, 373, 378–382 – substrates, 370, 371 – succession, 366, 370, 381, 382 – waste disposal, 366 mining wastes, 368, 369, 381, 586–588, 592, 594, 599 – restoration, see restoration under mining

852 minke whale (Balaenoptera acutorostrata), 91 Minnesota (USA), 170, 189, 196, 201, 203, 206–210, 213, 614, 615, 619, 620, 622, 623 miombo (Brachystegia ﬂoribunda), 374 mires (see also wetlands), 180, 331, 332 Mississippi River, 131, 334 Mississippi (USA), 353, 443 Missouri River, 131 Missouri (USA), 123, 297 mites, 372, 523, 532, 593, 639, 647 mitigation, 392, 393, 759 mixed conifer forests (see also conifers), 425 mixed forests (see also forests), 392 mixed-species, 480 mobilization, 527 mode locking, 740–742 modeling, 14, 15 models, 488, 489, 660 – CENTURY model, 292 – Connell’s Model IV, 499 – core–buffer–matrix models, 675 – hump-backed model, 488 – lottery model, 154 – prediction of change in species richness and measurements, 499 – press perturbation model, 499 – Revised Universal Soil Loss Equation (RUSLE), 513, 514 – structural, 707, 708, 711, 712, 719 – Swift’s models, 499 – three-phase conceptual model, 503 – three-phase soil model, 503–508 – Universal Soil Loss Equation (USLE), 513 – Water Erosion Prediction Project (WEPP), 513, 514 – Wind Erosion Prediction System (WEPS), 513 modiﬁed ecosystems, 659, 660 moisture, 257, 258 moisture availability – post-ﬁre, 574 moisture deﬁcit, 21 Mojave Desert (USA), 271, 278, 308, 311–313, 316, 317, 390, 420, 423, 428, 429, 439, 749 molluscs (see also slugs, snails), 68, 337, 350, 432 Mongolia, 580 monitoring, 660, 664, 698, 699, 703 monocultures, 259 Monona Ida silt loam, 127 monophagy, 771 monsoonal forests, 416 monsoons, 44 montados, 276 Montagne Pel´ee (Martinique), 142, 143 Montana (USA), 123, 124, 416, 420, 421, 460, 533 montane, 416, 418 – coniferous forests, 420 – forests, 205, 226, 231 – grassland, 420 – woodland brush, 386 Monteverde (Costa Rica), 227, 231, 233 moors (see also wetlands), 332 moose (Alces alces), 169, 170, 579, 593, 750

GENERAL INDEX moraines, see glaciers morphotypes of species, 492 mortality, 26, 461, 462, 613, 617, 619, 624 – in windstorms, 199–201 mosaic, 711, 718 Moses, 1 mosquitos, 334 mosses, 24, 55, 82, 87, 140, 142–144, 331, 562–564, 618 moths (see also Lepidoptera), 260, 262, 444, 455, 474, 552 Motmot (Papua New Guinea), 142 mounds, 193, 771 – treefall, 713 – wind throw, 202, 205, 206, 208–210, 212 Mount Erebus (Antarctica), 64 Mount Kenya (Kenya), 30 Mount Kula (Turkey), 142 Mount Lassen (USA), 142 Mount Ontake (Japan), 142, 145 Mount Pinatubo (Philippines), 137, 155, 156 Mount Rainier (Washington, USA), 139, 142, 145, 148, 155 Mount St. Helens (Washington, USA), 139–145, 147–156, 528, 531, 532, 588, 591, 593, 713 Mount Success (New Hampshire, USA), 552 Mount Takumbe (Japan), 139 Mount Taranaki (New Zealand), 142, 145, 147 Mount Tarawera (New Zealand), 142, 147, 148, 153, 154 Mount Tolbachik (Russia), 142 Mount Usu (Japan), 151, 154, 155 Mount Vesuvius (Italy), 140, 142, 155 mountain birch (Betula pubescens ssp. tortuosa), 171 mountain hemlock (Tsuga mertensiana), 454 mountainous, 124 mucilaginous coatings, 26 mud ﬂows (see also volcanoes), 51, 137 Muddus National Park (Sweden), 165, 177, 179–181 Muddy River lahar (Mount Rainier, USA), 139, 145 mulch, 534 Muldrow Glacier (Alaska, USA), 29 mule deer (Odocoileus hemionus), 697 multiple successional pathways, 678 multivariate statistics, 30 munitions, direct impacts, 385 Murmansk (Russia), 57, 65 musk orchids (Ophrys spp.), 371 muskeg (see also wetlands), 30, 31, 332, 577 muskoxen (Ovibos moschatus), 55, 83, 84, 90 muskrats (Ondatra spp.), 337, 351, 355, 645, 751 mussels, 350 mutualism, 523, 598 mutualistic–competitive dynamics, 525 Myanmar, 470 mycelial growth, 525 mycelial matrix, 522 mycoparasites, 530 mycophagy, 529 mycorrhizae [mycorrhizas], 153, 211, 212, 376, 521, 575, 597, 598, 647, 649, 681, 771 mycorrhizas, 311, 318, 472, 481, 751 mycostasis, 530

GENERAL INDEX mycotrophic species, 532 myth of steady-state efﬁciency, 735–739 myths, 1, 663, 735–739 myxomatosis, 69

15 N isotope dilution, 545 Namib Desert (Namibia) (see also deserts), 308, 311, 316, 324 Nan Shan (Tibet), 61 Napier silt loam, 127 narrow-leaf cattail (Typha angustifolia), 348 Natal mahogany (Trichilia emetica), 380 National Environmental Policy Act, USA (NEPA), 391, 690, 696, 697 National Forest Management Act, USA (NFMA), 694, 698–700 national parks, 32, 69 native – community, 497 – ﬂora, 404 – grasses, 536 – grassland, 416 – herbivores, 289 – peoples, 666 – plants, 398 – prairie, 416 – species, 404, 406, 481, 482 – – recovery, 417 – – species evolution versus environmental change, 492 – woodlands, 675 natural – colonization processes (see also colonization), 487 – disturbance regimes (see also disturbance), 272, 273, 333–337, 341, 343–345, 354, 356, 414, 432, 434 – disturbances, 40–56 – forest management, 470, 471 – forests, 469 – gas, 63, 365 – regeneration, 469, 470, 480, 481 – resources management, 725, 740 – selection, 710 – system, 499 naturalization, 404 naturalness, 675 Nature Conservancy agency, 2 nature conservation, 674 nature protection, 398 nature reserves, 2 NDVI-I (normalized difference vegetation index), 296, 300, 302 Nebraska (USA), 297, 639 nectarivores (see also herbivory), 636, 637, 642 needle ice, 30, 43 Negev Desert (Israel) (see also deserts), 131, 312, 313, 319, 323 neighborhood effects, 191, 192 nematodes, 130, 258, 332, 593 neo-malthusians, 723, 724, 744, 771 Neoglacial (see also glaciers), 18 neophytes, 405, 406 Nepal, 40, 49, 60 nests, 53

853 net nutrient mineralization (see also nitrogen under mineralization), 558, 562 net photosynthetic rate (see also photosynthesis), 571, 619 net primary production, 453 Netherlands, 340, 589 Nevada (USA), 418, 420–423 New Brunswick (Canada), 195, 548 New England peppermint (Eucalyptus nova-anglica), 264 New England Tablelands (New South Wales, Australia), 265 New England (USA), 193–195, 198, 201, 205, 338, 349, 350, 618, 622 New Guinea, 255 New Hampshire (USA), 51, 189, 195, 206, 210, 213, 615 New Jersey pine barrens, 701 New Jersey (USA), 190, 201, 208, 212, 613–615, 618, 622, 625 New Mexico (USA), 131, 318, 322, 378, 416, 526, 613 New South Wales (Australia), 263–265, 433 New York (USA), 130, 190, 193, 194, 208, 209, 213, 233, 401, 428, 455, 618, 619 New Zealand, 14, 17, 30, 31, 66, 90, 142, 145, 147, 148, 154, 188, 200, 206, 208–211, 214, 433, 436, 438, 441, 443, 461, 588, 591, 594, 676, 681 Nicaragua, 238–240, 636 niche, 625, 626 niche differentiation, 625–627 nickel, 365, 369 Nigardsbreen (Norway), 30 Niger River, 666 Nigeria, 124, 241 Nile River (Egypt), 2, 316 Nin Jin Kan Shan (Tibet), 47 nitrate, 262, 455, 545, 547–551 – adsorption, 553, 554 – leaching, 545, 546, 548, 552, 565 – losses, 546, 547, 550, 553, 554 – reductase, 524 – soil solution, 548, 550–552 – sorption, 546 – stream water, 546, 548, 549, 552 nitriﬁcation, 523, 545, 546, 548–551, 553, 554, 771 nitriﬁers, 533, 545, 565 nitrogen (see also nutrient cycling), 87, 258, 261, 262, 278, 281, 371, 398, 401–403, 406, 419, 442, 455, 457, 458, 472, 477, 587, 588, 600, 601 – assimilation, 558 – availability, 557 – budgets, 554 – concentrations, 293, 558, 560 – content in biomass, 292–294 – cycling, 523, 546 – – effects of rodents, 580 – – post-ﬁre, 575 – fertilization, 533, 549, 553 – ﬁxation, 53, 68, 143, 406, 521–523, 551–554, 562, 564, 586, 594, 597, 600, 771 – – post-ﬁre, 575 – immobilization, 550, 553, 557 – ineralization, 524 – inputs, 557

854 nitrogen (see also nutrient cycling) (cont’d) – losses, 546, 547, 549, 552, 553, 565 – – stream water, 548 – mineralization, 435, 548–551, 553, 557, 559 – oxidation, 522 – retention, 546, 553, 554 – saturation, 546, 549 – transformations, 353 – uptake, 552 nitrogenase, 563 niveo–aeolian erosion, 25 Niwot Ridge (Colorado, USA), 48, 50, 54, 84, 86 Noah, 1 nodulation (see also ﬁxation under nitrogen), 525 nomadic pastoralism, 62 non-equilibrium dynamics, 678 non-equilibrium state, 491 non-equilibrium systems, 493 non-governmental organizations (NGOs), 666, 667 non-melanized fungi, 532 non-mycorrhizal plants (see also mycorrhizae), 524 non-mycotrophic species, 532 non-native species, 399, 405, 408, 417, 422, 681 Nooksack Cirque (Washington, USA), 29 Noril’sk (Russia), 57 normalized difference vegetation index (NDVI-I), 296, 300, 302 Norse mythology, 1 North Africa, 308, 314, 324, 536, 663, 667 North America, 14, 15, 17, 57, 162–164, 167, 168, 170, 175, 187, 192, 193, 197, 206, 210, 214, 255, 259–261, 287, 288, 300, 308, 311, 312, 314, 322, 333–335, 343, 345, 350, 397, 417, 419, 425, 444, 455, 456, 574, 577, 580, 663, 681, 683, 693 – forests – – ﬁre, 572, 573 – – regeneration, 572, 573 – grassland, 574 – shrublands, 575 North American Great Plains, 556, 557 North Atlantic Treaty Organization (NATO), 391 North Carolina (USA), 204, 208, 212, 334, 336, 427, 428, 455, 553, 614, 617–619, 625 North Dakota (USA), 294, 639 North Pole, 39 North Sea, 68 North Vietnam, 385 northern bobwhite (Colinus virginianus), 639 northern desert, 386 northern hardwood forests, 192, 193, 206, 213 northern hardwood–conifer forests, 386 northern peatlands, 333 northern spotted owl (Strix occidentalis), 698, 704 Northwest Territories (Canada), 164, 577 Norway, 18–21, 23, 26, 28, 30, 63, 65, 67, 79, 164, 591, 592 Norway spruce (Picea abies) (see also spruce), 548 Nouragues (French Guiana), 231 Nova Scotia (Canada), 340 NOx , 401 nuclear weapons, 64 nucleation (see also islands of fertility), 151, 152, 591

GENERAL INDEX nudation, 17, 20, 586, 713 nu´ees ardentes, 138, 143 nunataks, 26 nurse logs, 211 nurse-plants, 153, 311, 317, 319, 320, 325 nutria (Myocaster coypus), 337, 342, 351, 751 nutrient cycling, 14, 15, 41, 244, 545–565, 752 – effect of beavers, 580 – erosion, 560 – nitrogen, 545, 552–554, 562, 563, 565 – soil terms, 547 – stimulation, 571 nutrient-retention processes, 679 nutrients, 26, 29, 31, 53, 90, 91, 211, 212, 244, 258, 262, 331, 333, 344, 352–354, 398, 402, 403, 458–462, 527, 589, 599, 620–622, 626, 711 – accumulation, 563 – availability, 331–333, 349, 351–353, 355, 558, 561 – – post-ﬁre, 572–575 – budgets, 547 – concentrations, 560, 563 – cycling, 331, 333, 352 – – rates, 555, 556, 564 – ﬁre, 352, 353 – ﬂush, 26 – grazing, 353 – immobilization, 31 – in succession, 31 – labile, 556, 557 – limitation, 572 – losses, 455, 459, 462, 472, 545, 550, 552, 561 – – calcium, 547 – – ﬁre, 575, 576 – – potassium, 548 – mineralization, 557, 558, 560, 561, 563, 565 – mobility, 523 – pollution, 337, 355 – pools, 560 – retention, 552, 562 – return to soil, 560 – transformations, 333, 351, 353 – transport, 565 – uptake, 523 Oak Ridge (Tennessee, USA), 552 oak savanna, 386 oak woodlands, 416, 438 oaks (Quercus spp.), 201, 206, 259, 261, 262, 271, 274–276, 278, 279, 377, 400, 401, 403, 455, 589, 591, 614, 635, 647, 710 oats (Avena sativa), 59, 127 Oceania, 434 oceanic circulation, 66 oceanic climate, 42 off-road vehicles (see also tracked vehicles and roads), 316, 324, 335, 337, 339, 340, 342 Ohio (USA), 190, 204, 206, 367, 428, 459, 613 oil, 63, 85, 365 oil spills, 67, 68, 85, 89, 90, 335, 337 Okefenokee Swamp (Georgia, USA), 336

GENERAL INDEX Oklahoma (USA), 2, 196, 294, 297, 377 old ﬁeld (see also competition intensity and forests), 612–615, 617–619, 621–626, 714 old-growth forests, 29, 187, 188, 190, 195, 197, 203–205, 207, 208, 210, 211, 213, 214, 471, 474, 478, 771 old-growth stage, 461 oligochaetes (see also earthworms), 332 oligohaline marshes (see also wetlands), 332, 348, 351, 356 olives (Olea europaea), 276 Olney’s three-square (Scirpus americanus), 343, 355 omnivores, 636 Omnsbreen (Norway), 26, 30 onion, 528 Ontario (Canada), 550, 617, 619, 620, 625 orchids, 637 ordination, 17, 30, 80 ore (see also mining), 588 Oregon (USA), 190, 198, 202, 203, 211, 421–423, 427, 463, 613, 618, 622, 701 – forests, 574 Organ Pipe Cactus National Monument (Arizona, USA), 312 organic acids, 523 organic carbon, 545–547, 549, 556, 558 – inputs, 555–558, 560–564 – labile, 555, 560, 564 – non-labile, 556 organic fallout, 151 organic matter, 89, 91, 126–130, 132, 314, 333, 342, 343, 352, 353, 355, 461, 472, 488, 496, 506, 555, 557, 558, 588, 590, 600, 601 – decomposition, 556, 557, 560, 561, 564 – inputs, 546, 555, 556, 559–561 – landscape patterns control inputs, 558 ortho-phosphate, 548 osmotica, 344 ostracods, 332 ¨ Otztal Alps (Austria), 22 outwash plain, 21 overﬁshing, 59 overgrazing, 60, 126 overharvesting, 472 overland ﬂow, 561 owls, 698, 704 oxalates, 524 Oxford ragwort (Senecio spp.), 406 oxygen partial pressure, 523 oyster (Ostrea spp.), 87 ozone, 64–66, 93, 257 Paciﬁc black brant geese (Branta bernicla nigricans), 70 Paciﬁc Northwest, 419, 429, 454, 694 Paciﬁc Ocean, 5, 128, 137, 195, 224, 225, 636 Paciﬁc walrus (Odobenus rosmarus divergens), 58 Pack Forest (Washington, USA), 551 pahoehoe, 138, 149 paleoecology, 191, 213 paleotropical forests, 226 paloverde (Cercidium spp.), 312 paludiﬁcation, 31

855 Pamir Mountains (central Asia), 56 pampas, 487–501, 771 Pampas grass (Cortaderia jubata), 426 Panama, 225, 230, 232, 255, 467, 475, 642 pandora moth (Coloradia pandora), 257 paper birch (Betula papyrifera), 167 paperbark (Melaleuca spp.), 392 Papua New Guinea, 142, 473, 474 Par´a (Brazil), 476 Para (Ivory Coast, Africa), 231 paraglacial disturbance, 21, 23, 28, 30 p´aramo, 60, 90 parasites, 4, 444 parasitism, 523, 602 particle segregation, 562 particle size, 504 partridge (Coturnix coturnix), 398 Pasoh Forest Reserve (Malaysia), 229 Pasoh (Malaysia), 231 Pasterze Glacier (Europe), 28 pastoralists, 44 pasture, 476, 477, 626 – old, 497, 499 Patagonia (Argentina), 66, 208, 287, 295, 311, 314 patch bodies, three-dimensional, 715 patch dynamics, 187, 191, 192, 202–206, 212, 223, 333, 336–339, 344, 347–351, 419–422, 462, 464, 493–497, 585, 595, 708, 715–718, 762, 763, 771 – deﬁnition, 716 patches, 3, 70, 71, 463, 526, 532, 613, 618, 623, 709, 714 – origin, 715, 716 – resource patch, 716 – size, 458, 463 patchiness, 141, 307, 588 Pathﬁnder Land Program data set, 300 pathogens, 259, 272, 442, 444, 678 pattern, 675 pattern analysis, 612 pattern of disturbance (see also disturbance) – spatial, 339, 340 – temporal, 340 patterned ground, 24, 26, 28 PCBs (polychlorinated biphenyls), 65 peak biomass, 296 peat, 89, 180, 331, 335 – accumulation, 336, 355 – burning, 342, 353, 355 – formation, initiated by ﬁre, 577 – harvesting, 335, 337 peatlands, 332, 333 Pechora Basin (Russia), 63 Pele, 1 penetration resistance, 510 penetrometer, 511 penguins, 30, 44, 53, 54, 62, 68, 70, 89, 91 Pennines (England), 375 Pennsylvania (USA), 190, 195, 198, 203, 205, 208–213, 378, 428, 433, 459 pepper tree (Schinus terebinthifolius), 441

856 per-gram effects, 621, 622 perennial, 771 perennial glasswort (Salicornia virginica), 340 perennial grasses, 498 perennial shrub species, 527 permafrost, 47, 52, 163, 167, 771 – effects of ﬁre, 577 persistence, 40 Peru, 56, 308, 469, 470, 479 Peruvian peppertree (Schinus molle), 432 pervection, 21, 23, 26, 29 pesticides, 499 petrels, 54, 59, 70, 91 petroleum hydrocarbons, 342, 353 pH, 262, 523 phalanx invasions, 152 phasmids, 637 pheasants, 639 phenotypic plasticity, 172, 348, 349, 495 Philippines, 124, 126, 137, 334, 436 phoresy, 529 phorid ﬂies, 647 phosphatase production, 523 phosphorus, 371, 398, 472, 521, 523, 548, 588, 594 photoinhibition, 42, 66, 87, 472, 594 photooxidation, 85 photoperiod, 89 photosynthesis, 42, 66, 87, 477 photosynthetic rate, 261, 617 phreatophyte, 771 phrygana, 272 physical amelioration, 141, 151 physical environment, 3, 17, 487 physical processes, 18 physiological acclimation, 344, 348, 349 physiological condition of plants, 253 phytobiont physiology, 530 phytogenic mounds, 314 phytophagous insects, 171 phytoplankton, 41 phytosociology, 17 phytotoxins, 508, 512, 574 pickerelweed (Pontederia cordata), 336 Piedmont (Georgia, USA), 126 Piedmont (North Carolina, USA), 550 pigeon wood (Trema orientalis), 380 pigeons (Columba livia f. domestica), 398, 407 pigments, 87 pigs (Sus scrofa) (see also alien species), 435, 442, 443, 636, 681 – feral, 435 pikas (Ochotona spp.), 55, 56, 58, 82 pin-hole borers, 636 Pinelands (New Jersey, USA) (see also New Jersey), 713 pines (Pinus spp.), 62, 162–170, 174, 177, 180, 182, 201, 212, 260–262, 274, 281, 333, 354, 400, 425, 440, 469, 470, 480, 573, 577, 760 – age–structure diagrams, 182 – bark beetles, 201 – northern limits, 169

GENERAL INDEX – types, 163–166 pingos, 167 pinyon–juniper–oak woodland, 386 pioneer species (see also succession), 26, 29, 141, 145, 147, 154, 319 pioneer zone (see also succession), 30 pipelines, 316 pitch pine (Pinus rigida) (see also pines), 190, 212, 713 pitcher plant (Sarracenia spp.), 340 pits, 209, 771 – animal, 713 – wind throw, 202, 208–210, 713 pitting, 536 plankton, 331 plant-available water, 507 Plantaginetea, 405 plantain (Plantago lanceolata), 374 plantations (of trees), 469, 470, 477, 479, 481, 482 plants – abundance, 257–259 – apparency, 258 – biochemical defenses, 254 – biomass, 558 – communities, 17, 490, 491 – condition, 257–259, 266 – cover, foliar, 561 – damage, 24–26, 28 – – by frost, 25 – defenses, 258, 260 – density, 266 – establishment, 311, 312 – growth, 253, 261, 266 – life histories, 595 – longevity, 310, 320 – management, 537 – material, 555, 556, 560 – – particle size, 556 – migration, 27 – morphotype, changes over time, 493 – mortality, 253, 254, 258 – nutrient uptake, 524, 550, 551, 553, 554, 557, 564 – nutrients, 126 – pathogens, 512 – physiological condition, 253 – plasticity, 627 – productivity, 531 – reproduction, 266 – reproductive strategies, 26 – roots, 23, 527 – spacing, 259 – stress, 258, 259 – succession (see also succession), 156, 556, 563 – survival, 266 plasticity, 771 playas, 313, 331, 334, 348, 771 Pleistocene, 280, 333 ploughing, 680 plow sole, 510 Plynlimon (Wales, UK), 548, 549

GENERAL INDEX pocket gophers, 82 – Geomys spp., 580, 645, 647 – Thomomys talpoides, 54, 55 Pococatepetl (Mexico), 1 pocosins, 331, 332, 336 podocarp, 188, 209 podzolization, 31 point richness, 496 Point Thomas, King George Island, South Shetland Islands (Antarctica), 53 Poland, 49, 188, 210, 213, 405, 430, 431, 589 polar bear (Ursus maritimus), 58, 64, 65, 68, 70 polar deserts (see also deserts), 45, 62, 76, 308 polar environments, 18 polar ice shelves, 66 polar vegetation, 32 policy, 391 – agriculture, 689, 691, 700 – developing countries, 690, 691 – economics, 701, 702 – forestry, 689, 698, 699 – human population, 692 – land-use, 691 – organizations, 689 – science, 702, 703 pollination, 318, 593, 681 pollinators, 43, 131 pollutants, 91 pollution, 2, 63, 85, 93, 333–335, 337, 353, 354, 403, 590, 664, 666–668, 675, 682 – thermal, 335 polychaetes, 332, 347 polychlorinated biphenyls (PCBs), 65 polygons, 46, 47 polymorphism, 344 polyphagy, 771 polyploidy, 42 Pompeii (Italy), 1 pond pine (Pinus serotina) (see also pines), 349, 355 ponderosa pine (Pinus ponderosa) (see also pines), 257, 261, 454, 456, 457, 459 poplar, see cottonwood population disturbance, 710 population growth, 668 population-level experiments, 612 porcupine (Hystrix indica), 312, 713 pore size, 561 porosity, 561 Portugal, 430 Poseidon, 1 positive associations, 626 possums (Trichosurus vulpecula), 441, 681 post-modern, 663 potassium, 371, 548, 587 – deﬁciency, 512 – stream water, 548 potato (Solanum tuberosum), 59, 129 poverty, 662, 663, 666 – environmental degradation, 662

857 powerlines, 316 Poznan (Poland), 403 prairie dogs (Cynomys spp.), 580, 645 prairie glacial marshes, 351 prairie potholes, 333, 334, 340 prairie vole (Microtus ochrogaster), 640 prairies (see also savannas) – Canada, 388 – Florida (USA), 439 – Iowa (USA), 617 – Kansas (USA), 575, 579 – Manitoba (Canada), 619 – mixed grass, 294, 295, 297, 298 – North America, 574 – Saskatchewan (Canada), 620 – tallgrass, 294, 295, 297, 298 precipitation gradient, 287–303 precipitation (see also rainfall), 44, 45, 262–264, 294–298, 308, 401 – temporal variability, 293–295, 303 predation, 257, 335, 337, 343, 347, 598 predators, 257, 259, 262, 264, 290 predawn water potential, 619 predictability, 14, 15, 151, 340, 748, 764 preserves, nature, 187 Presidential Range (New Hampshire, USA), 51 Pribilof Island fur seals (Callorhinus ursinus), 64 primary colonizers, 27 primary producers, 15 primary production, 14, 53–55, 65, 66, 89, 92, 261, 266, 287–303, 331, 333, 342–345, 352, 353, 558, 561, 564, 675, 752, 771 – components, 571 – grazing, 353 – precipitation, 287, 288 – problems of measurement, 571, 572 – response to – – ﬁre, 572–577 – – herbivory, 578–580 – – heterotrophs, 577, 578, 580, 581 primary succession (see also succession), 17, 27–31, 137, 144, 146, 149, 152, 153, 534, 611, 620, 626, 771 – patterns and processes, 585–602 primary volcanic surfaces, 152 Prince William Sound (Alaska, USA), 701 proactive resource management (see also management), 386 production of secondary metabolites, 495 productivity (see also primary and secondary production), 14, 15, 31, 457, 458, 461, 462, 593, 598, 748 progressive community development, 487 progressive model, 487 Prometheus, 1 propagules, 27, 527, 591, 592, 771 protected areas, 675 protective pigmentation, 87 proteinases, 524 Protozoa, 130 Prudhoe Bay (Alaska, USA), 63 psocopterans, 647 ptarmigans (Lagopus spp.), 58

858 Puerto Rico (USA), 1, 194, 226, 228, 230, 232, 235, 239, 241, 243, 245, 255, 257, 259, 477, 589, 590, 635–638, 642, 643, 749, 754, 762, 763 puffballs, 528 pulse–reserve systems, 310 pumice, 138, 140, 142, 143, 148 Pumice Plain (Mount St. Helens, Washington, USA), 528 purple loosestrife (Lythrum salicaria), 348, 431, 434 push-moraine, 21 pyroclastic – blasts, 140 – deposits, 143 – events, 138, 140, 156 – falls, 138 – ﬂows, 137–142, 771 – rocks, 138 Qilian Shan (Tibet), 61 Quaternary, 174, 273 Quebec (Canada), 175, 176, 189, 614, 618, 622 Queensland (Australia), 230, 264, 392, 472, 475 r-selection, 592, 596 rabbits (Oryctolagus cuniculus) (see also hares), 55, 68–70, 279, 282, 324, 398, 426, 443, 677, 681 radioactivity, 64 rain-loading, 227 rain-shadow deserts, 44 rain-shadow effect, 308 raindrops, 123, 513 – impact, 561 rainfall (see also precipitation), 123, 124, 126–128, 132, 235, 469, 472, 475, 481 – gradient, 234, 235 – intensity, 561 rainforests, 325, 419, 438, 716 – Amazon, 314 – Paciﬁc, 386 rainstorms, 224, 225, 227 Rajasthan (India), 579 Rakata (Indonesia), 138, 147–149 ramshorn snail (Marisa cornuarietus), 432 rangelands, 291, 546, 547, 565, 673 – disturbances, 546 rate of succession (see also succession), 27–29, 141, 143 ratio of plant species to fungal species (see also plants and fungi), 522 rats (Rattus spp.) (see also alien species), 70, 681 – brown rat (Rattus norvegicus), 69, 398 rattlesnakes, 644 reaction, 28, 586 reallocation, 674, 676 recalcitrant, 533 recently-deglaciated terrain, 17, 19 reclamation (see also restoration), 62, 324, 527, 674 recolonization, 527, 532 reconstruction, 674 recovery (see also succession and restoration), 70, 73–75, 78, 79, 84, 92, 230, 462, 475–477

GENERAL INDEX – from disturbance, 355, 356 – limiting factors, 679–681 – mechanisms, 137 – rates, 137, 139, 140 recreation, 57, 673 recruitment, 237, 239, 242, 461 red alder (Alnus rubra), 551–553 red brome (Bromus rubens), 422 red-cockaded woodpecker (Picoides borealis), 392 red mangrove (Rhizophora mangle) (see also wetlands), 341 red maple (Acer rubrum), 350 red oak (Quercus rubra), 261 red pine (Pinus resinosa) (see also pines), 573 red spruce (Picea rubens), 548 reed dieback, 403 reedswamps (see also wetlands), 332 reefs, 354 reference ecosystems, 676 reforestation (see also forests), 281, 480 regeneration, 461, 680, 771 – after disturbance by wind, 202–213 – strategies, 166 regeneration wave, 172, 771 regional-scale processes, 499, 682 – successional change (see also succession), 491, 492 regrowth, 237, 238 regulated rivers, 431, 432 rehabilitation (see also restoration), 324, 674 reindeer (Rangifer tarandus), 23, 54, 55, 58, 60, 62, 64, 69, 83, 84, 92, 168–170, 750 reinvasion (see also colonization), 527–529 – mechanisms, 528, 529 relative humidity, 258, 401 release, 237, 239, 242 reliability, 260 relicts, 152, 497–499 relocation, 666, 667 removal, 618, 673 replacement, 673 repression, 237, 241 reproduction, 43, 71, 92, 261, 535 – sexual vs. asexual, 43 – sexual vs. vegetative, 26 reptiles (see also alligators, lizards), 43, 276, 636, 644 reserves, 93, 94 residual sites, 141 residual species, 152 resilience, 14, 15, 40, 71–73, 92, 660, 661, 675, 678, 679, 694, 703, 748, 754, 755, 757, 771 resistance, 40, 71–73, 92 resource availability, 192, 202 Resource Conservation and Recovery Act, USA, 391 resource extraction (see also mining), 12 respiration rates, 524 responses to ecosystem disturbance, 237, 343–355, 662–665 – catastrophic wind disturbances, 237–241 – community-level, 349–352 – ecosystem-level, 352–354 – landscape-level, 354, 355

GENERAL INDEX – population-level, 348, 349 – species-level, 343–348 resprouting after disturbance, 203, 210, 211, 214 restoration ecology, 674, 675, 678, 679, 681, 683 restoration (see also succession and recovery), 14, 324, 434, 475–477, 482, 536, 585, 673–683, 694, 695, 699, 703, 718, 748, 759, 760, 771 – aquatic systems, 682, 683 – blasting, 373 – goals, 674–681 – management, 15 – terminology, 674 retention, 536 retrogression (see also succession), 27, 31, 771 return time, 191, 193, 214 R´eunion, see La R´eunion revegetation (see also restoration), 79, 85, 459, 511, 536, 589 rhizobial infection, 525 rhizomes, 80 rhizosphere, 27, 523, 771 Rhˆone Glacier (Switzerland), 30 rhyolite, 587 rice (Oryza sativa), 129, 481, 665 rice straw, 536 Richards Bay (KwaZulu–Natal, South Africa), 378 rills, 514 Ring of Fire (Paciﬁc), 137 R´ıo de la Plata (Argentina), 489 R´ıo Matanza (Argentina), 489 R´ıo Negro (Argentina), 495 R´ıo Paran´a (Argentina), 489 R´ıo Salado (Argentina), 489 riparian, 771 riparian buffers, 682 riparian ecosystems, 427–432, 439, 714 riparian vegetation, 403 riparian woodlands, 271 risk-assessment, 661, 662, 759 river channel, 25 river meanders, 229 river red gum (Eucalyptus camaldulensis), 429 roadbeds, 512 roads (see also vehicle tracks), 173, 278, 316, 460, 462, 511, 589, 593 roadsides, 415–417 Robinson Crusoe paradox, 744, 771 rock hyrax (Procavia capensis), 56 rock outcrops, 587, 589, 590, 594, 595, 599 rock slides (see also landslides), 20, 29 Rock Valley (Nevada, USA), 316 Rocky Mountain National Park (USA), 69 Rocky Mountains (North America), 48, 50, 54, 55, 58, 59, 63, 66, 67, 72, 78, 80, 84, 86, 88, 165, 171, 428, 429, 456, 457, 751 rodents, 69, 70, 169, 213, 276, 319, 398, 498, 529, 580, 639, 640, 647, 681 – burrowing, 577 – – site effects, 580, 581 – heteromyid, 645 Rodrigues Island, 420, 636

859 roller-drum chopping (RDC), 550 Rome (Italy), 1, 2, 275 rookeries, 54 root-borers, 254 roots, 23, 25, 227, 230, 244, 246, 555–557, 561, 562, 564, 565, 615, 620, 622, 626 – and gaps, 212 – and shoot competition, 626 – biomass, 555–557, 560, 561, 563, 564 – breakage, 25 – competition, 617, 620 – depth, 559 – distribution, 555, 557 – exudation, 523 – growth, 523 – hairs, 525 – turnover, 523 – uptake, 548 Ross Sea, 55, 91 ruderal, 772 ruderal species, 488 runoff, 513, 528, 565 runoff rates, 389 Russia (see also Soviet Union), 57, 63, 65, 74, 86, 93, 140, 142, 456, 460, 695 Russian Arctic, 62 Russian olive (Eleagnus angustifolia), 431 rutting, 472 rye-grass, 127 Saami, 62 Sabah (Borneo, Malaysia), 234, 473–475 safe sites, 142, 144, 145, 151–154 sagebrush (Artemisia spp.), 417, 562, 573 sagebrush steppe, 439, 547, 556–560, 562, 564, 565 saguaro (Carnegiea gigantea), 310, 317, 324 Sahara Desert (northern Africa), 308, 314, 323, 324 Sahel Region (Africa), 292, 312, 323, 324 Sakurajima (Japan), 140, 142, 143, 148, 154–156 Salamanca (Spain), 294 salinas, 331 saline grassland, 574 saline water table, rising, 677 salinity, 332, 335, 339, 340, 342–345, 349–351, 355, 588 salinity tolerance, 431 salinization, 128, 276, 281, 282, 674 saliva, 578 salt, 264 – adaptation, 344, 349 – tolerance, 356 salt-marsh snails (Melampus bidentatus), 349 salt marshes (see also wetlands), 332, 333, 336, 338, 340–342, 344, 345, 347–353, 355, 356, 403, 438, 614, 618, 622 salt-tolerant species, 677 salt water intrusion, 201 Salta (Argentina), 495 saltbush (Atriplex spp.), 417 saltcedar, see tamarisk Saltillo (Coahuila, Mexico), 312

860 Samarinda (East Kalimantan, Borneo), 226 Samoa, 420, 636, 638 San Antonio de Areco (Argentina), 490 San Diego (California, USA), 529 San Francisco (California, USA), 2 San Joaquin Valley (California, USA), 278 San Juan (Argentina), 495 San Luis (Argentina), 489 sand, 29, 505 – coarse, 536 sand and silt deposition, 29 sand birch (Betula pendula), 406 sand dropseed (Sporobolus cryptandrus), 648 sand-dune succession (see also succession), 532 sand dunes (see also dunes), 316 sand lovegrass (Eragrostis trichodes), 648 sand pine (Pinus clausa) (see also pines), 391 sand shinnery oak ecosystem (SSO), 647–650 sand shinnery oak (Quercus havardii), 635, 647 sandhills, 297, 391 sandur, 21, 23, 28 sandwort (Minuartia spp.), 374 Santa F´e (Argentina), 489 Santiago (Chile), 279 Santorini (Greece), 137 S˜ao Carlos (Venezuela), 225 sap-suckers, 254, 257, 260, 262, 266 saprobic microfungi, 526 saprophytes, 532 saprophytic fungi, 525 Sarawak (Malaysia), 225, 229 sarcophagid ﬂy larvae, 647 Sardinia (Italy), 365 Saskatchewan (Canada), 620, 622, 623 satellite population, 718 Saudi Arabia, 5 savanna/forest ecotone, 287 savanna-type vegetation, 499 savannas, 772 savannas (see also prairies), 5, 14, 15, 271, 274, 275, 279, 287–303, 331, 355, 421, 425, 438, 439, 477, 576 – climate, 287, 288 – deﬁnition, 287 – dehesa, 294 – distribution, 287 – disturbance, 288 – – effects on above-ground net primary production (ANPP), 295–303 – East African, 290 – model of disturbance effects, 302 – oak, 294 – precipitation gradient, 287–303 – soil, 288 – structure, 287, 288 – tropical, 296 saw grass (Cladium jamaicense), 351, 353, 355 saw-grass prairie, 433 saw-log harvesting (SAW), 547–554 sawdust, 536

GENERAL INDEX sawﬂy (Perga afﬁnis), 264 scale, 139, 260, 708, 710–712, 716 – spatial, 3, 4, 385, 386, 708 – temporal, 3, 4, 708 scale of disturbance, 334, 338, 340, 343, 344, 347, 354, 355 Scandinavia, 17, 57, 88, 169, 171 scarabs [scarabaeids] (Anoplognathes spp.), 264, 641, 647 scarifying, 680 scarlet oak (Quercus coccinea), 259 scenarios for agricultural development, 733, 734 Schoolroom Glacier (Wyoming, USA), 28 scientiﬁc activity, 22 scientists, effect of, 30 sclerophyll ecosystems, dry, 264 sclerophyll forests, 280 sclerophyllous leaves, 271 sclerophyllous vegetation, 573, 772 sclerotia, 530 scorched earth policy, 385 scoria, 138, 140, 142, 144, 148 Scotch broom (Cytisus scoparius), 422 Scotia Arc (Antarctica), 52 Scotland (UK), 163, 188 Scots pine (Pinus sylvestris) (see also pines), 549, 579 scouring, 339, 351, 356 – by ice, 335, 336, 342 scree, 51 scrub, 421 scurvy grass (Cochlearia pyrenaica), 374 sea ice, 50, 89 sea-level changes, 332, 334, 337, 352 sea mammals (see also mammals), 44 sea thrift (Armeria maritima), 374 sea waves, 137 seabirds, 68 seagrass, 354 seals, 30, 43, 54, 58, 64, 70, 89–91, 337, 751 seasonal phenology, 531 seasonality, 469, 472 secondary colonizers, 27 secondary disturbance, 32 secondary forests, 471, 476–478, 481, 772 secondary metabolites, 491, 500 – production of, 493, 495, 500 secondary production, 92 secondary succession (see also succession), 27, 137, 139, 149, 611, 616, 751, 752, 772 sedges, 60, 61, 68, 82 sediment (see also soils), 331, 335, 337, 338, 343, 345, 347, 353, 355 – burial, 341 – deposition, 334–337, 339, 342, 343, 349, 353, 354, 356 – ﬂow, 20 sedimentation, 131, 335, 354, 432, 441 – effects of structural barriers, 337 seed banks, 140, 205, 212, 213, 236, 237, 240, 342, 344, 348, 349, 352 seed-eaters, 254 seed-tree cuts, 461

GENERAL INDEX seedbed, 680 seedlings, 26, 343, 344, 348–351, 356 – emergence, 496 – establishment, 562, 563 – mortality, 28 – regeneration, 545 – survival, 25 seeds, 26, 344, 461, 537 – adaptation, 349 – dispersal, 26, 212, 213, 318, 331, 343, 344, 348, 349, 354, 431, 442 – – by birds, 498 – – by humans, 593 – – by vertebrates, 213 – – by wind, 213 – dormancy, 26, 212, 213 – ecology, 192, 212, 213 – pads, 212, 213 – pathogens, 212 – pool, 190 – predation, 211–213, 619, 681 – rain, 213, 602 – scariﬁcation, 573 selection-harvest regimes, 461 selection system, 470 selective cutting, 470, 772 selenium, 369 semi-arid [deﬁnition], 308 semi-arid habitats, 428, 531 semi-natural community, 498 seral species (see also succession), 154 sere, 772 Serengeti National Park (Tanzania), 291, 294, 295, 297, 556, 558, 560, 578 Seri Indians, 323 serotiny, 164–167, 169, 772 serpentine grassland, 421 serpentine soils, 278, 772 set-aside programs, 666 severity of disturbance (see also disturbance), 140, 191 sewage, 65 sexual reproduction, 26 Shaba Province (Zaire), 372, 374 shade-intolerant species, 236, 237, 240, 242, 243, 472, 478 shade tolerance, 202–208, 211–214, 236 shade-tolerant species, 236, 242, 243, 472 shading, 594 shear, piling, disking (SPD), 550 sheathbills (Chionis spp.), 65 sheep (Ovis spp.), 58, 60, 61, 83, 92, 264, 280, 281 – feral, 443 shelterwood cuts, 461 shifting cultivation, 467, 665 shifting mosaic, 223, 755 shivering, 43 Shoalwater Bay Training Area (Queensland, Australia), 392, 393 shoot competition, 620 short grass prairie, 416, 418, 421, 423 short grasses, 557

861 shrink–swell clays, 506 shrub-grassland, 533 shrub steppe, 557, 562, 564 shrublands, 14, 15, 416, 423, 438, 439 – Mediterranean, 5 – post-ﬁre productivity, 573 shrubs and windstorms, 201, 207, 208 Siberia, 56, 63, 163, 166, 167, 174 Siberian ﬁr (Abies sibirica) (see also ﬁr), 163, 167, 168 Siberian larch (Larix sibirica), 163, 167 Siberian silkworm (Dendrolimus sibiricus), 169 sideoats grama (Bouteloua curtipendula), 574 siderophores, 523 Sierra Nevada (California, USA), 67, 457, 459 Sierra region (Ecuador), 125 Signy Island (Antarctica), 54, 74 Sikes Act, 391 silica, 505 silt, 29, 505, 528 silt-loam, 127 silver, 279, 369 silvicultural treatment and wind damage, 200 silviculture, 167, 460, 469, 478, 481, 772 sink, 331, 718 siphonopterans, 647 site fertility, 553, 554, 561 site homogeneity, 527 site preparation, 460–462 site protection, 535 site recovery, 467 Sitka spruce (Picea sitchensis) (see also spruce), 454, 548, 550, 600, 601 size-frequency distribution, 26 size separates, 505 ski runs, 69 skuas (Catharacta spp.), 62, 65, 68, 70, 91 slash, 548–553 slash piling, 554 slash pine (Pinus elliottii) (see also pines), 552 slopes, 561 – angle, 513 – length, 513 – position, 716 – stability, 22 sloughing, 523 Slovakia, 189 slow-release irrigation, 536 slugs, 637, 638 small mammals (see also mammals), 388 snails, 131, 349, 353, 432, 637, 643 Snake River, 425 snakeweed (Gutierrezia sarothrae), 389 snapping, 232 snow, 25, 45–50, 307, 714 – creep, 30 snow goose (Anser caerulescens), 337, 751 snow-line, 39 snowbeds, 22 snowﬁelds, 29

862 snowshoe hares (Lepus americanus) (see also rabbits), 55 social capital, 703 social impacts, 662 socioeconomic development, 727 socioeconomic pressure, 725, 730–734 sodium, 587 soil-crust slurry, 536 soils, 200, 261–264, 266, 331, 332, 335–337, 339, 342, 343, 349, 351, 353, 355, 398, 399, 401, 402, 443, 471, 472, 480, 499, 503–515, 753 – aeration, 318, 506, 507, 512 – aggregation, 505, 506, 561, 564 – air, 507, 508 – arthropods, 523 – bulk density, 510 – cation exchange capacity, 594 – clay, 278 – CO2 emission, 496 – compaction, 388, 392, 393, 399, 401, 405, 461, 507, 509–511 – compression, 76–79 – cooling, 512 – crusts, 324, 535 – degradation, 665, 677 – degradation and plant physiotypes, 493 – depth, 128, 561 – destabilization, 309 – development, 17 – disruption, 444 – disturbance, 246, 508, 526 – drainage, 31 – erosion (see also erosion), 1, 388, 389, 513–515, 674, 678, 682 – – effects of ﬁre, 574, 577 – – post-ﬁre, 575, 576 – fauna, 443 – fertility, 262, 560, 564, 565 – formation, 23, 125 – freezing, 506 – friability, 505 – gas phase, 507, 508 – hardpans, 509 – heterogeneity, 288 – horizons, buried, 24 – hypoxia, 512 – impacts, 386, 387 – inﬁltration, 388 – invertebrates, 522 – leaching, 524 – liquid phase, 507 – loosening, 509–511 – management, 536 – microbes, 521–523, 526, 530, 560, 564 – mixing, 23, 26 – moisture, 20, 262, 556–558, 560–565 – nitrogen, 553, 554, 594 – organic carbon, 534, 556, 562 – organic matter (see also organic matter), 311, 510, 514, 531, 547, 549, 554–557, 594, 601 – – inputs, 554, 558 – organisms, 471, 472

GENERAL INDEX – oxygen, 507, 512, 562 – pH, 575 – physical model, 503–508 – pitting, 389 – pores, 527, 565 – reducing conditions, 512 – resources, 626 – seed stores, 680 – serpentine, 278 – size separates, 504 – slaking, 505 – solid phase, 504–507 – stability, 590 – strength, 507, 511, 512 – structural decline, 680 – structure, 388, 504 – surface crusts, 506 – temperature, 458, 460, 496, 558, 560, 564 – texture, 504, 505, 558, 564, 590, 594, 772 – thawing, 506 – tilth, 505 – volume, 562 – water, 21, 505, 507, 590, 594 – wind disturbance, 200, 209–212 solar radiation, 39, 42, 399 solid waste, 63, 64, 85, 86 soliﬂuction, 21, 24, 30, 47, 51, 772 Sonbhadra (India), 295 Sonora (Mexico), 310 Sonoran Desert (Mexico, USA) (see also deserts), 278, 308, 310, 311, 317, 320, 321, 323, 324, 416, 418, 423, 425, 428, 431, 432, 749 sorghum, 129 Sørsdal Glacier (Antarctica), 31 South Africa, 233, 271, 272, 274, 278, 282, 317, 366, 416, 424–426, 430, 436, 439–442, 695, 759 – human history, 274 – Johannesburg, 366 – Witwatersrand, 366 South America, 14, 15, 17, 53, 56, 57, 90, 125, 128, 188, 287, 291, 300, 308, 439 South Carolina (USA), 190, 195, 200, 201, 203, 210, 334, 343, 618, 622 South Dakota (USA), 294, 295, 297, 298, 423 South Georgia (Subantarctica), 55, 59, 69, 70, 74, 79, 83, 91 South Island (New Zealand) (see also New Zealand), 188 South Pole, 39, 65 South Shetland Islands (Antarctica), 48, 53, 77, 91 South Vietnam, 385 Southeast Asia, 225–229, 233, 695 southern beech (Nothofagus spp.) (see also New Zealand and Chile), 147, 188, 204, 206–208, 211, 287 southern desert scrub, 386 southern giant petrels (Macronectes giganteus), 59 southern grasshopper mouse (Onychomys torridus), 389 Southern High Plains (USA), 647 Southern Ocean, 59, 66 Southern Ocean Whale Sanctuary, 93 southern pine beetle (Dendroctonus frontalis), 259–261

GENERAL INDEX sovereignty, 659 Soviet Union (former), 57, 130 – steppeland rodents, 580 soybean (Soja max), 129, 490, 494, 512 spacing, 461 spadderdock (Nuphar luteum), 336 Spain, 365, 533, 664 Spanish colonization, 277–279 sparrows, 390 spatial heterogeneity, 3, 14, 15, 461, 527, 594, 597, 600, 760, 772 spatial organization, 18 spatial patterns, 526 spatial variability, 586, 595 SPD, see shear, piling, disking specialization, 626 species – composition, 301, 488 – constancy, changes in, 496 – density, 488 – diversity (see also richness under species), 1, 28, 140, 147, 271, 272, 340, 350, 351, 462, 464, 487, 500 – endangered (see also extinction), 392, 699 – extinction, 404, 407 – feedback, 435–444 – increment, rate of, 499 – interactions, 493, 675 – reintroduction, 681 – replacement, 497 – richness (see also diversity under species), 147, 148, 152, 276, 340, 343–345, 350, 379, 380, 489, 491–494, 499 – – control of seed germination by soil physical characteristics, 496 – turnover, 147, 148, 532 spiders, 153, 258, 593, 635, 636, 639, 640 spike dropseed (Sporobolus contractus), 318 Spirit Lake (Mount St. Helens, Washington, USA), 150 spontaneous vegetation, 499 spore velocity, 529 spores, 527 sporocarps, 529 spring sandwort (Minuartia verna), 374 sprouting, 237, 238, 242, 317 spruce bark beetle (Dendroctonus ruﬁpennis), 171, 260, 455, 751 spruce budworm (Choristoneura fumiferana), 167, 169–171, 175, 580, 750 spruce forests, 31 spruce (Picea spp.), 167–173, 180, 182, 205, 454, 456, 548, 550, 577, 580, 600, 601, 710, 751 – age–structure diagrams, 182 – northern limits, 169 – types, 163, 167–172 squash (Cucurbita pepo), 59 squatter settlers, 665 St. Lawrence Island (Beringia), 58 St. Pierre (Martinique), 1 St. Vincent (Caribbean), 241 stability, 14, 15, 40, 718, 748, 754, 755, 757, 772 stabilization, 27, 31, 586 – during succession, 31

863 stabilized dunes, 532 stakeholders, 690, 695, 696, 700 stand development, 460 stand initiation, 460 stand-replacing disturbances, 453 standing mortality, 227, 230, 233, 234 state-transition processes, 675–677 Stegholbreen (Norway), 20 Steller’s sea cow (Hydrodamalis gigas), 58 stem exclusion, 29, 461 stem girdling, 546, 552, 553 steppe, 294–297 Stepping Stone Island (Antarctica), 45 stewardship, 663, 664 Stillwater (Oklahoma, USA), 297 stinkwood (Celtis spp.), 380 stochastic, 772 stochastic events, 149, 154, 598 Stone Age, 280 stone marten (Martes foina), 398 stone stripes, 24, 25 Storbreen (Jotunheimen, Norway), 28, 30 storm surge, 335 storms, 425, 440 strategic environmental assessment, 668 strategies, plant, 26 straw, 536 strawberry guava (Psidium cattleianum), 443 stream-water chemistry, 548 streams, 711, 714, 715 stress, 39, 140, 585, 592, 596, 614, 627, 678, 679, 714, 772 stress tolerance, 40, 488, 499, 500 stressors, 679 strip cutting, 547, 554 strip mining (see also mining), 371, 376–378, 526, 588, 772 structural impoundment, 335 structure of ecosystems, 675 Studebaker Ridge (Washington, USA), 148 studies, long-term, 194, 196 stumps, 202, 210, 211 Styggedalsbreen (Jotunheimen, Norway), 24 sub-alpine vegetation, 30, 204 sub-fossil, 20 sub-tropical marsh/woodlands, 433 subalpine, 416 subalpine ﬁr (Abies lasiocarpa) (see also ﬁr), 458 Subantarctic, 55, 88 subglacial sediments, 27 subhumid grassland, 418 subsidence, 76, 337, 354, 588 subsidies, 691, 692 subsoiling, 510 substrates (see also soils), 29, 342, 343, 356, 523 – chemical changes, 29 – deglaciated, 18, 19, 28 – disturbance, 23, 24 – mixing, 23 – mobility, 25 – modiﬁcation, 23, 24

864 substrates (see also soils) (cont’d) – stabilization, 26, 28, 441 – textural change, 29 – volcanic, 28 substratum succession, 531 subtropical broadleaf forests, 206 subtropics, 469, 772 succession (see also primary and secondary succession, recovery, restoration), 2, 14, 15, 17, 75, 79–83, 137, 143, 144, 146, 147, 149, 151–154, 161, 169–171, 175, 191, 192, 203, 213, 258, 259, 261, 266, 279, 314, 319, 332, 333, 336, 344, 347, 348, 351, 352, 356, 397, 405–407, 409, 425, 427, 441, 444, 477, 487–501, 531–535, 558, 559, 561, 595, 675, 679, 717, 748, 760–763, 772 – after windstorms, 192, 201 – allogenesis, 30 – and disturbance, 31 – and productivity, 31 – animals, 593 – autogenesis, 30 – climax, 595, 600 – Connell and Slatyer models, 596 – convergence, 30 – cyclic, 585 – deﬁnitions of, 585, 586 – disturbance in, 17–32 – divergence, 27, 30, 31 – dynamics, 760 – geoecological model, 597 – grassland, 490 – Grime, 596 – herbivory, 169 – heterotrophic, 488 – initial ﬂoristics, 595, 598 – models, 487, 595, 602 – natural, 499 – nutrient distribution, 557 – on recently-deglaciated terrain, 17–32 – organic matter inputs, 564 – organismic analogy, 595 – pastures, 496 – primary (see also primary succession), 17, 27–31, 585, 586, 591, 595, 597, 713 – processes, 488, 498–500 – rate of, 27–29, 140–147, 150 – regressive, 499 – roots, 557 – secondary (see also secondary succession), 27, 488, 489, 493, 499, 586, 595, 597, 598, 713 – species replacement, 585 – stabilization during, 27, 31 – temperate forests, 192 – theory, 151, 585, 595–600 – Tilman’s resource-ratio model, 596 – trajectory of, 27, 29, 30 – trends, 493, 495, 500 – vital attributes, 596 Succulent Karoo (South Africa), 271 succulents, 308, 310, 311

GENERAL INDEX Sudan, 365, 470, 665 Sudbury (Canada), 369 sugar beet (Beta vulgaris), 515 sugar fungi, 532 sugar maple (Acer saccharum), 203–205, 403, 548 sulfur, 523 sulfur dioxide, 401 Sumatra, 691 summer rains, 312 sunﬁsh, 432 sunﬂower (Helianthus annuus), 490 supercooling, 43 supraglacial ecosystems, 27 surface depressions, 561 surface detention, 561 surface heterogeneity, 680 surface texture, 561 surface wash, 23 surface water, 390 Surinam, 469 Surtsey (Iceland), 83, 137, 142, 143, 147, 148, 154 survival, 261 survival of wind damage, 199, 200 survivors, 151 sustainability, 469, 477, 718, 723–745 – epistemological predicament of, 741, 742 – myths, 736–739, 743, 744 sustainable development, 482, 660, 663, 669, 772 sustainable management, 469 Svalbard (Norway), 28, 30, 58, 69, 88 Svartisen (Norway), 30 Svellnosbreen (Jotunheimen, Norway), 25 swamp cyrilla (Cyrilla racemiﬂora), 355 swamps (see also wetlands), 190, 331–336, 338, 339, 341, 343, 345, 347, 350, 352, 355, 772 Swaziland, 469 Sweden, 26, 74, 75, 164–166, 168, 172–174, 177, 179, 188, 204, 548–550, 553, 618 sweet birch (Betula lenta), 552 sweet thorn (Acacia karoo), 380 Switzerland, 56, 59 symbiosis, 523, 529 symbiotic bacteria, 521 synergy, 455, 457 Syracuse (New York, USA), 372 tabonuco (Dacryodes excelsa), 257, 638 tabonuco forests (see also Luquillo Experimental Forest), 642, 643, 749 Tabor oak (Quercus ithaburensis), 276 Tafua Rain Forest Reserve (Samoa), 636–638 Tahiti, 426 Tai (Ivory Coast, Africa), 231 taiga, 163, 167, 772 Taiwan, 188, 201 Takla Makan Desert (China) (see also deserts), 316 tallgrass prairie, 340 talus slopes, 75 tamarisk (Tamarix spp.), 324, 431, 432, 434, 445

GENERAL INDEX Tanana River (Alaska, USA), 601 Tanzania, 421, 667 tap roots, 25 tapirs (Tapirus spp.), 473 tarbush (Flourensia cernua), 318 tardigrades, 523 Tareya (Russia), 74 Tasman Glacier (New Zealand), 30 Tasmania (Australia), 188 taxes, 691, 692, 699–701 taxonomic diversity, 526, 756 teak (Tectona grandis), 470, 480 tectonic forces, 3 temperate forests (see also forests), 187–214, 428, 433, 453–464 – coniferous, 429 – deciduous, 416, 433 – harvesting, 459–464 – rain forests, 420 temperate grassland, 418 temperate zone, 772 temperature, 257, 258, 264, 310, 398, 399, 401, 402, 405 – environmental conditions, 258 – ﬂuctuations, 42–44, 552 – high, 310 – low, 310, 311 – variation, 310 temporal and spatial patterns, 690 temporal changes, 492 temporal disturbances, 693 temporal heterogeneity, 597 temporal pattern, 491 temporal variability, 586 tenebrionids, 316 Tenerife (Canary Islands), 147 Tennessee (USA), 208 tephra, 137–144, 148–151, 153, 154, 156, 772 termites, 264, 474, 475, 593, 645, 646 – soil-feeding, 475 – wood-feeding, 475 terra ﬁrme, 225 terrain-age, 28 terrestrial invasion (see also invasion), 521 Teton Range (Wyoming, USA), 28 Texas high plains (Texas, USA), 298 Texas (USA), 190, 195, 196, 201, 203, 334, 348, 378, 420, 425, 428, 432, 433, 635, 647, 648 Texas wintergrass (Stipa leucotricha), 574 textural classes, 505 Thailand, 334, 470 – grassland, 573 thatching ants, 350 Thera (Greece), 156 thermokarst, 46, 47, 62, 78, 81, 167, 772 thermokarst activity, 577 Thessaloniki (Greece), 294, 298 thickets, 598, 599 thinning, 461 thrashers, 390 threatened and endangered species (see also extinction), 392

865 three-phase soil model, 503–508 thresholds, 14, 15, 70, 661, 676, 678, 748, 756, 759, 772 thunderstorms, 189, 190, 196, 197, 201, 210, 225 Tibet, 40, 44, 47, 56–58, 60, 61, 63, 88 Tibetan Plateau, 39, 40, 44, 50, 56, 58–61, 65, 67, 78, 80, 82, 84, 88, 93 ticks, 639 tidal forests, 332 tides, 331, 332, 335, 338, 340, 343, 348, 353 Tien Shan (China), 28, 56, 67 Tierra del Fuego (Argentina), 432 Tigris River (Turkey, Iraq), 2, 663 till, 18 tillage, 506, 535 – systems, 493, 494 tillage erosion, 513 tillage pans, 510 timberlines (see also tree line), 143, 147 – depression, 147 titi (Cliftonia monophylla), 355 toad (Bufo bufo), 407, 645 Tokyo (Japan), 405, 407 Tolbachik (Kamchatka, Russia), 140 tolerance, 596–598 tolerant forest management, 665 topographic position, 231 topography, 195, 200, 227 – wind damage, 194, 199, 200 topsoil, 534 topsoil retention, 535 tornadoes, 4, 5, 50, 189, 190, 195, 196, 200, 201, 203, 205, 208–211, 213, 713 tourism, 22, 57, 58, 62, 69, 70, 92, 277, 324, 673 Toutle River lahar (Mount St. Helens, USA), 145 toxic contaminants, 390 toxicity, 525 toxins, 331, 335, 337, 341–343, 345 tracked vehicles (see also off-road vehicles), 385, 387, 393 trade-offs in technological development, 726, 736, 737, 744 trafﬁc pans, 510 tragedy of change, 740, 743, 744 training activities (military), 385–387, 392, 393 trampling, 555, 556, 558, 561, 562, 564, 644 transboundary impacts, 659, 667, 668 transhumance, 59, 659 translocation, 574 translocation of assimilates, 573, 574 Transmigration Programme, 667 transplanting, 537 transplants, 619, 621, 623 transportation, 4, 10, 62, 63, 587, 589 trawling, 59 tree falls, 3, 4, 172, 635, 641–643, 749, 763 tree ferns, 209 tree line (see also timberlines), 39, 50, 51, 57, 59, 88, 171, 174 tree-ring analysis, 193 tree-rings, 25 trees, 88, 619, 622, 626 – anchorage, 230–232

866 trees (cont’d) – crown size, 228 – defoliation, 635 – height, 228, 229 – invasion, 88 – mortality, 234, 453 – plantations (see also plantations), 469 – wind damage – – effect of tree size, 199, 201 – – effect of tree species, 199–201, 210, 213 – wind-training/shearing of, 25 trenching, 546 trichomes, 563, 565 trim-line, 20 Trinidad, 469 trophic level, 14 Tropic of Capricorn, 392 tropical cyclones, see cyclones, hurricanes tropical-forest landscape, 467 – anthropogenic disturbance, 467–469, 471, 481 – disturbance, 467–469, 471, 474, 477, 480 tropical forests, 212, 213, 223–246, 271, 416, 419, 635–638 – dry, 227 tropical grasslands (see also savannas) – ﬁre effects, 573, 574 tropical mountains, 18, 588 tropical systems, 679 tropical vegetation, 386 tropical zone, 772 tropics, 5, 469 tsunamis, 137, 138, 334 Tucson (Arizona, USA), 310, 312 tulip poplar (Liriodendron tulipifera), 552 tundra, 5, 52, 55, 73, 82, 89, 90, 171, 174 Tunisia, 436, 441 Tunstead Limestone Quarry (Derbyshire, England), 373 Turkey, 142 turnover due to wind disturbance, 197, 198 turnover rates, 147, 148, 535 tussock vegetation (see also tundra), 45, 52, 71, 73, 75, 80, 82, 83, 89–91 Tyndall Glacier (Kenya), 30 typhoons, see cyclones, hurricanes

Uganda, 170, 475 Ukraine, 189 ultraviolet radiation, 41, 42, 65, 66, 85, 87, 92, 93 understory, 457 – plants, 206 – – windstorms, 201, 203, 204, 206–209, 213 – reinitiation, 29, 461 undisturbed habitats, 426, 427, 434, 444 ungulates, 169, 547, 554, 557, 565 United Kingdom, see Great Britain United Nations, 392 United States National Forest Service, 690 United States of America, 2, 69, 125–127, 137, 142, 148, 172, 189, 196, 214, 259–262, 308, 309, 324, 334, 366, 387, 403,

GENERAL INDEX 432, 439, 453, 454, 459, 460, 499, 500, 573, 588, 589, 666–668, 691, 692, 694–697, 734, 738, 759 unrest, 667 unsuitable ﬁre regimes, 678 unsustainability, 725 uprooting, 209–211, 231, 232, 243 uptake, 546–550, 553, 554 – nitrogen, 548, 551, 560 – nutrients and water, 552 – rates, 545 Ural Mountains (Russia), 189 uranium, 369 urban, 4, 14, 15, 589, 673 – climate, 399–401 – development, 673 – ﬂora, 403–407 – forests, 398, 401 – habitats, 397 urbanization, 1, 2, 15, 272, 276–279, 334, 754 urea, 43 urine, 578, 581 Uruguay, 291 Urumqi River (China), 67 Utah (USA), 123, 124, 320, 417, 418, 420, 423, 428, 442 utilization of native ecosystems, 673 valley glacier, 20 valley grassland, 421 valued ecosystem components (VEC), 694 variation in temperatures, 310 vascular plants, 399, 405 vegetation, 397–399, 401, 402, 405, 408, 534 – clearance, 682 – mapping, 17 – patterns, 147 – structure – – precipitation, 287, 288 vegetative reproduction, 26, 79 vehicle tracks (see also off-road vehicles, tracked vehicles), 62, 83 vehicles, 69, 73, 80 Venezuela, 5, 225, 230, 232, 233, 244, 255, 423, 468, 470, 477 Veobreen (Jotunheimen, Norway), 23 Veracruz (Mexico), 532 Vermont (USA), 190, 205, 700 vertebrates, 172, 213 vervet monkeys (Cerceopithecus aethiops), 379 Vestfold Hills (Antarctica), 31 Victoria Land, 64 vicu˜nas (Vicugna vicugna), 58, 60, 92 Vienna (Austria), 405 Vietnam, 385, 470, 660 vines, 227, 233, 234, 237, 241 – wind damage, 200 violets (Viola spp.), 372, 374 Virginia (USA), 190, 336, 378, 425, 428 viruses, 258 vital ecosystem attributes, 676 vital landscape attributes, 682 vivipary (see also reproduction), 26, 43, 772

GENERAL INDEX volatilization of nutrients (see also nutrients), 575 volcanic substrates (see also volcanoes), 28 volcanoes, 1, 3, 4, 15, 47, 51, 52, 64, 65, 137, 139, 143, 153, 156, 272, 334, 586–589, 594, 599, 713, 750, 757 – differential effects, 149, 150 – disturbance regimes, 139–141 – effects of weeds, 154, 155 – effects on animals, 150 – impacts, 137–139 – mechanisms of recovery, 150–153 – predictability, 153, 154 – primary succession (see also primary succession), 141 – volcanic hazard management, 155, 156 voles, 640 Vorkuta (Russia), 57 Vulcan, 1 Wales (UK) (see also Great Britain), 430, 431 Walker-on-Tyne (England), 366 walkingsticks, 257, 637 wallows, 713 walrus (Obodenus spp.), 58, 90 Warren Woods (Michigan, USA), 189 Washington (USA), 29, 170, 190, 203, 211, 416, 420–422, 427, 454, 588 – forests, 574 wasps, 639–641, 645 waste products, 4 wastelands, 497–499 water, 258, 262–264, 311–314 – availability, 311–313 – dispersal, 143 – erosion, 311 – ﬂux, 548, 553, 554 – holding capacity, 561, 562 – inﬁltration, 126, 389, 459 – potential, 614, 619 – predictability, 311–313 – quality, 546, 547 – – forest harvesting, 547, 548 – resources, 692 – retention, 561 – supply, 29 – table, 26, 29, 30, 576 water bears, 523 water buffalo (Bubalus bubalis), 443, 681 water hyacinth (Eichhornia crassipes), 432 water management, 536 water movement, 528 water-repellent layers, 576 waterfowl, 335, 336, 343, 354, 390 waterlogging, 128, 227, 231, 232 wave action, 334–336, 351, 354 wave regeneration, 172 weather, 264 Weddell Sea (Antarctica), 59 weeds (see also invasive species), 88, 154, 407, 488, 489, 491, 492, 589, 677, 681, 772 – community, 487, 492, 500

867 – control of, 681 – dynamics, 499 – ﬂora, 487 – germination, 496 Wellton (Arizona, USA), 312 West Africa, 224, 226, 229, 231, 233, 576 western hemlock (Tsuga heterophylla), 454, 600 wet-ﬂush effect, 29 wet meadow, 79 wetland management, see management of wetlands wetlands, 14, 15, 55, 331–356, 391, 432, 433, 438, 617, 619, 623, 624, 682, 772 – beaver activity, 580 – disturbance types, 335–337 – forests, 226 – history of disturbance, 333, 334 – loss, 334, 356 – riparian, 334 whales, 58, 59, 68, 90, 91, 93 whaling, 59, 70 Wharram Chalk Quarry (Great Britain), 372 wheat (Triticum aestivum), 59, 126, 129, 130, 281, 490, 515 wheatgrass (Agropyron spp.), 694 white ﬁr (Abies concolor) (see also ﬁr), 261 white-footed mice (Peromyscus leucopus), 389 white ironwood (Vepris lanceolata), 380 White Mountains (New Hampshire, USA), 51 white oaks (Quercus alba) (see also oaks), 259 white spruce (Picea glauca) (see also spruce), 545 white stinkwood (Celtis africana), 380 Whiteface Mountain (New York, USA), 172 whole-tree harvesting, see WTH wildlife, 131, 150, 462 wildlife management, 303, 336 willows (Salix spp.), 48, 55, 140, 144, 170, 531 wind, 3, 23–26, 50, 51, 123, 126, 132, 172, 262, 307, 314–316, 334–336, 340, 354, 454, 455, 528, 713, 716, 750 – action, 21 – damage, 199–201, 203, 210, 213 – dispersal, 143 – disturbance, 223 – in forests, 187–214 – shearing, 25 – speeds, 194–196, 200, 228 – training, 25 wind-blown salts, 31 wind-chill, 50 windrowing, 461 windstorm studies, methodology, 201 windstorms, 1, 224, 228, 260, 262, 266, 333, 334, 347, 352 – history, 193, 194 – mounds, 193 windthrow, 22, 26, 30, 31, 223, 235, 258, 335, 338, 347, 350 – mounds, 202 – pits, 202 winter dessication, 25 winter moth (Operophtera brumata), 444 winter rains, 312, 313 wiregrass (Aristida stricta), 333

868 Wisconsin (USA), 124, 170, 190, 193, 196, 197, 205, 415, 416 Witwatersrand (South Africa), 366 wolves (Canis spp.), 44 wood chips, 536 wood cutting, 30 wood strength, 192, 201, 204 woodlands, 29, 271, 272, 405, 406, 423, 424, 438, 439, 469, 676, 677 – acacia, 279 – dry sclerophyll, 271 – eucalypt, 271 – oaks (Quercus spp.), 274, 275, 278, 279 – riparian, 271 – sclerophyll, 279 – wet sclerophyll, 271 woodpeckers (Picoides spp.), 392 woodrats, 645 woody debris, 169 woody-plant encroachment, 288 woody plants, 617, 618, 624 World War II, 275, 281 wrack, 333, 335, 336, 338, 339, 345, 347, 349, 351, 356, 772 Wrangell Mountains (Alaska, USA), 75 WTH (whole-tree harvesting), 547, 549–554 Wyoming (USA), 28, 190, 212, 294, 296, 458, 527, 533

GENERAL INDEX xeric, 772 xeric species, 561 yaks (Bos grunniens), 60, 61 Yakutia (Russia), 62 yellow birch (Betula alleghaniensis), 203, 552 Yellowstone National Park (Wyoming, USA), 289–291, 296, 457–459 Yellowstone Plateau (Wyoming, USA), 713 Yenisey River (Russia), 63 Yucatan (Mexico), 226, 233, 235, 239–241, 244, 245, 636 Yukon Territory (Canada), 28–31, 165 Yuma (Arizona, USA), 312 Zagne (Ivory Coast, Africa), 231 Zambia, 366 zebra mussels, 432 Zhong Zhang (Tibet), 60 zinc, 365, 369, 375 zinc violet (Viola calaminaria), 374 zokors (Myospalax spp.), 56, 82 zonation, 340, 348 zoning, 701 zooplankton, 41, 68 Zulu cherry-orange (Teclea gerrardii), 380