Ecosystems of disturbed ground

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

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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). 

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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

18

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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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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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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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

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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

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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

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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).

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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

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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

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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

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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

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(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).

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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).

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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

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

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

67

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).

68

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

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

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).

69

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

70

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).

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

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

71

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

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

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).

STRESS AND DISTURBANCE IN COLD REGION ECOSYSTEMS

85

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

86

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

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).

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(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).

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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

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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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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

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Erosion (t/ha)

20

10

0 0

20

40

60

80

100

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% 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;

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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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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

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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).

128

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.

REFERENCES Aina, P.O., Lal, R. and Taylor, G.S., 1977. Soil and crop management in relation to soil erosion in the rainforest of western Nigeria. In: National Conference on Soil Erosion, Purdue University, Soil Erosion: Prediction and Control, Pub. 21. Soil Conservation Society of America, Ankeny, Iowa, pp. 75–84. Alexander, M., 1977. Introduction to Soil Microbiology. 2nd ed. Wiley, New York, 467 pp. Allen, W., 1994. The flood of 1993 may rank as the worst weather disaster in U.S. history. St. Louis Post Dispatch, St. Louis, Missouri. Allison, F.E., 1973. Soil Organic Matter and its Role in Crop Production. Elsevier, New York, 639 pp. Alonso, C.V. and Combs, G.S., 1989. Streambank Erosion Due to Bed Degradation. The American Society of Agricultural Engineers, St. Joseph, Michigan, 78 pp. Anderson, J.M., 1978. Inter- and intra-habitat relationships between woodland cryptostigmata species diversity and the diversity of soil and litter microhabitats. Oecologia, 32: 341–348. Anderson, J.M., 1988. Spatiotemporal effects of invertebrates on soil processes. Biol. Fertil. Soils, 6: 216–227. Atlavinyte, O., 1964. Distribution of earthworms (Lumbricidae) and larval insects in the eroded soil under cultivated crops. Pedobiologia, 4: 245–250. Atlavinyte, O., 1965. The effect of erosion on the population of earthworms (Lumbricidae) in soils under different crops. Pedobiologia, 5: 178–188. Barrow, C.J., 1991. Land Degradation. Cambridge University Press, Cambridge, 295 pp. Barrows, H.L. and Kilmer, V.J., 1963. Plant nutrient losses from soils by water erosion. Adv. Agron., 15: 303–315. Bennett, H.H., 1939. Soil Conservation. McGraw-Hill Book Co., New York, 993 pp. Bohac, J. and Pokarzhevsky, A., 1987. Effect of manure and NPK on soil macrofauna in chernozem soil. In: J. Szegi (Editor), Soil Biology and Conservation of Biosphere. Akademiai Kiado, Budapest, Hungary, pp. 15–19. Bruijnzeel, L.A., 1990. Hydrology of Moist Tropical Forests and Effects of Conversion: A State of Knowledge Review. Free University and UNESCO, Amsterdam, 244 pp. Byington, E.K., 1986. Grazing Land Management and Water Quality. American Society of Agronomy and Crop Science Society of America, Harpers Ferry, West Virginia, 323 pp. Chaney, K. and Swift, R.S., 1984. The influence of organic matter on aggregate stability in some British soils. J. Soil Sci., 35: 223–230. Clark, E.H., 1987. Soil erosion: Offsite environmental effects. In: J.M. Harlin and G.M. Bernardi (Editors), Soil Loss: Processes, Policies, and Prospects. Westview, New York, pp. 59–89. Cox, P.A., Elmqvist, T., Pierson, E.D. and Rainey, W.E., 1991. Flying foxes as strong indicators in South Pacific Island ecosystems: a conservation hypothesis. Conservation Biol., 5: 448–454.

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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

VOLCANIC DISTURBANCES AND ECOSYSTEM RECOVERY

141

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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Roger DEL MORAL and Sergei Yu. GRISHIN

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).

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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-

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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.

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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,

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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.

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(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

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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

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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.

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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

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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.

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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.

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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

Ola ENGELMARK

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

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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

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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

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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).

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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

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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

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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

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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

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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

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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

Sara L. WEBB

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).

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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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(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).

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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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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

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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

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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)

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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

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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

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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-

264

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

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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

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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

Philip W. RUNDEL

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

Philip W. RUNDEL

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),

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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.

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

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

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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

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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),

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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

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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

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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

James A. MacMAHON

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 occ