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Michel concludes 'with a vibrant call for an indis- pensable ... Prize winner 1986 in marine ecology. Quotation of the Jury (Chairman: John Gray, Oslo, Norway):.
Editor O. Kinne

O. Kinne, Editor

The Challenges of Biodiversity Science

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Michel Loreau

The Challenges of Biodiversity Science

Michel Loreau

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Published 2010 by International Ecology Institute 21385 Oldendorf/Luhe Germany

OTTO KINNE Editor

Michel Loreau

THE CHALLENGES OF BIODIVERSITY SCIENCE

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Introduction (Otto Kinne) Michel Loreau: A Laudatio (Anders Pape Møller)

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Publisher: International Ecology Institute Nordbünte 23, 21385 Oldendorf/Luhe Germany

Michel Loreau Department of Biology McGill University 1205, avenue Docteur Penfield Montréal Québec H3A 1B1 Canada

ISSN 0932-2205 Copyright © 2010, by International Ecology Institute, 21385 Oldendorf/Luhe, Germany All rights reserved No part of this book may be reproduced by any means, or transmitted, or translated without written permission of the publisher Printed in Germany Typesetting by International Ecology Institute, Oldendorf Printing and bookbinding by Druckerei Wulf, Lüneburg Printed on acid-free paper

Contents Introduction (Otto Kinne) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Michel Loreau: A Laudatio (Anders Pape Møller) . . . . . . . . . . . . . . . . . . . . . . XXVII Preface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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THE EMERGENCE OF BIODIVERSITY SCIENCE . . . . . . . . . . . . . .

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WHAT IS BIODIVERSITY? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A Diversity of Diversities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Taking the Measure of Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimating Species Richness Using Species Accumulation and Rarefaction Curves . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Measuring Species Diversity Based on Species Abundance Distributions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Measuring Species Diversity Using Non-Parametric Diversity Indices . Species Diversity in Space: α, β, and γ Diversity . . . . . . . . . . . . . . . . . . Species Diversity in Space: Species–Area Relationships . . . . . . . . . . . . Biodiversity beyond Species . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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BIODIVERSITY IN CRISIS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Balance between Origination and Extinction . . . . . . . . . . . . . . . . . . Heading for a Sixth Mass Extinction . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Causes of the Current Biodiversity Crisis . . . . . . . . . . . . . . . . . . . . .

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THE VALUES OF BIODIVERSITY . . . . . . . . . . . . . . . . . . . . . . . . . . . . Why Does Biodiversity Matter? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use Values of Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecological Values of Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Intrinsic and Cultural Values of Biodiversity . . . . . . . . . . . . . . . . . . . . . . Economic Value of Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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FRONTIERS OF BIODIVERSITY SCIENCE . . . . . . . . . . . . . . . . . . . . Earth’s Exploration Is Still to Be Done . . . . . . . . . . . . . . . . . . . . . . . . . . Discovering Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Monitoring and Predicting Biodiversity Changes . . . . . . . . . . . . . . . . . . Assessing the Ecological and Social Consequences of Biodiversity Changes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conserving and Managing Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . Building Integrated Predictive Models of Biodiversity Changes . . . . . . .

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LINKING BIODIVERSITY SCIENCE AND POLICY . . . . . . . . . . . . . The Strained Relations between Science and Policy . . . . . . . . . . . . . . . . International Scientific Assessments as a Means to Strengthen the Science–Policy Interface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Building an International Scientific Assessment Mechanism for Biodiversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . BUILDING A NEW RELATIONSHIP BETWEEN HUMANITY AND NATURE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Limits of Biodiversity Conservation . . . . . . . . . . . . . . . . . . . . . . . . . Reintegrating Humans into Nature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biodiversity and Environmental Ethics . . . . . . . . . . . . . . . . . . . . . . . . . .

References

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INTER-RESEARCH SCIENCE PUBLISHER Journals Books EE Books Top Books ESEP Books Eco-Ethics International Union Orders

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Introduction Otto Kinne International Ecology Institute, Nordbünte 23, 21385 Oldendorf/Luhe, Germany

About the Book Professor Michel Loreau — winner of the ECI Prize 2002 in Terrestrial Ecology — addresses one of the most significant topics of our time: the importance of biodiversity for the functioning of ecosystems and the need to protect diversity for the sake of present and future human generations. Michel’s work includes ground-breaking modeling studies at the community level, which he uses to make predictions at the ecosystem level. He integrates different branches of ecology, and by elucidating the importance of biodiversity in ecosystem functioning, Michel contributes significantly to the advancement of biological conservation, and to applied ecology in general. His numerous publications in top journals have shaped the thinking of the current generation of ecologists. In EE Book 17 Michel aims at communicating his vision of biodiversity science to the broadest possible audience. He draws strongly from his experience as a leader of a variety of international organisations and activities devoted to research on and protection of biological diversity. In particular, he conveys his insights gained as Chair of DIVERSITAS, an international research programme founded in 1991 and sponsored by the International Council for Science (ICSU), the United Nations Educational, Scientific and Cultural Organization (UNESCO), and other international organisations. Michel argues that Earth is facing the 6th mass extinction in its history, this time not because of any natural disaster but as a repercussion of human activities, and he analyzes both the mechanisms and the consequences of massive biodiversity loss. Michel concludes ‘with a vibrant call for an indispensable re-examination of the philosophical and economic premises on which modern society is built’, expounding ‘that a reconciliation of humans with their own nature is key to resolving this crisis’ (p. 3).

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Ecology Institute Prize 2002 in the field of terrestrial ecology. Reproduction of the prize awarding document

ABOUT THE INTERNATIONAL ECOLOGY INSTITUTE

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About the International Ecology Institute The international Ecology Institute (ECI) was founded in 1984. It is a non-profit-making organization of research ecologists, sponsored by Inter-Research Science Publisher. The ECI’s aims and activities have been described in detail in my introduction to EE Book 3 (Gene E. Likens, The Ecosystem Approach: Its Use and Abuse, 1992). The ECI strives to achieve its aims by setting out awards to honor outstanding scientists: the ECI Prize (with associated EE Books) and the IRPE Prize. The Institute also supports postgraduates in eastern European countries via the Otto Kinne Foundation (OKF). ECI and IRPE Prizes. The ECI Prize honors the sustained high performance of outstanding research ecologists. It is awarded annually, in a rotating pattern, for the fields of marine, terrestrial and limnetic ecology. We realize that the division into such general fields is not very satisfactory; however, so far it has worked quite well. Laureates are elected by a jury of seven ECI members appointed by the ECI Director. The IRPE (International Recognition of Professional Excellence) Prize honors a young (not more than 40 years of age) research ecologist who has published uniquely independent, original and/or challenging papers representing an important scientific breakthrough and/or who must work under particularly difficult conditions. The prize recipients are elected by the ECI Jury mentioned above. OKF. The Otto Kinne Foundation supports promising young environmental scientists in eastern European countries. It aids postgraduates — without distinction of race, religion, nationality, or sex — by providing financial assistance for research projects, educational travel, and purchase of scientific equipment or published information. Details are available from the President of the Foundation: Dr. Anna F. Pasternak, P. P. Shirshov Institute of Oceanology, Russian Academy of Sciences, Nakhimovskii prospekt 36, Moscow 117 851, Russia (Email: [email protected]). Nominations. Nominations for ECI and IRPE Prizes (accompanied by the nominee’s CV, list of publications, and a statement why, in the opinion of the nominator, the nominee qualifies for the prize) are invited from research ecologists worldwide. They should be sent to the chairperson of the respective ECI Jury (see www.int-res.com/ecology-institute/call-fornominations/) or, alternatively, to the ECI’s director, who will then forward them to the chairperson. Eligible are all ecologists engaged in scientific research (except the ECI’s director, the Jury’s chairperson, and previous Laureates; Jury members nominated will be replaced by other ECI members). The Jury selects prize winners using the nominations received as well as their own knowledge of top performers and their own professional judgement. Nominations for OKF Fellows, to be addressed to Dr. Anna F. Pasternak (address given above) and accompanied by a letter of support as well as a brief documentation of the nominee’s performance, are invited from scientists worldwide.

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INTRODUCTION ECI Prize Winners, Their Major Scientific Achievements and Their Books*

Tom Fenchel (Helsingør, Denmark), ECI Prize winner 1986 in marine ecology. Quotation of the Jury (Chairman: John Gray, Oslo, Norway): The Jury found Professor T. Fenchel’s contribution to ecological knowledge in a variety of research fields to be of the highest international class. In particular, the Jury cites his brilliant and uniquely important studies on the microbial loop which have opened up a fundamentally new research field. Professor Fenchel is, in addition, an excellent publicizer in his field of research with authorship of a number of standard works in marine ecology. Book 1: Ecology – Potentials and Limitations. (Published 1987) Edward O. Wilson (Cambridge, MA, USA), ECI Prize winner 1987 in terrestrial ecology. Quotation of the Jury (Chairman: Sir Richard Southwood, Oxford, UK): Professor E. O. Wilson is distinguished for his many contributions to different aspects of ecology and evolutionary biology. His life-time love of nature, a theme explored in his book ‘Biophilia’, has been particularized in his study of ants leading to major new insights on the evolution of castes and the operation of social systems. His seminal ‘Sociobiology’, derived from this work, has founded a new branch of science, between ecology and the social sciences. With the late Robert MacArthur he was the originator of the modern theories of island biogeography that have contributed not only to the understanding of island biota, but to community and population ecology. Book 2: Success and Dominance in Ecosystems: The Case of the Social Insects. (Published 1990) Gene E. Likens (Millbrook, NY, USA), ECI Prize winner 1988 in limnetic ecology. Quotation of the Jury (Chairman: William D. Williams, Adelaide, Australia): Gene Likens is a distinguished limnologist who has made salient contributions to many fields of limnology. In 1962 he initiated and developed (with F. H. Bormann) the Hubbard Brook Ecosystem Study in New Hampshire. Comprehensive investigations in this study provided a model for ecological and biogeochemical studies worldwide. A major finding of the study was that rain and snow are highly acidic. ‘Acid rain’ is now recognized as one of the major environmental hazards in North America, Europe and elsewhere. Elected to the American Academy of Sciences in 1979, and the National Academy of Sciences in 1981, Gene Likens is a highly worthy recipient of the 1988 ECI Prize in Limnetic Ecology. Book 3: The Ecosystem Approach: Its Use and Abuse. (Published 1992) Robert T. Paine (Seattle, WA, USA), ECI Prize winner 1989 in marine ecology. Quotation of the Jury (Chairman: Tom Fenchel, Helsingør, Denmark): Robert T. Paine has made substantial and original contributions to marine biology and to ecology in general. In particular the Jury mentions the discovery of the role of patch formation and properties of food web structure in shaping communities of sedentary organisms. These studies (of which several have become classics of marine ecology) have fundamentally changed the way in which we view marine benthic communities. This work has also served as an inspiration for innovation in the mathematical description of community processes and has had a lasting impact on our understanding of ‘landscape dynamics’, of equal importance *For details on how to order EE Books consult the information at the end of this book

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to the development of the science of ecology and to conservation ecology. Book 4: Marine Rocky Shores and Community Ecology: An Experimentalist’s Perspective. (Published 1994) Harold A. Mooney (Stanford, CA, USA), ECI Prize winner 1990 in terrestrial ecology. Quotation of the Jury (Chairman: John L. Harper, Penmaenmawr, UK): Professor Harold A. Mooney is distinguished for his studies of the physiological ecology of plants, especially of arctic-alpine and mediterranean species. He has explored the ways in which plants allocate carbon resources and expressed this allocation in terms of costs, benefits and trade-offs. This has given a quantitative dimension to the study of plant-animal interactions and acted to integrate physiological ecology with population biology, community ecology, and ecosystem studies. Book 5: The Globalization of Ecological Thought. (Published 1998) Robert H. Peters (Montreal, PQ, Canada), ECI Prize winner 1991 in limnetic ecology. Quotation of the Jury (Chairman: Jürgen Overbeck, Plön, Germany): Professor R. H. Peters’ contributions to the fields of limnology and ecology have been numerous and far reaching. His work on phosphorus cycling in lakes provides examples of excellent research illuminating a number of important aspects regarding the movement and availability of phosphorus in aquatic systems. His book ‘The Ecological Implications of Body Size’ gives a powerful overview of the utility of allometric relationships for the study of ecological problems and for building ecological theory. Book 6: Science and Limnology. (Published 1995) Authors: The late F. H. Rigler and R. H. Peters David H. Cushing (Lowestoft, UK), ECI Prize winner 1992 in marine ecology. Quotation of the Jury (Chairman: John Costlow, Beaufort, NC, USA): Dr. David H. Cushing has, for many years, made an enormous contribution to the field of marine ecology through his numerous publications and his original ideas. His work continues to be highly influential in fisheries and plankton ecology. Although first published over ten years ago, his pioneering studies on the dynamics of a plankton patch, the feeding of copepods, the ‘match-mismatch’ theory of recruitment and the climatic influences on plankton and fisheries remain of central importance. Book 7: Towards a Science of Recruitment in Fish Populations. (Published 1996) Paul R. Ehrlich (Stanford, CA, USA), ECI Prize winner 1993 in terrestrial ecology. Quotation of the Jury (Chairman: Harold A. Mooney, Stanford, CA, USA): Dr. Paul Ehrlich’s scientific contributions have been substantial and sustained. The quality and depth of his interpretation of environmental issues to students, the general public, and to policy makers is unrivaled. His concern for both environmental quality and environmental justice has rarely been matched. He has made fundamental contributions to the study of population biology utilizing butterflies as a model system. These studies have had a large impact on how we view the population structure of organisms and have provided important guidelines on the conservation of wild populations. Book 8: A World of Wounds: Ecologists and the Human Dilemma. (Published 1997) Colin S. Reynolds (Ambleside, UK), ECI Prize winner 1994 in limnetic ecology. Quotation of the Jury (Chairman: William D. Williams, Adelaide, Australia): The research of Dr. Colin S. Reynolds in algology has several components interfacing with other biological disciplines, and, indeed, other sciences. He has pursued his research in depth and great detail and yet been able to relate and apply findings to holistic analysis of ecosystem function. Thus, his contributions to our understanding of the dynamic controls and

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responses of planktonic algae have provided new insight into several controversial areas of ecology. Dr. Reynolds’ ideas have also been widely applied in the water industry to reduce the impact of algal growth in reservoirs. His conversion of these ideas into mathematical models has resulted in a vastly increased application of his knowledge to the benefit of society. Book 9: Vegetation Processes in the Pelagic: A Model for Ecosystem Theory. (Published 1997) Ramon Margalef (Barcelona, Spain), ECI Prize winner 1995 in marine ecology. Quotation of the Jury (Chairman: Ernest Naylor, Menai Bridge, UK): Ramon Margalef is generally acknowledged to be the most prominent marine ecologist that Spain has produced. He has excelled in the study of unicellular algae, developing the paradigm of phytoplankton organization when previously such organisms were considered to be in unstructured suspension. He also pioneered the use of multidimensional statistical analyses in wider studies of marine plankton. The contributions which he has made to theoretical ecology have brought him particularly high international prestige, and have made him one of the most frequently quoted contemporary ecologists. As one reviewer of his work has noted, ‘Margalef ’s ideas have provoked thought, an enviable encomium for any scientist’. Book 10: Our Biosphere. (Published 1997) John H. Lawton (Ascot, UK), ECI Prize winner 1996 in terrestrial ecology. Quotation of the Jury (Chairman: Ilkka Hanski, Helsinki, Finland): The ECI Jury selected John Lawton for his distinguished conceptual, theoretical and empirical contributions to population, community and ecosystem ecology since the early 1970s. Lawton is perhaps best known for his elegant and comprehensive work on folivorous insects, including a long-term research project on the food web structure of bracken insects. He is also one of the leaders in macroecology, the study of large-scale patterns in animal and plant communities. Book 11: Community Ecology in a Changing World. (Published 2000) Z. Maciej Gliwicz (Warsaw, Poland), ECI Prize winner 1997 in limnetic ecology. Quotation of the Jury (Chairman: Winfried Lampert, Plön, Germany): Since the early 1970s, Maciej Gliwicz has exerted an outstanding impact on freshwater plankton ecology. He has laid the foundations for our current understanding of the mechanisms of phytoplankton-zooplankton interactions. For example, he pioneered the work on selective feeding by zooplankton and mechanical interference by inedible algae. His studies on the relative impact of bottom-up and top-down forces in plankton communities and on the evolution and constraints of defense mechanisms have greatly influenced modern lake ecosystem and trophic cascade theory. Maciej Gliwicz has always been in favor of unusual ideas, and, with his legendary enthusiasm, he has become a creative and thoughtful leader in evolutionary aquatic ecology. Book 12: Between Hazards of Starvation and Risk of Predation: The Ecology of Offshore Animals. (Published 2003) Richard T. Barber (Beaufort, NC, USA), ECI Prize winner 1998 in marine ecology. Quotation of the Jury (Chairman: B.-O. Jansson, Stockholm, Sweden): Richard T. Barber studied the productivity of coastal-upwelling ecosystems off California, Ecuador, Peru, North Africa and the Arabian Sea. One of the strongest advocates of the integration of biological oceanography with physical and chemical oceanography both in field work and numerical modelling as early as the 1960s, he has had a large share in the shaping of modern oceanography to a unified discipline. His own intensive time series work on the eastern side of the Pacific led to the first detailed documentation of the biological effects of the El Niño episode 1982/83, long before El Niño phenomena attracted general attention. Barber also documented the 1991/92 episode. He played a leading role in the testing of the iron hypothesis,

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which is based on iron being one of the limiting elements for primary production in the open ocean. The international test expedition he initiated to the tropical eastern Pacific was an outstanding success, receiving considerable attention in leading scientific journals. Dr. Barber has been able to develop constructive working relationships with scientists and research institutes in disadvantaged countries—in particular Peru and Ecuador—and in China. Book 13: The Response of Oceanic Ecosystems to the Climate of the 21st Century. (tentative title; in preparation) I. Hanski (Helsinki, Finland), ECI Prize winner 1999 in terrestrial ecology. Quotation of the Jury (Chairman: F. A. Bazzaz, Cambridge, MA, USA): Ilkka Hanski’s research in population and community ecology spans 20 years. His work is characterized by a combination of theory, modelling and empirical research. Though he has worked on a range of taxa and questions, most of his research has focussed on the spatial structure and dynamics of populations. Spatial ecology has developed into an important area of ecology over the past 10 years, and Hanski’s research has made significant contributions to this development. In the 1970s and early 1980s, much of Hanski’s research was on small-scale spatial structure of populations, the empirical work being done on insects living in ephemeral microhabitats. Hanski demonstrated how intraspecifically aggregated spatial distributions of species may facilitate coexistence of competitors, and how a generalist predator using the spatially aggregated prey species may have the same effect. Experimental work on blowflies supported these predictions and may help to explain high species richness in insect communities. In the late 1980s, Hanski’s research shifted to larger spatial scales and to metapopulations. He has made major contributions to both theoretical and empirical aspects in the field, and edited, together with Michael Gilpin, the two most widely read volumes in metapopulation biology (Metapopulation Dynamics, 1991, and Metapopulation Biology, 1997). Book 14: Habitat Loss and its Biological Consequences. (tentative title; in preparation) Stephen R. Carpenter (Madison, WI, USA), ECI Prize winner 2000 in limnetic ecology. Quotation of the Jury (Chairman: Wolfgang Wieser, Innsbruck, Austria): Stephen R. Carpenter, Halverson Professor of Limnology and Professor of Zoology at the University of Wisconsin-Madison, USA, is a leader in ecosystem science. His demonstrations of the role of fish in controlling lake productivity and nutrient cycling have brought wholeecosystem experimentation to a new level of sophistication. Carpenter leads a multidisciplinary research team that has manipulated fish community structure and nutrient inputs of entire lake ecosystems. These large-scale experiments demonstrated trophic cascades through the food web, altering primary production, nutrient cycling, and gas exchange between lakes and the atmosphere. Carpenter has received many awards for distinguished research, including the R.E. Hutchinson Medal of the American Society of Limnology and Oceanography, the R.H. MacArthur Award from the Ecological Society of America, and the Per Brinck Award in Limnology. In 1999 he assumed leadership of the North Temperate Lakes LTER site, a collaboration of more than 20 principal investigators from 7 academic departments and 2 agencies, studying long-term lake dynamics, land-water interactions, and the interactions of people and lakes in rural and urban regions. Recently Carpenter was elected President of the Ecological Society of America. Book 15: Regime Shifts in Lake Ecosystems: Pattern and Variation. (Published 2003) Louis Legendre (Villefranche-sur-Mer, France), ECI Prize winner 2001 in marine ecology. Quotation of the Jury (Chairman: Richard T. Barber, Beaufort, NC, USA): Professor Louis Legendre has advanced our understanding of how ocean ecosystems func-

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tion. Elegantly integrating observation, experimentation and theory, Legendre’s work contributes both pragmatic and theoretical advances. He pioneered the concept of hydrodynamic control of biogenic carbon fluxes in open ocean and coastal regimes, an advance that has importance for the future course of carbon partitioning in a world significantly altered by anthropogenic activities. His wide-ranging investigations relating physical processes to biological responses led him to develop the concept of ‘Dynamic Biological Oceanography’. Based in part on the seasonal physical progression that characterizes high temperate and polar oceans, this broad concept involves a mechanistic understanding of species succession, photoadaptation, nutrient limitation, temperature responses, grazing and sedimentation. That Legendre’s contributions are characterized by unusual quantitative rigor is evidenced by his participation in the creation and development of the new discipline of ‘numerical ecology’. In conclusion, we cite Prof. Legendre's long-term project of developing a unified theoretical framework for biological oceanography, an ambitious undertaking that is still in progress. Book 16: Scientific Research and Discovery: Process, Consequences and Practice. (Published 2004) Michel Loreau (Paris, France), ECI Prize winner 2002 in terrestrial ecology. Quotation of the Jury (Chairman: Paul Ehrlich, Stanford, CA, USA): The Jury found that Professor Michel Loreau has contributed in a number of significant ways to important increases in ecological knowledge. His ability to use models to explore the effects of phenomena at the level of communities to make predictions about the ecosystem level provides predictions for novel experimental approaches to fields of ecology with a strong descriptive basis. In this way, his theoretical and empirical insights have contributed significantly to the integration of different branches of ecology. The studies of Michel Loreau will have great impact on applied ecology and issues of conservation biology related to the ecosystem effects of climatic change. His many important papers in international journals of the highest rank are widely known and have strongly influenced a new generation of ecologists. Michel Loreau is a highly worthy recipient of the 2002 ECI Prize in Terrestrial Ecology. Book 17: The Challenges of Biodiversity Science. (Published 2010) Jonathan J. Cole (Millbrook, NY, USA), ECI Prize Winner 2003 in limnetic ecology. Quotation of the Jury (Chairman: Colin Reynolds, Ambleside, UK): The innovative studies of Jonathan Cole have achieved distinction in several areas of aquatic biology, especially in microbial ecology, nutrient biogeochemistry and carbon cycling. Challenging and controversial, his investigations of bacterial activity and its contribution to the gas balance in lake systems have succeeded in establishing fresh insights into the relationships between primary productivity, microorganisms and the metabolism of lakes, within the broader context of their hydrological landscapes. Jonathan Cole has a proven ability to synthesise his own work and that of others, melding a wider, integral understanding of the ways in which lakes function. He is among the true leaders of contemporary ecology. Book 18: Multiple Roles of Freshwater Ecosystems in the Carbon Cycle. (tentative title; in preparation) Bo Barker Jørgensen (Bremen, Germany), ECI Prize winner 2004 in marine ecology Quotation of the Jury (Chairmann: Victor Smetacek, Bremerhaven, Germany): Bo Barker Jørgensen's qualitative and quantitative description of the microbial sulphur cycle in marine sediments revolutionised microbial ecology and biogeochemistry. Ever since his early career he has continued to play a leading role in aquatic microbiology. In 1992 Jør-

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gensen was invited to found and co-direct the Max Planck Institute for Marine Microbiology in Bremen. His combination of skills, especially his ability to stimulate and inspire scientists around him, established the institute at the cutting edge of international research from the start. Together with his co-workers, he has demonstrated the limiting role of molecular diffusion for process rates, showed how different element cycles interact and discovered novel physiological types of microorganisms. Bo is amongst the 'highly cited scientists' of ISI. In 2004 he was presented the Hutchinson Award by the American Society of Limnology and Oceanography. In addition to his outstanding scientific attributes he is a very likeable man, known to his students as the 'silent giant'. Book 19: Anoxic ecosystems and biogeochemical cycles. (in preparation) Robert D. Holt (Gainsville, FL, USA), ECI Prize winner 2005 in terrestrial ecology Quotation of the Jury (Chairman: Michel Loreau, Paris, France): Robert D. Holt has made significant theoretical contributions to population, community, and evolutionary ecology during the last three decades. His early work on the role of predation on the structure of ecological communities has had a long-lasting influence on the way ecologists view communities, by showing that predation strongly affects the outcome of competitive interactions and is even the vehicle of new competitive interactions among species. Robert Holt’s work on the spatial dynamics of populations and communities has been equally influential; it has contributed to the emergence of the current interest in metacommunities and the relationship between local and regional ecological processes. The evolutionary dimension of ecological processes is one of his credos; he has had a special interest in the evolutionary response of species to a spatially heterogeneous environment. Although mainly theoretical, his work also has substantial implications for conservation biology. The Jury particularly valued the breadth of his views, the innovative nature of his approaches, and his efforts to link different areas of ecology and evolutionary biology. It awards the 2005 ECI Prize in Terrestrial Ecology to Robert Holt for these outstanding qualities. Book 20: On the conceptual unification of ecology: an unfinished agenda. (tentative title; in preparation) Winfried Lampert (Plön, Germany), ECI Prize winner 2006 in limnetic ecology Quotation of the Jury (Chairman: Nelson G. Hairston, Jr., Ithaca, NY, USA): Winfried Lampert is one of the most influential international limnologists of the past four decades. His strong advocacy for including adaptive evolutionary considerations in studies of the functioning of plankton populations and their role in limnetic ecosystem dynamics has been fundamental in shaping research on lakes. His choice of Daphnia as a model organism for his own investigations, and his success in demonstrating by example the progress in understanding gained through rigorous laboratory experiments using this ecologically critical species, has established a research program that now encompasses a large network of scientists around the globe. Winfried Lampert’s studies, and those of his students, of Daphnia feeding behavior, growth rate, population dynamics, reaction to predators, and impacts on primary production, have identified the central role that Daphnia plays in the dynamics of the pelagic zone. His generosity as a host for visitors to the Max-Planck-Institute for Limnology is legendary, and has established Plön as a world center for the study of evolutionary ecology where there is always a stimulating mixture of ideas and research in progress. Book 21: Daphnia — The development of a model organism in aquatic ecology. (tentative title; in preparation)

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Daniel Pauly (Vancouver, Canada), ECI Prize winner 2007 in marine ecology Quotation of the Jury (Chairman: Bo Barker Jørgensen, Bremen, Germany): Daniel Pauly is Professor of Fisheries and Director of the Fisheries Centre at the University of British Columbia. He is a distinguished scientist in marine fisheries biology and the management of fisheries resources and marine ecosystems. Among his important contributions, he pioneered the assessment of the global impact of fishing on marine ecosystems. His life-work and ideas have widened the scope of fisheries ecology and contributed to the harmonious coexistence of humans with nature. His work on marine ecosystem conservation has met widest acceptance through his 'Marine Trophic Index, MTI' (i.e. the mean trophic level of fisheries landings). MTI was endorsed in 2004 by the Convention of Biological Diversity as one of eight ecological/biodiversity indicators for immediate testing. He is one of the most highly-cited marine ecologists and was recently ranked among the world's 50 leading scientists (Scientific American's 50 Research Leaders in 2003). Book 22: Gasping Fish and Panting Squids: Oxygen, Temperature and the Growth of Water Breathing Animals. (Published 2010) Monica Turner (Madison, WI, USA), ECI Prize winner 2008 in Terrestrial Ecology Quotation of the Jury (Chairman: Robert D. Holt, Gainsville, FL, USA): Dr. Monica Turner has been a driving force in the development of the field of landscape ecology. She has carried out exemplary field studies and developed methods, models, and theories that permit this field to go beyond mere description to make testable, quantitative predictions. She has provided important and timely syntheses of landscape ecology, grounded in her significant empirical studies in many systems, most notably in the Greater Yellowstone Ecosystem, where she has developed key insights into long-term vegetation dynamics across many scales that incorporate the interplay of changing disturbance regimes, vertebrate grazing, and soilmicrobe-nutrient interactions. This integrated body of work sets a standard for research in landscape ecology, and has led to important advances in addressing applied problems. Her many publications are highly cited, and together with her excellence as a researcher, her warm personal qualities have made her an inspiration for many young scientists. Book 23: Lessons from Landscape Ecology. (tentative title; in preparation) Brian Moss (Liverpool, UK), ECI Prize winner 2009 in Limnetic Ecology Quotation of the Jury (Chairman: Morten Søndergaard, Copenhagen, Denmark): Professor Brian Moss has invested his scientific and scholarly skills and enthusiasm for research within a variety of limnological fields for more than four decades, and his studies have covered freshwater ecosystems from the Arctic to the tropics. The Jury especially emphasize the many excellent studies Brian Moss has carried out on eutrophication of shallow lakes, particularly with respect to the problems of the Norfolk Broadland. One of the many virtues reflected in his research is the ability to work with entire ecosystems, and to understand how different groups of organisms interact and how freshwater ecosystems behave under pressure from human activities, whether in the form of excess nutrient loads or changed climate. Brian Moss was one of the first scientists to carry out full-scale biomanipulation studies, thereby closing the gap between scientific understanding of trophic cascading and treatment of ecological problems. The excellence in research is mirrored by his abilities to describe cutting-edge science in textbooks and to present science to the general public; often even with a touch of poetry and art. Professor Brian Moss is a most worthy recipient of the ECI Prize in Limnetic Ecology 2009. Book 24: Liberation Ecology: The Reconciliation of Natural and Human Cultures. (tentative title; in preparation)

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IRPE Prize Winners and Their Major Scientific Achievements Not in all years did the Jury elect an IRPE Prize winner. The nominations received were either too few or not sufficiently strong Colleen Cavanaugh (The Biological Laboratories, Harvard University, Cambridge, MA 02138, USA), IRPE Prize winner 1986 in marine ecology. Quotation of the Jury (Chairman: John Gray, Oslo, Norway): The Jury found the research of Dr. C. Cavanaugh on chemosynthesis—initially concerning hot-vent fauna but extended to other sulphide-rich habitats—to be highly original and to represent a major scientific breakthrough. Her hypothesis, formulated whilst a beginning graduate student, met severe opposition from established scientists with opposing views, but nevertheless proved to be correct. The Jury acknowledge Dr. Cavanaugh’s brilliant and independent research in understanding chemosynthetic energetic pathways. ˇ ˇ Karel Simek (Hydrobiological Institute, Czech Academy of Sciences, 370 05 Ceské Budˇejovice, Czech Republic), IRPE Prize winner 1991 in limnetic ecology. Quotation of the Jury (Chairman: Jürgen Overbeck, Plön, Germany): ˇ Dr. Karel Simek belongs to the generation of young limnologists in Eastern Europe who— despite lack of international information exchange—published, under difficult conditions, excellent contributions to the field of Aquatic Microbiology. He enjoys a high international ˇ reputation. Under the present, improved conditions Simek is likely to proceed even more successfully to new professional horizons. Richard K. Grosberg (Department of Zoology, University of California, Davis, CA 95616, USA), IRPE Prize winner 1992 in marine ecology. Quotation of the Jury (Chairman: John Costlow, Beaufort, NC, USA): Richard K. Grosberg has not only published extensively on fundamental issues relating to marine ecology, but has also demonstrated his understanding of marine ecology through superb teaching of invertebrate zoology to undergraduate and graduate students. He is acknowledged as a leader in adapting molecular techniques for the study of marine larvae and in developing information on extraordinarily detailed mapping studies of the genetic structure of adult populations of marine organisms. Nikolai V. Aladin (Zoological Institute, Russian Academy of Sciences, St. Petersburg 199034, Russia), IRPE Prize winner 1993 in terrestrial ecology. Quotation of the Jury (Chairman: Harold A. Mooney, Stanford, CA, USA): Dr. Nikolai V. Aladin is one of Russia’s most eminent young ecologists. He has researched environments in the former Soviet Union, particularly in Kazakhstan where he and a small team have focussed upon the area of the Aral Sea. Dr. Aladin’s studies were performed during a period of change, both in the patterns of organismic assemblages and in the political structure of his country. These studies were undertaken in his own time and at his own expense. It is only over the past few years that his studies have been officially supported and their value recognized. Stephen J. Hawkins (Centre of Environmental Sciences, University of Southampton, UK), IRPE Prize winner 1995 in marine ecology. Quotation of the Jury (Chairman: Ernest Naylor, Menai Bridge, UK): At the start of his research career Stephen Hawkins resisted pressures to work on topics for which funding was known to be available, preferring to develop his own ideas and to be judged on those. It was a brave stance when trying to formulate a Ph.D. programme in a harsh financial climate, but he was successful and began imaginative field experiments on

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rocky shore communities to test his ideas on species/area concepts and on the temporal basis of mosaic distributions. He followed these with detailed studies of intertidal gastropods as models for ideas on niche theory and competition. The outcome of his work has been to generate important new insights into quantitative sampling techniques and environmental impact assessments of rocky shore communities. Susan Harrison (Division of Environmental Studies, University of California, Davis, CA 95616, USA), IRPE Prize winner 1996 in terrestrial ecology. Quotation of the Jury (Chairman: Ilkka Hanski, Helsinki, Finland): Susan Harrison has made distinguished contributions to population ecology and conservation biology. Her Ph.D. research on the Edith's checkerspot butterfly has had a major impact on the way ecologists think about spatially structured populations. Her more recent empirical research has tested intriguing theoretical predictions about the spatial dynamics of animal and plant populations. Susan Harrison has been exceptionally influential in clarifying and interpreting the implications of population biological research to conservation. Jef Huisman (Department of Biological Sciences, Gilbert Hall, Stanford University, Stanford, CA 94305, USA), IRPE Prize winner 1997 in limnetic ecology. Quotation of the Jury (Chairman: Winfried Lampert, Plön, Germany): Jef Huisman has made significant contributions to resource competition theory. Through a unique blend of theory and elegant experiments, he has included light as a resource into the framework of Tilman's mechanistic competition models. Light is a unique resource as it cannot be intermixed like nutrients. Models of light-limited growth and competition, therefore, require a vertical light gradient and a ‘critical light intensity’ for different species. Jef Huisman has developed the existing theory further and has discovered that not only the ratio of nutrients and light determines the outcome of competition, but also the absolute supply. This outstanding work will have a strong influence on general ecological theory. Philip Boyd (NIWA Centre for Chemical and Physical Oceanography, Department of Chemistry, University of Otago, Dunedin 9001, New Zealand), IRPE Prize winner 1998 in marine ecology. Quotation of the Jury (Chairman: B.-O. Jansson, Stockholm, Sweden): Dr. Philip Boyd is an internationally acknowledged authority in the fields of phytoplankton community dynamics and oceanic productivity. His thesis ‘The flow of carbon in marine microbial ecosystems’ provided him with a broad base for launching his scientific career. From postdoctoral positions in the UK and Canada and from his present position in New Zealand he has participated in and contributed significantly to the international Joint Global Ocean Flux Study. His work on the relationship between sinking flux and pelagic community composition is truly interdisiplinary as it links pelagic ecology with marine geochemisty. Throughout his career he has been involved in large-scale field measurements of the open ocean that have applied methods and techniques ranging from molecular biology to ocean physics. He has collaborated with leading scientists and with great success amalgamated complex data sets and modelled major processes from photosynthesis to sinking flux. Currently Boyd has taken up the challenge of organising and leading the first iron fertilisation experiment to test the iron hypothesis in the Southern Ocean. Since this is the first joint South African - New Zealand scientific undertaking its significance for ocean ecology is complemented by its political dimension. Indeed a heavy burden to be carried by a young scientist! Kevin J. Gaston (Department of Animal and Plant Sciences, University of Sheffield, Sheffield S10 2NT, UK), IRPE Prize winner 1999 in terrestrial ecology. Quotation of the Jury (Chairman: F. A. Bazzaz, Cambridge, MA, USA):

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Dr. Kevin J. Gaston has played a major role in the development of macroecology and biodiversity. He established the existence of several major macroecological patterns in insect assemblages. His thesis (which appeared as a series of papers in Nature, Journal of Animal Ecology, American Naturalist and Oikos) clearly demonstrated the need for a macroecological viewpoint. Since then Dr. Gaston has sought to determine the generality of a wide diversity of patterns in macroecology, the fundamental structure of these patterns, the mechanisms which generate them, and their wider implications. Features of these studies have been the development of appropriate statistical tools, the use of null models, and a drive to distinguish real ecological patterns from artefacts and to establish the fundamental mechanisms which cause the real patterns. A number of macroecological hypotheses have been tested for the first time, and others have been subject to more detailed scrutiny than had previously been the case. He has challenged our thinking about the form of macroecological patterns (e.g. measurement of population variability) and their determinants (e.g. abundance-occupancy relationships). This has resulted in stimulating debates. Ruben Sommaruga (Institute of Zoology and Limnology, University of Innsbruck, Technikerstr. 25, 6020 Innsbruck, Austria), IRPE Prize winner 2000 in limnetic ecology. Quotation of the Jury (Chairman: Wolfgang Wieser, Innsbruck, Austria): Ruben Sommaruga, a native of Uruguay, received his Ph.D. at the University of Innsbruck, Austria, where he is now an Associate Professor. He is known for his outstanding scientific contributions to two fields, microbial ecology of hypertrophic lakes, and UV-photobiology. He has proposed a general theory about the roles of Microcystis aeruginosa and Planktothrix agardhii in shaping the structure of the microbial food web and the ecological importance of filamentous grazing-resistant bacteria in hypertrophic lakes. Among his most influential achievements in the second field are the findings that ultraviolet-A (UVA) radiation is as effective in reducing bacterial activity as UVB, that UV radiation has the potential of affecting carbon flow between protists and bacteria, and that intracellular UV-absorbing compounds, called mycosporine-like amino acids, are present not only in marine organisms but also in inhabitants of alpine lakes. Sommaruga has published extensively on plankton ecology, water chemistry, and UV-physics. He is still conducting and supervising projects in Uruguay, particularly with respect to the effects of UV radiation in coastal lagoons and practical problems of eutrophication. David M. Post (Department of Ecology and Evolutionary Biology, Yale University, New Haven, CT 06520 USA), IRPE Prize Winner 2003 in limnetic ecology. Quotation of the Jury (Chairman: Colin Reynolds, Ambleside, UK): David Post has quickly established a reputation for interesting and innovative research in the variability inherent in food-web structure, especially that which is attributable to the ontogenies and variable demographies of the main components. Well-versed in trophic cascades, the dynamics of fish populations and their roles in nutrient cycling, David has already registered an outstanding contribution to the development of limnetic ecology. His recognition that food-chain length is related primarily to the size of the supportive ecosystem, rather than to the efficiency of energy transfer, as hitherto supposed, has confirmed him as one of the brightest young limnetic ecologists. Members of the Jury consider him especially worthy of the IRPE Laureate and the encouragement for further studies that it brings. Markus G. Weinbauer (Laboratoire d'Océanographie de Villefranche-sur-Mer, France), IRPE Prize winner 2004 in marine ecology. Quotation of the Jury (Chairman: Victor Smetacek, Bremerhaven, Germany):

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Dr. Markus G. Weinbauer has made major contributions towards elucidating the role of viruses in microbial food webs and their effect on prokaryotic diversity. His work has provided new insights into the role of viruses as pathogens of prokaryotes in oxic and anoxic systems, the relationship between viral lysis and dormant viral infection in surface, deep and anoxic waters, the role of repair of ultraviolet radiation-induced DNA damage in the sustenance of high viral activity, and the relative effect of viral lysis and protist grazing of prokaryotes in food web processes in marine and freshwater systems. Markus has worked in Europe and the USA and is actively engaged in training students and young scientists. Andrew Hector (Institute of Environmental Sciences, University of Zürich, Switzerland), IRPE Prize winner 2005 in terrestrial ecology. Quotation of the Jury (Chairman: Michel Loreau, Paris, France): Andrew Hector is one of the most talented experimental ecologists of his generation. His publication record in top-quality journals is outstanding for his age. He is recognised throughout the world as an expert in the new biodiversity and ecosystem functioning area that has emerged during the last decade. He has played an important role in the design, analysis and propagation of large-scale experiments that test the effects of biodiversity changes on ecosystem processes in terrestrial systems, and he continues to do so through his responsibilities in DIVERSITAS, the international programme of biodiversity science. Thanks to his scientific understanding, open-mindedness, and honesty, he was also one of the persons who contributed to resolve the controversy over the interpretation of biodiversity experiments, and thereby to provide the bases for the continued development of this scientific area. These achievements make him a worthy recipient of the 2005 IRPE Prize in Terrestrial Ecology. M. Jake Vander Zanden (Center for Limnology, University of Wisconsin, Madison, Wisconsin, USA), IRPE Prize winner 2006 in limnetic ecology. Quotation of the Jury (Chairman: N. G. Hairston Jr., Ithaca, NY, USA): Jake Vander Zanden, seven years after receiving his doctorate, is already a major contributor to substantive ideas about the structure and functioning of lake ecosystems. Using stable isotope analyses, his research has provided significant new insights to the structure of limnetic food webs, the substantial role that benthic pathways play in energy transfer within lakes, and the impacts of introduced species on these dynamics. Jake Vander Zanden has been remarkably successful in joining basic and applied ecology, using his studies in each sphere to inform his understanding in the other, and he is an active advisor to various governmental and private management organizations with missions to conserve and restore aquatic ecosystems. Marcel M. M. Kuypers (Max Planck Institute for Marine Microbiology, Bremen, Germany), IRPE Prize winner 2007 in marine ecology. Quotation of the Jury (Chairman: Bo B. Jørgensen, Bremen, Germany): Marcel M. M. Kuypers is a marine organic chemist who during his short career has made major breakthroughs in environmental science — first on the analysis of large-scale ocean anoxia in the geological past and second on the modern oxygen minimum zones and their function as a major sink for nitrogen in the ocean. Marcel Kuypers applies diagnostic membrane lipids of prokaryotic organisms as molecular signals for microbial communities and their environment. He combines such information with process studies using isotope tracers and with DNA/RNA-based probes and sequence analyses in a unique manner. His recent work on anammox has changed the way we understand the nitrogen balance in the ocean and thereby also the regulation of ocean productivity. Marcel Kuypers is now the head of a dynamic, independent research group of the Max Planck Society and continues to do internationally outstanding research in aquatic biogeochemistry and microbial ecology.

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Campbell O. Webb (Harvard University Herbaria, Cambridge, MA, USA), IRPE Prize winner 2008 in terrestrial ecology. Quotation of the Jury (Chairman: Robert D. Holt, Gainesville, FL, USA): Dr. Campbell Webb has been instrumental in developing a fresh approach to community organization — community phylogenetics — which uses the tools of phylogenetic systematics to add a historical dimension to issues of community structure and assembly. His innovative empirical studies in plant communities have helped lead ecologists to reconsider the role of niche of differentiation in both mature plant communities, and in communities experiencing invasion. He has contributed to the emergence of community phylogenetics both through both conceptual syntheses and the development of software tools that help community ecologists apply phylogenetic perspectives to their datasets. At the same time, he is deeply devoted to understanding and conserving tropical forests, and he has conducted much of his research while based in Borneo, contributing there to human capital development in local communities and to conservation. For all these reasons, the jury awards Cam Webb the 2008 IRPE Prize in Terrestrial Ecology.

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Alexei Essenin (Yesenin), 1994 Institute of Evolutionary Morphology and Ecology of Animals of Russian Academy of Sciences, Laboratory of Bioindication, 33 Leninsky Prospekt, Moscow 117071, Russia Supported project: Metal accumulation patterns of terrestrial invertebrates Vojtech Novotny, 1994 Institute of Entomology, Academy of Sciences of the Czech Republic, Branisˇ ovská 31, 370 05 Cˇ eské Budeˇ jovice, Czech Republic Supported project: Ecological research on insect groups in a Papua New Guinea rainforest Inna M. Sokolova, 1995 White Sea Biological Station, Zoological Institute of the Academy of Science of Russia, Universitetskaya nab., 1, St. Petersburg 199034, Russia Supported project: Influence of salinity on marine bivalve populations Andrej V. Grischenko, 1995 Faculty of Biology and Soil Sciences, St. Petersburg State University, 16 Linia, 29, St. Petersburg 199178, Russia Supported project: Ecology of intertidal bryozoans of the Commodore Islands shelf region Ferenc Baska, 1995 Veterinary Medical Research Institute, Hungarian Academy of Sciences, POB 18, 1581 Budapest, Hungary Supported project: Identification of fish coccidia and myxosporea using MAb and PCR techniques Andrei A. Filippov, 1997 Laboratory of Brackishwater Biology, Zoological Institute of the Russian Academy of Sciences, Universitetskaya nab., 1, 1999034 St. Petersburg, Russia Supported project: Benthic fauna of the Aral Lake under impact of changed environment Konstantin A. Lutaenko, 1997 Institute of Marine Biology, Far East Branch of the Russian Academy of Sciences, Palchevsky Street 17, Vladivostok 690041, Russia Supported project: Paleoclimatology as reflected by molluscan fauna of Eurasian coastal regions Sergey V. Dobretsov, 1997 Laboratory of Invertebrate Zoology, Biological Research Institute, St. Petersburg University, Oranienbaumskoye sch. 2, Stary Peterhof 198904, Russia Supported project: Successional community dynamics of hydrozoan and mollusc larvae Marina I. Orlova, 1997 Laboratory of Brackishwater Biology, Zoological Institute of the Russian Academy of Sciences, Universitetskaya nab., 1, 1999034 St. Petersburg, Russia Supported project: Salinity responses and their molecular background in euryhaline and stenohaline species of dreissenid bivalves Marie Alexandrovna Saburova, 1998 Institute of Biology of the Southern Seas, National Academy of the Ukraine, Prospekt Nakhimova 2, Sevastopol 335001, Crimea, Ukraine Supported project: Studies on the distribution, migration, cell cycle and nutrition of microalgae in marine intertidal sediments and environmental factors which may affect them Jan Zukal, 1998 Institute of Landscape Ecology, Academy of Sciences of the Czech Republic, Kvetná 8, 60365 Brno, Czech Republic Supported project: Research on the ecology of bats, their foraging activity, changes in abundance and problems of bat conservation

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András Báldi, 1998 Animal Ecology Research Group of the Hungarian Academy of Sciences and the Hungarian Natural History Museum, Baross u. 13, H-1088 Budapest, Hungary Supported project: Testing of indicators of biodiversity in Hungarian nature reserves Alexey Valerievich Smirnov, 1999 St. Petersburg State University, Faculty of Biology, Department of Invertebrate Zoology, St. Petersburg, Russia Supported project: Elucidation of distribution and dynamics of abundance of gymnamoebae in bottom sediments of a freshwater lake, using a newly developed original method Andrei Nikolaevich Ostrovsky, 1999 St. Petersburg State University, Faculty of Biology, Department of Invertebrate Zoology, St. Petersburg, Russia Supported project: Reproductive patterns and life cycles of White Sea bryozoans, with special emphasis on the structure and dynamics of populations of different species Alexey Vladimirovich Rybakov, 1999 Chorology Department, Institute of Marine Biology, Far Eastern Branch of the Russian Academy of Sciences, Vladivostok, Russia Supported project: Larval characters in the taxonomy of parasitic cirriped barnacles (Rhizocephalid crustaceans) Vojtech Novotny, 2001 Faculty of Biological Sciences, University of South Bohemia, >eské Budeˇjovice, Czech Republic Supported project: Beta diversity of herbivorous insects in lowland rainforests of Papua New Guinea. Development of conservation strategies for the lowland tropics Yuri Kvach, 2001 Department of Zoology, Odessa University, 2 Shampansky Provulok, 65058 Odessa, Ukraine Supported project: Biodiversity of helminths of gobiid fishes from the North-Western Black Sea region Csaba Székely, 2001 Veterinary Medical Research Institute of the Hungarian Academy of Sciences, 21, Hungaria krt, Budapest, H-1142, Hungary Supported project: Parasitoses in fresh-water and pond-farm fisheries, including control measures Alexander V. Kirdyanov, 2003 Institute of Forest, Siberian Branch of Russian Academy of Sciences (SB RAS), Akademgorodok, Krasnoyarsk 660036, Russia Supported project: Tree-ring growth — climate interactions along the longitudinal transect in Siberia in the context of global and regional climatic changes Piotr Kuklinski, 2003 Institute of Oceanology, Polish Academy of Science, ul. Powstancow Warszawy 55, Sopot 81–712, Poland Supported project: Ecology of bryozoans from Svalbard waters Kasia Salwicka-Ciaputa, 2003 Department of Antarctic Biology, Polish Academy of Sciences, Ustrzycka 10, 12, 02141 Warsaw, Poland Supported project: Genetic study of Southern elephant seal (Mirounga leonina) population in Admiralty Bay, King George Island, Antarctica Bohdan Prots, 2005 State Museum of Natural History, National Academy of Sciences of Ukraine, 18 Tetralna Str., Lviv 79008, Ukraine Supported project: The Transcarpathian Riverine Forests Nature Park — Saving the Future of Unique Floodplain Forests

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Ruslan Salamatin, 2005 Department of General Biology and Parasitology, Medical University of Warsaw, Warsaw, Poland and Department of Parasitology, I. I. Schmalhausen Institute of Zoology, National Academy of Sciences of Ukraine, Chmielnicki 15, Kyiv, Ukraine Supported project: Comparative studies on helminth fauna of wild ducks (subfamily Anatinae) in coastal regions of Poland and Ukraine in ecological aspect István Szentirmai, 2005 Department of Ethology, Eotvos Lorand University, Pazmany P. Setany 1/C, 1117 Budapest, Hungary Supported project: Habitat selection and reproductive success of the endangered Montagu’s harrier in agricultural landscapes Martina Petrú, 2007 Mnisek 95, Trebon 37901, Czech Republic Supported project: Regeneration capacity and resilience of Madagascar’s arid spiny forest in a rehabilitation perspective Csilla Stenger-Kovács, 2007 University of Pannonia, Department of Limnology, Egyetem str. 10, Veszprém 8200, Hungary Supported project: Development of equilibrium phases and the role of herbivory in periphyton composition of streams with special emphasis on diatom communities Tina N. Molodtsova, 2007 P.P. Shirshov Institute of Oceanology RAS, 36 Nakhimovskiy prospect, Moscow 117218, Russia Supported project: Deep-sea Antipatharia (Anthozoa: Cnidaria) of the North-East Atlantic: continental slope vs. seamount faunas Arevik Minasyan, 2009 st. Marrie 5, apt. 3, Yerevan 0079, Armenia Supported project: Dynamics of microbial self–cleaning power of Lake Sevan and its main tributaries Zoltan Tóth, 2009 105 Kassai str., H-8000 Székesfehérvar, Hungary Supported project: Investigation of the emergence of non-random following networks in house sparrow flocks Lenka Trebatická, 2009 Kauppatie 58 A 2, Konnevesi, 443 00 Finland Supported project: The effect of Puumala hantavirus on survival of its host, the bank vole (Myodes glareolus)

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*Ecology Institute Staff 2009 (in brackets: year of appointment)* Director and Founder: Professor O. Kinne, Nordbünte 23, 21385 Oldendorf/Luhe, Germany Marine Ecology Prof. F. Azam, La Jolla, CA, USA (1985) Prof. R. T. Barber, Beaufort, NC, USA (1999) Prof. J. Cebrian, Dauphin Island, AL, USA (2009) Prof. S. W. Chisholm, Cambridge, MA, USA (1993) Prof. T. Fenchel, Helsingør, Denmark (1985) Dr. N. S. Fisher, Stony Brook, NY, USA (1985) Prof. B.-O. Jansson, Stockholm, Sweden (1989) Prof. B. B. Jørgensen, Bremen, Germany (2004) Prof. D. M. Karl, Honolulu, HI, USA (2006) Prof. L. Legendre, Villefranche-sur-Mer, France (2002) Prof. E. Naylor, Menai Bridge, UK (1984) Prof. S. W. Nixon, Narragansett, RI, USA (1989)

Prof. W. Nultsch, Hamburg, Germany (1994) Prof. R. T. Paine, Seattle, WA, USA (1990) Prof. D. Pauly, Vancouver, Canada (2007) Dr. T. Platt, Dartmouth, NS, Canada (1984) Acad. Prof. G. G. Polikarpov, Sevastopol, Ukraine (1985) Dr. F. Rassoulzadegan, Villefranche-sur-Mer, France (1997) Prof. V. Smetacek, Bremerhaven, Germany (1993) Prof. A. Underwood, Sydney, Australia (2005) Prof. B. B. Ward, Princeton, NJ, USA (2006) Dr. M. Weinbauer, Villefranche-sur-Mer, France (2004)

Terrestrial Ecology Prof. T. N. Ananthakrishnan, Chennai, India (1984) Prof. R. Barbault, Paris, France (2007) Prof. F. A. Bazzaz , Cambridge, MA, USA (1998) Prof. F. S. Chapin, III, Fairbanks, AK, USA (1986) Prof. J. Ehleringer, Salt Lake City, UT, USA (1986) Dr. P. Ehrlich, Stanford, CA, USA (1994) Prof. M. Gadgil, Bangalore, India (1985) Prof. G. Glatzel, Vienna, Austria (1995) Prof. I. Hanski, Helsinki, Finland (1993) Prof. A. Hector, Zürich, Switzerland (2006) Prof. Robert D. Holt, Gainesville, FL, USA (2005)

Prof. Ch. Körner, Basel, Switzerland (2006) Prof. E. Kuno, Kyoto, Japan (1986) Prof. J. Lawton, Heslington, UK (1997) Prof. M. Loreau, Montreal, QC, Canada (2003) Prof. A. P. Møller, Paris, France (1995) Prof. H. A. Mooney, Stanford, CA, USA (1991) Prof. C. Nilsson, Umea, Sweden (2006) Prof. B.-E. Saether, Trondheim, Norway (2000) Prof. D. Schimel, Boulder, CO, USA (1998) Dr. M. Shachak, Sede Boker, Israel (1989) Prof. M. G. Turner, Madison, Wisconsin, USA (2005) Dr. D. H. Wall, Fort Collins, CO, USA (2005) Prof. E. O. Wilson, Cambridge, MA, USA (1988)

*Following their receipt of the ECI Prize document, laureates are invited to join the institute’s staff

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Limnetic Ecology Prof. N. V. Aladin, St. Petersburg, Russia (1994) Prof. J. Benndorf, Dresden, Germany (2009) Prof. T. Berman, Tiberias, Israel (1998) Prof. J. Cairns, Blacksburg, VA, USA (1998) Prof. S. R. Carpenter, Madison, WI, USA (2000) Prof. J. J. Cole, Millbrook, NY, USA (2004) Prof. P. W. Cullen, Gunning, NSW, Australia (2000) Prof. J. I. Furtado, London, UK (1985) Prof. Z. M. Gliwicz, Warsaw, Poland (1998) Prof. N. G. Hairston, Ithaca, NY, USA (1998) Dr. J. E. Hobbie, Woods Hole, MA, USA (1986) Prof. Ewa Kamler, Piaseczno, Poland (1993) Prof. W. Lampert, Plön, Germany (1993) Prof. G. E. Likens, Millbrook, NY, USA (1989) Prof. K. Lillelund, Hamburg, Germany (1985) Dr. D. M. Livingstone, Duebendorf, Switzerland (2009)

Prof. B. Moss, Liverpool, UK (2009) Prof. J. Overbeck, Merkendorf, Germany (1984) Prof. T. J. Pandian, Madurai, India (1985) Prof. E. Pieczy´nska, Warsaw, Poland (1993) Prof. M. Power, Berkeley, CA, USA (2006) Dr. C. S. Reynolds, Ambleside, UK (1995) Prof. D. W. Schindler, Edmonton, Alberta, Canada (2006) Dr. T. Simé-Ngando, Aubière, France (2009) Prof. R. Sommaruga, Innsbruck, Austria (2008) Prof. M. Søndergaard, Hillerød, Denmark (2009) Prof. R. Sterner, St. Paul, MN, USA (2005) Prof. L. Tranvik, Uppsala, Sweden (2005) Prof. J. G. Tundisi, São Paulo, Brazil (1990) Prof. D. Uhlmann, Dresden, Germany (1989) Prof. W. Wieser, Innsbruck, Austria (1987) Prof. M. Yamamuro, Tokyo, Japan (2009)

OKF Trustees 2009 Dr. Anna F. Pasternak (President) Moscow, Russia

Prof. F. Buchholz (Vice President), Germany Prof. J. Padisák, Vezprém, Hungary Prof. M. Plinski, Gdynia, Poland Dr. M. Pavlicev, St. Louis, MO, USA A. Fromm (Administration), Germany

Michel Loreau: Recipient of the Ecology Institute Prize 2002 in Terrestrial Ecology. A Laudatio Anders Pape Møller Laboratoire de Parasitologie Evolutive, CNRS UMR 7103, Université Pierre et Marie Curie, Bât. A, 7ème étage, 7 quai St. Bernard, Case 237, 75252 Paris Cedex 05, France

The vast number of living organisms and the resulting potentially immense ways of interactions among these species are puzzling fact for lay men and scientists alike. The field of community ecology attempts to understand the assembly of communities and the ways in which such assembly through interactions among species and interactions with the environment affect the functioning and the composition of the community. Michel Loreau has had an exceptional influence on bringing order to this seemingly chaotic field. Creation of order is a first step to allow scientific enquiry into underlying processes and resultant outcomes. Such enquiry, it being theoretical, observational or experimental, is the only way to make scientific progress. Michel Loreau has managed to tread the first steps that have allowed a large number of collaborators and colleagues to follow in their endeavor to better understand the complex web of ecological interactions in communities. The discipline of community and ecosystem ecology has experienced dramatic fluctuations in success and popularity. A period of 15 years from the late 1950’s to the early 1970’s was extraordinary heydays by providing a tremendous burst in new theories about communities and how they came about. Robert MacArthur’s classical studies of parulid warblers was followed by a whole suite of disciples such as M. Cody, J. Diamond, S. Levin, J. Terborgh, E. O. Wilson, and many others. The level of enthusiasm for community ecology could barely find room, until the discipline almost faded away during the 1970’s. In a famous blues song entitled “Martin Cody’s blues”, published in the obscure ecological publication Edema in Finland, this demise of the field, and the stark contrast to the heydays of the 1960’s, was lamented. Light first appeared again when a new generation of ecologists, with Michel Loreau at the front started to make simple mathematical models attempting to explain how different species, with different ecological roles, affected the abundance, composition and functioning of communities (e. g. Loreau 1998). These models were breakthroughs, as is often the case for theoretical work, because they in a

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simple, but elegant way managed to distill the essentials from an ecological cocktail of apparent chaos. The assumptions and predictions of these models were directly amenable to experimental falsification, and this opened up for a level of understanding beyond mere description. Order had appeared from chaos, and experimentation based on a theoretical approach was once again the way forward. Loreau, M. 1998. Biodiversity and ecosystem functioning: A mechanistic model. Proceedings of the National Academy of Science USA 95:5632–5636

Preface This book ventures off the beaten track. It provides an overview of what I see as the main challenges of biodiversity science today, and, in doing so, it drifts inexorably beyond the traditional bounds of the scientific enterprise. Most scientists would probably view research as their primary mission, and teaching as their secondary mission. Accordingly, most scientific books either report new research findings or are textbooks tailored for students. Yet, the social practice of science is much broader and multifaceted than this. Thus, science also involves communicating to the general public. This activity yields yet another type of books, that of popular scientific books. Perhaps less widely appreciated is the fact that science is also a collective activity performed by a community of scientists who interact, deliberately or unwittingly, to develop new concepts, new ideas, and new directions in research. Much of this collective activity is invisible, sometimes even to the scientists themselves. Successful scientists are traditionally portrayed as solitary geniuses; anything that suggests otherwise tends to be hidden or downplayed. Yet, in my own research, I have always been struck by the fact that new ideas often pop up simultaneously in different minds and in different places as if by spontaneous generation. The scientific genius is the one who is first to crystallise these new ideas in a particularly clear and innovative way. The generation of new scientific ideas is truly spontaneous in the sense that it is not the outcome of some sort of preconceived plan, but it is nevertheless often a reasonably predictable process, which can be channelled through collective means such as conferences, workshops and research programmes. Organised collective activity is particularly important in the environmental sciences, where new findings often have important implications for policy decisions. Creating appropriate processes through which scientific knowledge is generated, synthesised beyond differences in opinion, and made available to policy can be viewed as an integral part of the social mission of science, and this task requires some level of collective organisation. For these collective facets of the scientific process, there is relatively little appreciation overall, and no special category of books. Although, like most of my peers, I have devoted much of my scientific career to research and teaching, I have also had an interest in the organisation and transformation of science as a social practice. Thus, when the controversy over the interpretation of biodiversity and ecosystem functioning experiments was raging, I took the initiative to organise, together with my colleagues Shahid Naeem and Pablo Inchausti, an international conference to resolve differences in opinion among the major players in this field. This conference resulted in a book (Loreau et al. 2002) and two influential consensus papers on biodiversity

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and ecosystem functioning (Loreau et al. 2001; Hooper et al. 2005). At the same time, I set up and then chaired a European Science Foundation programme, Linking community and ecosystem ecology, to address what I saw as a major research need in the field of biodiversity and ecosystems. Based on this freshly acquired experience, the International Council for Science asked me to chair a Task Force commissioned to rethink and reorganise DIVERSITAS, the international programme of biodiversity science. I took up the challenge, and this led me to one of the most rewarding experiences in my career. During seven years (first as chair of the Task Force, then as chair of the DIVERSITAS Scientific Committee) I had the opportunity to lead the emergence and development of a new biodiversity science transcending disciplinary boundaries. This venture eventually brought me to the boundary between science and policy, when I came to chair the international conference Biodiversity science and governance and then the international consultative process toward the establishment of a new International Mechanism of Scientific Expertise on Biodiversity, which resulted from this conference. When offered the opportunity to author a book in the series Excellence in Ecology, I first considered, quite naturally, writing a book about my research linking community ecology, evolutionary ecology, and ecosystem ecology. But gradually I came to the conclusion that to some extent this would be a missed opportunity because I could easily publish a research-oriented book elsewhere. In contrast, the series Excellence in Ecology allowed me to publish a book with a more personal touch that I probably would not have even considered writing otherwise. Therefore, I decided to publish a synthesis of my research work at Princeton University Press (Loreau 2010), and to focus this book on my vision of biodiversity science as it emerged from my experience at the head of DIVERSITAS and other international initiatives. The result is a book that does not fit neatly into traditional categories. My book is written in a style that should make it accessible to a broad audience of scientists, students, decision makers, managers, and indeed any person who has a scientific background and an interest in biodiversity. The first chapter serves as an introduction; it explains how and why an integrative biodiversity science is emerging today. The rest of the book examines the huge but exciting challenges that biodiversity science is facing. There are many challenges. The first requirement for any science is to define the subject it studies and how it studies it. The multifaceted nature of biodiversity constitutes a basic challenge for the science of biodiversity. In Chapter 2, however, I show that some order can be brought into the apparent chaos of the many different approaches to biodiversity. My goal in this chapter is to present some unifying frameworks to define and measure biodiversity while maintaining the intrinsically diverse nature of biodiversity. A second requirement for the emerging biodiversity science is to

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clearly explain why biodiversity is an important issue today, and hence why a science of biodiversity is vitally needed. This is the objective of the following two chapters. Chapter 3 summarises our current knowledge about the extent and causes of the current global biodiversity crisis. It shows how we are heading for a sixth mass extinction in the history of Earth because of the current rapid expansion of the human population and economy. Chapter 4 then examines the potential consequences of biodiversity loss for human societies. These potential consequences are much more important than appears at first sight because of the multiple ways in which humans depend, directly or indirectly, on biodiversity. Thus, the first part of the book contains the basic scientific information that is necessary to understand what biodiversity is, how it can be measured, how and why it is threatened, and why it matters. As a consequence, it overlaps with material that can be found in other books, such as Wilson’s (1992) classic The diversity of life, and Gaston and Spicer’s (2004) textbook Biodiversity: an introduction. The overlap is only partial, however, as I provide my own perspective on these topics as well as some new approaches to address them. The second part of the book is definitely non-traditional. In Chapter 5, I turn to what I see as the main scientific challenges for the emerging biodiversity science. Although there is broad consensus on the main causes and consequences of biodiversity changes, there are still major gaps in our knowledge that prevent current biodiversity science from being predictive. Chapter 5 examines these knowledge gaps and offers some avenues to fill them. But the challenges of biodiversity science are not just scientific. As a science that arises from a crisis situation, biodiversity science must also enter into dialogue with policy and society at large if it is to contribute to the resolution of this crisis. This is a major challenge, which brings biodiversity science on, and at times even beyond, the dividing line between science and other human social practices. I address this challenge in the last two chapters. Chapter 6 discusses the needs and ways to strengthen the links between biodiversity science and policy, in particular through the establishment of an intergovernmental process akin to the IPCC for climate change. Chapter 7 concludes this book with a vibrant call for an indispensable re-examination of the philosophical and economic premises on which modern society is built. I show that the current biodiversity crisis is the historical result of the fundamental divorce between humankind and nature that is deeply rooted in modern philosophy, science and economy, and that a reconciliation of humans with their own nature is key to resolving this crisis. I also argue that biodiversity science can contribute to reaching this goal by fostering the emergence of a genuine evolutionary ecology of the human species. Human evolutionary ecology can help understand the creation and maintenance of the deepest needs and values in humans, and thereby contribute to linking the var-

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ious forms of knowledge and to establishing the bases for a new philosophy and a new ethics that fully acknowledge humans as part of nature. I am aware that some of the material contained in the last two chapters may be controversial. But I strongly believe that science will be able to contribute to the resolution of the major environmental issues that humankind is facing today only if it does not stay shut away in an ivory tower and if it is open to a dialogue with the rest of society and other forms of knowledge. In no way do these chapters intend to provide the final word on these difficult topics. Instead, they are meant to offer food for thought on key issues that are too often ignored to my mind in discussions of the biodiversity crisis. Although I am the only one to blame for the ideas expressed in this book, I am particularly grateful to Claire de Mazancourt, Andy Gonzalez, Anne Larigauderie, Shawn Leroux, Greg Mikkelson and Meryl Williams for providing helpful comments on the manuscript on very short notice. I also thank Rob Colwell, Claire de Mazancourt, Helen Elina, Andy Hector, Allen Larocque, Chris Steiner, Maja Weilenmann and Boris Worm for kindly providing data or helping to design some of the figures.

1 THE EMERGENCE OF BIODIVERSITY SCIENCE Biodiversity is a broad concept that encompasses the variety of life in all its manifestations. The term itself is quite recent — it was born during a US National Forum on BioDiversity in 1986 (Wilson 1989). Since then it has gained popularity with phenomenal speed, both in the scientific community and in the general public. The number of scientific papers concerning biodiversity increased abruptly after the term was coined, and shows steady exponential growth during the last 15 years (Fig. 1). The term is now also commonly used in the media. Yet the concept of biological diversity — of which the neologism “biodiversity” is simply a contraction — is much older. Patterns of species diversity and the processes that shape them had been studied by community ecologists decades before the term “biodiversity” was first coined. The maintenance of genetic diversity within populations had been a familiar topic in population genetics for a long time, just as had the diversification of evolutionary lineages in paleontology and phylogenetics. The scientific literature on biological diversity is still today much broader than the one referring explicitly to the term “biodiversity”, although the latter is gradually closing the gap (Fig. 1). What, then, made the new term “biodiversity” different? Why did it become so popular, and spread well beyond the boundaries of science? Many factors have contributed to its success, but the main one, I believe, is that it helped unify the concept of diversity that was used in different scientific disciplines as well as the message that scientists were conveying to society. Biodiversity is a common banner under which scientists from all disciplines can gather as a community and speak to the general public about the threats that hang over life on Earth as a result of the fast-growing impacts humans exert on their environment. At first, perhaps paradoxically, the new term had more impact on society than on science itself. The public message grew stronger, and a few years later, in 1992, the international Convention on Biological Diversity was created during the United Nations Conference on Environment and Development in Rio. The international scientific community started to organise itself at roughly the same time. It established DIVERSITAS, an international programme devoted to fostering scientific research into all the dimensions of biodiversity and biodiversity loss, in 1991. DIVERSITAS and other international initiatives were instrumental in identifying and launching new research themes. In particular, they identified the relationship between biodiversity and ecosystem functioning as an important topic to explore (Schulze and Mooney 1993). This topic gave birth to one of the hottest research fields in ecology and environmental sciences

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Fig. 1. Growth of the scientific literature on biodiversity during the last 25 years. Curve A shows the log-transformed number of papers that use the term “biodiversity” in their title, abstract or keywords. Curve B shows the log-transformed number of papers that use either the term “biodiversity” or the term “diversity” in their title, abstract or keywords, after exclusion of the papers that deal with medicine, humanities, pure social sciences, computer sciences, and all natural and engineering sciences that do not have any biological component. A constant slope on the log scale indicates exponential growth. Data collected from the Web of Science

during the next decade (Loreau et al. 2001, 2002; Kinzig et al. 2001; Hooper et al. 2005; Naeem et al. 2009). Despite the impetus and new directions provided by these international initiatives, the global biodiversity scientific landscape was still relatively fragmented in the 1990s. Taxonomists focused on the “taxonomic impediment”: the global decline in the number of taxonomy experts and its dire consequences for the discovery and inventory of global species diversity. Ecologists sought to understand the ecological processes that lead to the maintenance or loss of species diversity and the functional consequences of biodiversity in ecosystems. Conservation biologists discussed the best ways to protect endangered species and set up nature reserves. A few economists were exploring new ways to value biodiversity and assess its economic consequences for human societies. In short, scientists in each scientific discipline were concerned by the questions that delimit their own discipline.

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Fragmentation was glaring even within each scientific discipline. For example, in ecology, terrestrial ecologists, freshwater ecologists and marine ecologists rarely mixed — and still rarely mix today — each group having its own traditions, journals and scientific societies. Similarly, microbial ecologists, plant ecologists and animal ecologists constituted distinct scientific communities — and still largely do, even though differences are blurring gradually. Thus, a wide range of scientific disciplines and subdisciplines were involved in biodiversity research, but there was no unified biodiversity science agenda. This situation was reflected in the functioning of DIVERSITAS, which was conceived as an umbrella programme covering various activities sponsored by different organisations. Scientific diversity mirrored biological diversity. The common theme of biodiversity had produced a diversity of biodiversity-related sciences, not an integrated biodiversity science. Scientific diversity is to some extent inevitable. It is even a source of wealth and dynamism, just as is biological diversity itself. But the benefits of diversity can only be reaped if it contributes to a common enterprise. Fragmentation typically leads to lack of cohesion, impaired functioning, and even polarisation between groups with conflicting viewpoints. Efficient functioning requires unity in diversity, or unity through diversity. The question is: Is such unity feasible? Can there be an integrated biodiversity science given the wide range of scientific disciplines involved in biodiversity research? My answer to these questions is yes, and my opinion is based, not just on abstract thinking, but also on my real experience with DIVERSITAS. After a difficult period in which the programme went almost extinct, we reorganised it completely in 2001 around the concept of integrated biodiversity science and around a unified set of objectives and activities. Since then DIVERSITAS has flourished, and has become a major international player, not only in the biodiversity scientific arena, but also in the biodiversity policy arena. For me, this is concrete proof that, although biodiversity science is still in its infancy, it has real substance and tremendous potential. Its power comes from the fact that it provides unity of purpose through integration of different perspectives. What, then, is the nature and scope of biodiversity science? Clearly, biodiversity science is not a new stand-alone scientific discipline. Biodiversity science deals with all aspects of biodiversity, including its natural and social aspects, and thus is interdisciplinary by nature. It is a science that arises from the need to provide broader perspectives on issues of societal relevance that cannot be addressed by traditional scientific disciplines in isolation. In this respect, it is similar to other new integrative approaches in environmental sciences, such as Earth system science and sustainability science. It becomes increasingly clear that global environmental problems in general, and biodiversity loss in particular, cannot be solved by traditional disciplinary approaches.

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Environmental problems have to do with the relationship between humans and their natural environment, and hence are inherently social and natural simultaneously. Even their social or natural dimensions cannot be reduced to a single discipline. The mere prediction of global climate change, for instance, is not the prerogative of climatologists, but requires the collaboration of an increasing range of physical, chemical and biological sciences because of the multiple physical, chemical and biological feedbacks that operate in the climate system. Similarly, understanding the dynamics of biodiversity and its multiple natural and social causes and consequences is a daunting task that can only be faced by a broad coalition of scientific disciplines. To better understand this challenge, let us briefly consider why biodiversity is an issue today, and what kind of scientific information is needed to address this issue. We shall examine various components of this issue in greater detail in subsequent chapters. Interest in biodiversity as a unifying concept began with the realisation that biodiversity is being lost globally at accelerating rates, that this loss may have important long-term consequences, and hence that measures should be taken to halt or reduce it. This summary statement, which is now accepted virtually universally by the scientific community, contains three main propositions. The first asserts that biodiversity is being lost, or more generally changes (since biodiversity may increase at some spatial and temporal scales), at alarming rates. To substantiate this proposition and make it more precise, detailed information is needed on the extent of existing biodiversity, on the ecological and evolutionary processes that drive biodiversity changes under natural conditions, on the factors that drive present changes, and on the rates at which these changes occur in comparison with the past. Answering these questions is a huge task that is well beyond the reach of a single scientific discipline. Addressing the natural science component of these questions alone requires at least the contributions of ecology, taxonomy, systematics, evolutionary biology, genetics, and paleontology. But it turns out that the ultimate factors that drive current biodiversity changes are social: biodiversity is being lost today because of the growing environmental impacts of human activities, and these in turn are governed by patterns of human population growth, production, consumption and trade. Therefore biodiversity changes cannot properly be understood and predicted without consideration of the global social and economic system, which requires the contributions of social sciences such as economics, politics, law, sociology, anthropology, and psychology. The second proposition in the above statement speaks about the potential consequences of biodiversity loss. Although these consequences were largely speculative until recently, we now have strong evidence that biodiversity changes affect the functioning of managed and unmanaged ecosystems, their

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ability to deliver ecosystem services to humans, and many aspects of human well-being. Understanding and predicting the ecological consequences of biodiversity changes are tasks that fall first and foremost to ecology, but as soon as human well-being is concerned, health sciences, economics, and other social sciences such as sociology, anthropology, and psychology enter the scene. The third and final proposition in the above statement is perhaps that which is most evidently cross-disciplinary. Conserving and managing biodiversity in ways that prevent or minimise further biodiversity loss call for policy decisions that involve economic, management and political sciences. These decisions must be informed by sound knowledge coming from the natural sciences, and assessed by their effectiveness at maintaining biodiversity, a task that again falls to natural sciences. Disciplines such as ecology, conservation biology and population genetics are directly concerned by such assessments. Furthermore, these various components of the biodiversity issue interact with each other. Biodiversity changes, the consequences of these changes, and the ways to prevent or mitigate them are obviously interrelated. In particular, there is a cycle of interactions that connects biodiversity changes to their consequences, their consequences to policy decisions, and policy decisions back to biodiversity changes (Fig. 2). Understanding and resolving the global biodiversity crisis can only be achieved by considering the totality of this cycle of interactions. Although natural scientists may focus on biodiversity itself, its changes and its ecological consequences, they cannot ignore society as both a cause and a potential solution to the phenomena they study. Similarly, social scientists may focus on the way society perceives, values and manages biodiversity, but only if they take into account the natural dynamics of biodiversity and its effects. Thus, a cross-disciplinary approach is needed for both the description and the solution of the biodiversity crisis. Integration across scientific disciplines is particularly critical and challenging for the study of biodiversity loss in comparison with many other environmental issues. Take for example two other major environmental challenges that have attracted a lot of attention so far, stratospheric ozone depletion and climate change. Stratospheric ozone depletion — the so-called “ozone hole” — is essentially a problem of atmospheric chemistry with a simple anthropogenic cause (the release of chlorine- and bromine-containing chemicals by a few industrial sectors) and straightforward consequences for human health and ecological systems through increased levels of ultraviolet radiation. Although a comprehensive assessment of the causes and consequences of stratospheric ozone depletion required the collaboration of several natural and social sciences, the issue was simple enough that its study could be organised around a single discipline, atmospheric chemistry. Climate change is a much more complex environmental issue with a broad array of causes, mechanisms, and conse-

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Fig. 2. The main components of the current biodiversity crisis and their interactions

quences involving natural systems and human societies. Therefore the study of climate change has mobilised an increasingly broad coalition of disciplines from both the natural and the social sciences. This coalition, however, revolves around a core discipline — climatology — and even a core mathematical tool — general circulation models, which describe the main physical features of the global climate. The role of general circulation models in condensing information coming from a wide range of disciplines around a physical nucleus is largely responsible for the success of the climate change scientific community in presenting a common front and a set of clear quantitative predictions to policy makers and the general public. What is perhaps less apparent to the layperson is that this strength is also a weakness. Like any model, general circulation models are simplifications of reality that focus on some processes and properties but ignore others. These models do a very good job at incorporating the main physical and chemical constraints on the global climate, but they still do a rather poor job at incorporating the biological, ecological, and social feedbacks that affect the global climate. Biodiversity loss is an even more complex and challenging environmental issue than is climate change. Biodiversity is a multifaceted, multi-scale property of biological systems that is affected by and affects a wide range of natural and social processes at all scales. Although the study of its proximate causes

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and consequences seems to fall naturally to ecology, and although ecology has indeed played a central role in the emergence of biodiversity science, the range of issues and processes involved in biodiversity loss is simply too broad to be tackled appropriately using a single disciplinary approach. The mere documentation of past and present biodiversity requires the collaboration of ecology, systematics, evolutionary biology, and paleontology. The persistence or conservation of small populations hinges not only on ecological constraints, but also on genetic constraints, the understanding of which requires the contribution of population genetics. In some cases, resource management and policy decision interact so closely with changes in biodiversity that decoupling the natural system from the social system would be counterproductive. The creation of an integrative biodiversity science is unquestionably a challenging enterprise, but it is also an exciting opportunity to build a new approach that has an interdisciplinary nature from the outset. The emergence of an integrative biodiversity science, however, in no way diminishes the importance of the various disciplinary sciences that contribute to it. Strong interdisciplinary approaches require strong disciplinary foundations. There is sometimes temptation to compromise on the quality and rigour of the sciences that are used to build an interdisciplinary framework. But this is a recipe for failure. In fact, successful interdisciplinary approaches are more demanding than disciplinary ones. Not only do they require thorough knowledge of the various disciplines involved, they also require developing new conceptual and theoretical frameworks at the interface between these disciplines. Failure to meet these requirements often results in weak science, and therefore also in ineffective policy and management. Biodiversity science does not substitute itself for the various sciences that study biodiversity, it connects and integrates them into a coherent framework and a coherent set of questions and objectives.

2 WHAT IS BIODIVERSITY? A Diversity of Diversities I started the previous chapter with a broad definition of biodiversity as the variety of life in all its manifestations. My task now is to make this definition more concrete and more precise. What are the manifestations of life that make it so diverse? Life is a mode of existence of matter and energy characterised by the formation of extremely complex, self-organised dissipative systems maintained by external sources of free energy and able to self-reproduce. The basic unitary entity of life is the individual organism. An organism usually has clear physical boundaries; it functions and reproduces as a whole. Organisms show an astonishing diversity of forms, from relatively simple unicellular prokaryotes to highly complex multicellular eukaryotes, including higher plants and animals. Most of this diversity is readily apparent and has for a long time been the basis for the classification of life. Therefore the diversity of types of organisms, or organismal diversity, is the most common way in which biodiversity is defined and measured in practice. More specifically, since organisms of the same type are grouped into species, organismal diversity is generally expressed in the form of species diversity, by far the most widely used definition of biodiversity in both the natural and the social sciences. Organismal diversity, however, is a broader concept than species diversity because it potentially encompasses other differences between organisms. First, species diversity is but a subset of the more general concept of taxonomic diversity, i.e., the variety of units in the taxonomic classification, which includes many other hierarchical levels than the species level, such as subspecies, genera, families, orders, classes, and phyla. The classification of life is still very much a work in progress. New organisms are being discovered continuously, leading to constant refinement and at times major changes in the taxonomic classification. Taxonomic diversity is a way to capture the diversity of types of organisms that is not bound exclusively to a single taxonomic level. Second, as we shall see below, organisms differ in many other ways than in their taxonomic label, which opens up the possibility of defining and measuring other forms of diversity. By analogy with organismal diversity, a diversity of structures can also be defined at both lower and higher levels of the biological organisation. Moving down to lower levels of organisation, a multicellular organism is composed of organs, which themselves are composed of cells, which in turn are composed of molecules. Because there are a relatively limited number of different types of

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organs and cells, biodiversity has yet to be defined at these levels of organisation even though this is a distinct possibility. By contrast, a huge variety of molecules contribute to the structure and metabolism of living organisms. Molecular diversity is an inherent property of life. Among these molecules are the deoxyribonucleic acids, which carry the genetic information necessary to ensure the functioning and reproduction of the whole organism in the form of sequences of elementary nucleotides called genes. Because the number of potential sequences of nucleotides is virtually infinite for all practical purposes, the diversity of these genes, or genetic diversity, is also virtually infinite. And since the information that encodes other molecules in living cells is contained in the genes, genetic diversity underpins much of the diversity of living organisms. Humans have exploited this genetic diversity since the advent of agriculture to get thousands of different varieties of each of the main crops and select some of those fittest for human use. Therefore genetic diversity is often viewed as the most fundamental form of biodiversity. But genetic diversity may also be viewed alternatively as a specific form of molecular diversity, a broader concept that includes the variety of all molecules that partake in life. Moving up to higher levels of organisation, each individual organism is both part of a population of interbreeding organisms from the same species and part of a remarkably complex network of interactions between organisms from different species and their abiotic environment, called an ecosystem. These higher-level complex ecological systems also show a great variety of forms at a wide range of spatial scales. Tropical rainforests, temperate grasslands, lakes and coral reefs, for instance, are types of ecosystems that differ strikingly not only in their physical environment, but also in the types of organisms they contain and in the way these organisms interact with each other and with the physical environment. At smaller scales, different patches of the same forest may contain different species, or similar species but with different relative abundances and different strengths of interactions. The manifold differences between populations and ecosystems are often lumped together under the generic term ecological diversity, although there is an obvious potential for more subtle distinctions here. Ecological diversity is both the cause and the consequence of the maintenance of genetic and organismal diversity in a heterogeneous world. It also underpins the diversity of interactions between humans and nature. Cultural diversity is a term often reserved for describing the variety of cultures in human populations, but it can also apply to other social animals. Genetic diversity, organismal or species diversity, and ecological diversity are the three levels of biodiversity that are distinguished conventionally (Heywood and Baste 1995; Gaston and Spicer 2004). I have already alluded to some of the gaps of this traditional classification above. The most important limita-

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tion of traditional approaches to biodiversity, however, is that they reduce, implicitly or explicitly, diversity to a counting issue. Diversity at each level of the biological organisation is often equated with the number of distinct entities at that level. Although there are many problems with the definition of the species concept, a species is in principle a well-defined, countable entity. Therefore inventorying and counting species is undoubtedly the most straightforward way to describe and measure organismal diversity. Similarly, each gene is a specific sequence of nucleotides; the number of existing sequences can be counted in principle, even though this number may be extremely large. The spatial, temporal, and even conceptual boundaries of the ecosystem concept are more blurred. But once a consistent specific set of criteria is used, ecosystems are also, in principle, countable entities. Biodiversity, however, is a broader and richer concept than the mere number of different biological structures of a kind. The richness or multiplicity of elements in an ensemble is only one component of its diversity. Another component of diversity, which has been formally distinguished from richness in ecology, is the evenness or equitability of the elements of an ensemble. A temperate forest is typically less diverse than a tropical forest not only because it harbours fewer species, but also because a larger proportion of total plant density or biomass is concentrated in a single or a few dominant tree species, even after accounting for differences in species numbers. A third component of diversity, which has not been formally identified and yet which is gaining growing attention in ecology, is the disparity or heterogeneity of the elements of an ensemble. The concept of disparity or heterogeneity encapsulates the intrinsic, qualitative differences that make elements distinct from one another. Some authors have proposed to reduce the concept of diversity to its quantitative components alone, and make disparity a separate concept (Maclaurin and Sterelny 2008). Although greater precision is useful, I believe that this reduction of the diversity concept emasculates it. Disparity is a key component of biodiversity. In fact, many of the values of biodiversity result from qualitative differences between species, not species number per se. Qualitative differences between biological systems are manifold. Differences between organisms are often classified into two main categories, in line with evolutionary theory: genotypic differences are differences in their genetic constitution, while phenotypic differences are differences in any observable trait of organisms, including their morphology, development, physiology, and behaviour. Since genotypic differences can be ascribed directly to genetic differences at the molecular level, this source of variation between organisms is already taken into account in the concept of genetic diversity. By contrast, phenotypic differences between organisms are not included in the concepts of taxonomic diversity and species diversity, which only account for the number

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and relative abundances of different units in the taxonomic classification. Phenotypic or trait diversity is a key component of organismal diversity that governs many of the ecological consequences of species diversity because the traits of the various species determine both their impacts on ecosystem functioning and their responses to endogenous and exogenous environmental changes (Chapter 4). The term functional diversity is sometimes used to denote the diversity of traits that govern the effects of organisms on ecosystem functioning or their responses to environmental changes. Phylogenetic diversity provides an alternative way to describe the disparity of organisms. Phylogenetic diversity accounts for the different evolutionary histories of the various species. Since a species’ evolutionary history determines its current genotype and phenotype, this form of diversity overlaps with both genetic diversity and phenotypic diversity. Ecological diversity is largely driven by qualitative differences between ecosystems. Although ecology has often focused on species diversity, the main properties of ecosystems are, in fact, governed by the way species interact with each other and with their physical environment. Patterns of species interactions are typically more important than the mere number and relative abundance of species. The diversity of ecological interactions and processes, which lies at the core of ecological diversity, has received relatively little attention as such, although it has obvious connections to a number of descriptors of food webs and other ecological networks. A recent attempt to move beyond the focus on species diversity in ecology proposes to distinguish the horizontal and vertical dimensions of ecological diversity (Duffy et al. 2007; Loreau 2010). Horizontal diversity is the diversity of organisms that belong to the same trophic level or functional group (e.g., plant diversity, herbivore diversity, or decomposer diversity), and is essentially equivalent to organismal diversity within a trophic level or functional group. By contrast, vertical diversity is the diversity of trophic levels or functional groups in an ecosystem (ecological differentiation between plants, herbivores, carnivores, parasites, decomposers, etc.). But many other aspects of ecological diversity are not captured by this dichotomy and would deserve more attention. This brief examination of the biodiversity concept shows clearly that this concept is multifaceted. This diversity of diversities can be organised conceptually along two main orthogonal dimensions (Fig. 3). One dimension describes the level of biological organisation that defines the basic elements or aspects whose diversity is being considered. The three main levels of organisation that are considered traditionally are molecules, organisms and ecological systems (populations and ecosystems). The corresponding forms of diversity are molecular, organismal, and ecological diversity. Genetic diversity and cultural diversity may be viewed as specific forms of molecular diversity and eco-

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Fig. 3. The various levels of organisation and components that define the multiple facets of biodiversity

logical diversity, respectively, in this classification. The second dimension describes the various components of diversity at each level of organisation, i.e., richness, evenness, and disparity. This dimension currently distinguishes different forms of diversity only at the organismal level. Species diversity and taxonomic diversity include the richness and evenness components of organismal diversity, while phenotypic diversity and phylogenetic diversity encapsulate its disparity component. The concepts of molecular diversity and ecological diversity are currently broad enough to encompass the richness, evenness, and disparity components of diversity at the sub-organismal and supra-organismal levels, respectively. This classification seeks to organise the most common existing concepts but does not pretend to be exhaustive. More specific forms of diversity could easily be defined at the molecular and ecological levels. Cultural diversity could also easily be separated from ecological diversity by distinguishing the population and ecosystem levels. Note that each form of diversity defined at one level of organisation can also be studied at higher levels of organisation. The vertical axis in Fig. 3 shows the level of organisation at which the basic elements of each form of diversity are defined, but the diversity of these elements can then easily be scaled up to higher levels of organisation. For instance, genetic diversity is ascribed to the molecular level because genes are defined at this level, but it is also commonly studied at the organismal and ecological levels since organisms, populations and ecosystems alike differ in their genetic make-up. Similarly, species diversity is traditionally used in comparative studies of communities and ecosys-

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tems. Phylogenetic and trait-based approaches are receiving growing attention in community ecology and ecosystem ecology (Webb et al. 2002; McGill et al. 2006; Suding et al. 2008) and provide new opportunities to use phylogenetic diversity and phenotypic diversity as predictors of community and ecosystem processes (Heemsbergen et al. 2004; Cadotte et al. 2008).

Taking the Measure of Biodiversity Since the concept of biodiversity is multifaceted, one might expect its measurement to be a fairly complicated matter. In fact, the problem is at the same time simpler and even more complicated than expected. It is simpler because once an appropriate measure has been devised for one component of diversity at one level of biological organisation, it can also be used at other levels of organisation. It is more complicated because there is no universally accepted measure of diversity. Just as the concept itself, the measurement of biodiversity is multifaceted. There is a wealth of different approaches and measures in the scientific literature, and even entire books on this topic (Magurran 2004). Despite this complexity, however, there are some unifying approaches that bring order in this apparent chaos. Ecology has a long tradition of measuring species diversity. Therefore I will focus here on species diversity, provide a brief summary of the most common ways to measure it, and present a formal framework that unifies many of these measures. The same framework can easily be applied to other levels of organisation. Although the treatment that I provide of this topic in this section is fairly synthetic, it inevitably involves some technical considerations. Therefore, readers uninterested in technical details might wish to skip them or move directly to the next section. At first sight, it might seem that once biodiversity has been reduced to a single component — here, species diversity — the task of measuring it should be relatively straightforward. In many taxonomic groups species are rather well defined entities; therefore the problem seems to boil down to counting the number of such entities. Unfortunately, this simple solution quickly turns out to be impractical. The extent of the difficulty can be most clearly illustrated by the extent of our current ignorance about the total number of species on Earth. About 1.5 to 1.8 million species have been discovered — we do not even know exactly how many because specimens are spread all over the world and many lack detailed description. But based on various extrapolations from current knowledge of well-studied geographical areas and taxonomic groups and patterns of species description through time, the number of extant species is estimated to lie somewhere between 3.5 and 111.5 million species, with a working

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estimate of 13.5 million species (Hawksworth and Kalin-Arroyro 1995). Thus, our ignorance spans two orders of magnitude, and even the most reasonable working estimate implies that we have yet to discover almost ten times as many species as those described so far. Of course, some geographical areas and some taxonomic groups are better known than others. For obvious reasons, temperate faunas and floras are much better known than their tropical counterparts, and large-sized vertebrates are much better known than small-sized invertebrates. Even in the best known groups in the best known areas, however, we only have estimates of numbers of species, not actual numbers. And the huge diversity of microbes, for which the species concept itself is inappropriate, is only beginning to be uncovered using new molecular tools. There are two main reasons why species are difficult to count. The first reason is that there are simply too many organisms on Earth to count them all, even if we had the appropriate technical means to do so. In many cases, technical limitations would not even allow considering this option. Species diversity must generally be estimated from incomplete samples obtained using imperfect techniques, which inevitably introduces errors and uncertainties in its estimates. Typically, the smaller the sample, the larger the uncertainty. The second reason is that species have widely different relative abundances. Most communities typically contain a few abundant species and many rare species. Since abundant species are much more likely to be sampled than rare species, they are also much more likely to be recorded. Rare species are particularly difficult to detect, even with large sample sizes. Thus, species evenness — the degree to which species are similar in their abundances — affects the accuracy of estimates of species richness — the number of species — jointly with sampling effort. This explains why these two components of species diversity have been distinguished but at the same time why they need to be considered simultaneously. There are basically two ways to resolve this difficulty: either by keeping the focus on species richness and in some way getting around the problem of sample size, or by devising some integrative measures of species diversity that combine its richness and evenness components meaningfully.

Estimating Species Richness Using Species Accumulation and Rarefaction Curves The first path is based on species accumulation and rarefaction curves (Gotelli and Colwell 2001) (Fig. 4). Species accumulation curves show the number of species obtained by successively censusing either individual organisms (indi-

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WHAT IS BIODIVERSITY?

Fig. 4. Species accumulation and rarefaction curves. Species accumulation curves show the number of species obtained by successively censusing either individual organisms (individual-based accumulation curves) or samples (sample-based accumulation curves). Smoothed species rarefaction curves represent the statistical expectation of the corresponding accumulation curves. Credit: Rob Colwell, after Gotelli and Colwell (2001)

vidual-based accumulation curves) or samples (sample-based accumulation curves). The sample-based curve lies nearly always below the individual-based curve because samples usually contain several individuals that are nearby in space or consecutive in time. Since individuals from the same species are often aggregated in space and time, fewer species are represented in a sample on average than by an equal number of individuals censused independently in the same habitat. Species rarefaction curves are smoothed versions of species accumulation curves, obtained by repeatedly re-sampling the final pool of either individuals (individual-based rarefaction curves) or samples (sample-based rarefaction curves) at random at each sample size. A species rarefaction curve, whether based on individuals or on samples, is simply the statistical expectation of the corresponding accumulation curve. Species rarefaction curves are typically used to compare species richness among communities that have not been fully inventoried or have been inventoried with unequal effort. They are then interpolated backwards to smaller sample sizes (whence the term rarefaction) such that all communities are compared

SPECIES DIVERSITY BASED ON ABUNDANCE DISTRIBUTIONS

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based on the same sample size. In these applications, the objective is not to obtain an absolute estimate of species richness, but rather to obtain fair comparative values across different communities. Species accumulation and rarefaction curves can also be extrapolated forwards to estimate their unknown asymptote, which gives the true value of species richness in a community. The simplest non-parametric estimator of species richness, Sest, augments the number of species observed, Sobs, by a term that depends only upon the observed number of singletons (species each represented by a single individual), a, and doubletons (species each represented by two individuals), b (Chao 1984; Colwell 2009): a2 (1) 2b This estimator provides a fairly good estimate of true species richness as soon as sample size is large enough to detect the majority of species in a community. Unfortunately, it does not provide reliable estimates of species richness in hyperdiverse communities with large numbers of rare species, nor can it be applied to heterogeneous systems where there are large differences in species composition among sites. Therefore it does not help to resolve the hotly debated issue of the total number of species on Earth. Sest = Sobs +

Measuring Species Diversity Based on Species Abundance Distributions The second way to deal with the interaction between species richness and species evenness consists in incorporating the two components in more integrative measures of species diversity. Again, there are two alternative approaches to doing this. The first approach historically is based on species abundance distributions. Species abundance distributions describe patterns of species relative abundances within communities. There are two traditional ways to plot empirical data on species abundances. Rank-dominance curves plot the abundance of each species, usually on a logarithmic scale, against its abundance rank (Whittaker 1965) (Fig. 5). The slope of these curves is an inverse measure of species evenness: steep curves are indicative of strong dominance and low evenness (highly unequal abundances), whereas flat curves are indicative of high evenness (highly equal abundances). Abundance frequency histograms plot the number of species against abundance categories, usually in powers of two, thus also yielding a logarithmic scale of abundances (Preston 1948) (Fig. 6). Motomura (1932) first suggested a simple model to account for the fact that species in a community often have strongly unequal relative abundances. His

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Fig. 5. Rank-dominance curves for three well-known species abundance distributions: (A) the geometric series, fitted to empirical data from a vascular plant community in a subalpine fir forest in the Great Smoky Mountains, Tennessee; (B) the lognormal distribution, fitted to empirical data from a vascular plant community in a deciduous forest, also in the Great Smoky Mountains; and (C) the broken-stick distribution, fitted to empirical data from a bird community in a deciduous forest, West Virginia. Lines are theoretical distributions; points are empirical data. These rank-dominance curves plot the relative abundance of each species (measured as a percentage of total abundance) on a logarithmic scale, against its abundance rank. After Whittaker (1970)

model implicitly assumes a competitive hierarchy in which the single most abundant species appropriates some fraction of the available habitat or resources, the second most abundant species appropriates the same fraction of what is left, and so on ad infinitum. The ratio between the abundances of two consecutive species in the competitive hierarchy is then a constant, and the resulting species abundance distribution is a geometric series. The geometric series yields a straight rank-dominance curve when abundance is plotted on a logarithmic scale. It fits empirical data from very species-poor communities, often in extreme environments (Fig. 5). Many other statistical distributions have since been proposed to describe patterns of species relative abundance. The three best known such distributions are probably the logarithmic series, the lognormal distribution, and the brokenstick distribution. The logarithmic series was proposed by Fisher et al. (1943) to describe the distribution of species relative abundances in insect communi-

SPECIES DIVERSITY BASED ON ABUNDANCE DISTRIBUTIONS

23

Fig. 6. Abundance frequency histogram showing the lognormal distribution (curve), fitted to empirical data (histogram) from Sonoran desert plant communities on lower mountain slopes of the Santa Catalina Mountains, Arizona. The histogram plots the number of species against their coverage values (in %) on a logarithmic scale. Based on data from Whittaker (1965)

ties. It predicts a large number of rare species, as is typically observed in insect samples, which yields a rank-dominance curve that becomes flatter for rarer species (with a larger rank number). A significant advantage of the logarithmic series is that species richness, S, is related to the total number of individuals in a sample, N, by a simple relation that has a single free parameter, α: N S     =     α ln ⎛ 1 + ⎞ ⎝ α⎠

(2)

This parameter is independent of sample size, and thus provides a natural measure of species diversity, known as Fisher’s α diversity index. Note that this index makes the distinction between species richness and species evenness superfluous: the two components of species diversity are inextricably linked through the above relation and can be measured simultaneously by the single parameter α. The logarithmic series is more flexible than the geometric series and can often accommodate empirical data fitted by the geometric series. Unfortunately, not all species abundance patterns in nature can be described by the logarithmic series, and hence Fisher’s α diversity index does not provide a universal measure of diversity. In particular, large-scale patterns of species abundances in plants often obey a lognormal distribution, characterised by a bell-shaped curve when the number

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of species is plotted against the logarithm of species abundance (Preston 1948, 1962) (Fig. 6). Thus, in contrast to the logarithmic series, which predicts more and more species as they become rarer, the lognormal distribution predicts a maximum number of species at intermediate abundances and a declining number of rare species. This pattern yields a sigmoid rank-abundance curve, the inflexion point of which corresponds to the peak in the abundance frequency histogram (Fig. 5). Lastly, some bird communities appear to obey a distribution known as the broken-stick distribution. The broken-stick distribution owes its name to the rule used to generate it. As in the geometric series, the relative abundance of each species is assumed to be proportional to the fraction of habitat or resources available to it, but here resources are allocated at random, like a stick that would be broken simultaneously and randomly in as many segments as there are species in the community (MacArthur 1957). Since no intrinsic dominance is assumed in this model, the broken-stick distribution predicts highly even species abundance patterns (Fig. 5). There is a large literature on species abundance distributions. Some, following the lead of MacArthur, have interpreted them as indicative of different rules of resource allocation within communities (Tokeshi 1999). But others have shown that they have purely statistical interpretations (Caswell 1976; Cohen 1968; Hubbell 2001; May 1975). Since it is often difficult to distinguish between the various distributions (Engen 1978) and statistical explanations are more parsimonious because they do not invoke specific mechanisms, it seems reasonable to stick to the statistical interpretation of species abundance distributions unless there is strong evidence for specific resource allocation rules operating in some communities. The advantage of species abundance distributions in studies of species diversity is that they can be fitted to empirical data and yield one or a few parameters that have a clear interpretation and whose value can be compared easily among communities that obey the same distribution. Their corresponding disadvantage is that they are not universal, and thus it is almost impossible to compare communities that obey different distributions.

Measuring Species Diversity Using Non-Parametric Diversity Indices Given the limitations of species abundance distributions, more flexible integrative measures of species diversity were developed in ecology, and rapidly enjoyed considerable popularity. These measures are called diversity indices, just like Fisher’s α. But unlike Fisher’s α, which is the parameter of a specific

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25

statistical distribution, these indices are non-parametric, i.e., they can be used irrespective of the species abundance distribution. Non-parametric diversity indices are mathematical functions of species relative abundances that combine the effects of species richness and species evenness in a single measure. Although there are many others, the most commonly used non-parametric diversity indices in ecology are Shannon’s and Simpson’s. Shannon’s diversity index (Shannon and Weaver 1949) was actually devised as a measure of entropy or uncertainty in information theory, and has the form: S

H = – ∑ pi ln pi i =1

(3)

where pi is the relative abundance of species i, i.e., its proportion in the total number of individuals (or any other measure of quantitative importance, such as biomass and plant cover) in the community. Its lower bound is zero when the community is strongly dominated by a single species (p1 ≈ 1, pi ≈ 0 for all i > 1, and species are ranked in decreasing order of abundance), and its upper bound for a given species richness S is lnS when all species have equal relative abundances (pi = 1/S for all i). This shows clearly that Shannon’s diversity index incorporates both species evenness (which generates variations between zero and lnS for a given value of species richness) and species richness (which generates variations in the upper bound, lnS). Simpson (1949) proposed a concentration index, λ, defined as: λ =

S

∑ pi2 i =1

(4)

This index has an upper limit of one when the community is strongly dominated by a single species (p1 ≈ 1, pi ≈ 0 for all i > 1), and a lower bound of 1/S when all species have equal relative abundances (pi = 1/S for all i). Thus, Simpson’s concentration index measures the inverse of diversity. Two corresponding diversity indices are commonly used: Δ1 = 1–λ

(5)

Δ2 = 1/λ

(6)

Δ1 has a simple interpretation: it measures the probability that two individuals drawn at random in the community belong to different species. Δ2 can be interpreted as a measure of the “effective number of species” in the community, i.e., the number of species that would be found in a community that has the same value of Δ2 but equal species abundances. Just as Shannon’s, Simpson’s diversity indices incorporate both species evenness and species richness. The two types of indices, however, do not

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always rank communities in the same order. Simpson’s indices are more sensitive to abundant species and less sensitive to rare species than Shannon’s index. The latter, in turn, is more sensitive to abundant species and less sensitive to rare species than species richness, S, which gives equal weights to all species. At the other extreme lies Berger and Parker’s (1970) dominance index, which is simply p1, the relative abundance of the single most abundant species. This index depends only on the most abundant species, and is insensitive to rare species. Thus, these various diversity indices appear to be variations on the same theme. As a matter of fact, their common theme can be formalised mathematically using Rényi’s (1961) generalised measure of entropy of order a: Ha =

S 1 ⎞ ⎛ ln ⎜ ∑ pia ⎟ 1 − a ⎝ i =1 ⎠

(7)

or, equivalently, Hill’s (1973) diversity index of order a: 1

Sa = e H a

S

⎞ 1− a ⎛ = ⎜ ∑ pia ⎟ ⎝ i =1 ⎠

(8)

Species richness, Shannon’s diversity index, Simpson’s concentration index and Berger and Parker’s dominance index turn out to be specific instances of Rényi’s entropy or Hill’s diversity index of different orders (Table 1). Thus, Rényi’s generalised measure of entropy provides a theoretical framework that unifies the most common measures of diversity. The order a in Rényi’s entropy governs the balance between the weights given to abundant species and rare species. When a = 0, relative abundances are raised to the power 0 in the above equations, and hence their differences are erased. Hill’s diversity index then reduces to species richness. In contrast, when a is very large, relative abundances are raised to a very high power in the above equations. The single most abundant species then greatly outweighs all other species, and its relative abundance entirely determines the value of Rényi’s entropy. Table 1. Relationships between the most common measures of species diversity (S, species richness; H, Shannon’s diversity index; λ, Simpson’s concentration index; and p1, Berger and Parker’s dominance index) and Rényi’s entropy or Hill’s diversity index Order a 0 1 2 ∞

Rényi’s entropy Ha

Hill’s diversity index Sa

ln S H – ln λ – ln p1

S eH Δ2 = 1 / λ 1 / p1

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27

This unifying framework is not only intellectually satisfying; it also provides an effective way to overcome one of the main limitations of diversity indices, i.e., the fact that no single number can capture the full multidimensional nature of the diversity concept. Rényi’s entropy spans a whole family of diversity indices that give different weights to species abundances. Instead of applying a single diversity index, which to some degree is an arbitrary choice, one may apply the whole family of indices and produce a complete diversity profile by plotting the value of Rényi’s entropy or Hill’s diversity index against its order over a whole range of orders for each community (Patil and Taillie 1982; Tothmérész 1995) (Fig. 7). If the whole profile yields a consistent ordering of communities, we can be confident that the differences in species diver-

Fig. 7. Diversity profiles showing Rényi entropy as a function of its order parameter in three forest ground-beetle communities sampled using pitfall traps from 1978 to 1981 in Lembeek, Belgium. Arrows show how Rényi entropy unifies the most common measures of species diversity: S, species richness; H, Shannon’s diversity index; λ, Simpson’s concentration index; and p1, Berger and Parker’s dominance index. The whole diversity profile of the community from the successional forest (n) lies above the other two, which clearly establishes that this community is the most diverse. The diversity profile of the community from the beechwood (d) lies above that of the community from the pinewood (h) for small orders, and below it for large orders, yielding mixed conclusions on which community is most diverse. Species richness is highest in the beechwood community, but species evenness is highest in the pinewood community. The x-axis is interrupted between 2.5 and ∞. Based on data from Loreau (1984)

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sity thus revealed are robust. By contrast, if indices of different orders yield different orderings of communities (such that their diversity profiles intersect), the differences in species diversity revealed by the various indices should be interpreted with caution. Rényi’s generalised measure of entropy also allows unification of measures of diversity, richness and evenness. Although a large number of species evenness indices have been proposed, most of them are incommensurate with measures of species diversity and species richness. Since species richness and species evenness are two components of species diversity, one would ideally like to have measures that account for these relationships between the three concepts. Rényi’s generalised measure of entropy allows precisely this. Since Hill’s diversity index of order zero is equal to species richness, it is straightforward to define measures of species evenness, Fa and Ea, that yield an explicit partition of species diversity into its two components (Buzas and Hayek 1996; Kindt et al. 2006): Ha = lnS + Fa

(9)

Sa = SEa

(10)

or, equivalently:

where the two measures of evenness are related by Fa = ln Ea. Thus, Rényi’s entropy of order a is simply the sum of species richness and species evenness of the same order, while Hill’s corresponding diversity index is the product of species richness and species evenness. The evenness index Ea is comprised between zero and one, and has a simple interpretation. Hill’s diversity index of order a, Sa, measures the effective number of species in the community under the transformation of species relative abundances prescribed by the index. The corresponding evenness index, Ea, is the fraction of species richness that represents the effective number of species under this transformation.

Species Diversity in Space: α, β, and γ Diversity One of the most striking features of biodiversity at all levels is its strong spatial organisation. Different species are present in different localities, and these differences in species composition typically increase as the spatial scale increases from small plots to entire continents, resulting in a steady increase in species diversity. At the global scale, this spatial differentiation yields a set of distinct biomes, i.e., types of ecosystems characteristic for different climatic conditions. Several factors contribute to this increased spatial differentiation and

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29

amount of biodiversity with area. One factor is statistical in nature, and should ideally be removed as far as possible: small sampling units may be simply too small to reveal all the species characteristic of the habitat from which these units are taken. Other factors are ecological, and include spatial heterogeneity of the abiotic environment (climate, soil, nutrients), dispersal limitation (species are unable to reach geographically distant habitats), and distinct evolutionary histories on different islands or continents. There are different ways to study and measure the spatial organisation of species diversity. One common approach, proposed by Whittaker (1972), is to define three nested spatial levels of species diversity, called alpha, beta and gamma diversity. Alpha diversity, or local diversity, is the diversity of a local community, i.e., of an assemblage of coexisting species in a relatively homogeneous habitat. Note that alpha diversity is not equivalent to Fisher’s α diversity index: the latter is just one possible measure of the former. Beta diversity is the increase in diversity arising from differences in species composition among different localities as one moves from local to regional spatial scales. Lastly, gamma diversity, or regional diversity, is the combined diversity of all habitats in a region. Thus, under these definitions, gamma diversity is simply a combination of alpha diversity and beta diversity. Alpha and gamma diversity are typically measured using the same metric — for instance, by the number of species or any of the other metrics presented in the previous sections. Whittaker initially proposed to measure beta diversity as the ratio between gamma and alpha diversity, which implicitly assumes a multiplicative partition of gamma diversity into its alpha and beta components. An alternative, powerful approach to integrate the three spatial levels of diversity is to use an additive partition of gamma diversity into its alpha and beta components (Lande 1996; Loreau 2000a). Take, for instance, Rényi’s entropy of order a, Ha, as a measure of diversity. This metric can be applied to the various local communities in the region, yielding an estimate of alpha diversity Haαj in each locality j. The same metric can be applied to the region as a whole, yielding an estimate of gamma diversity, Haγ . It is then possible to define a corresponding measure of beta diversity, Haβ , such that: 33 Haγ = H aα + Haβ

(11)

33 where H aα is the average alpha diversity across the region. The advantages of this additive partition are many. First, alpha, beta and gamma have the same measure and can be compared quantitatively. Second, this additive partition can be applied consistently at multiple nested spatial scales, from that of single sample to different habitat types to landscapes to entire regions (Loreau 2000a; Fournier and Loreau 2001; Crist et al. 2003). Gamma diversity at one spatial scale then becomes alpha diversity at the next

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higher spatial scale. Since the separation between the local and regional scales is to some extent arbitrary, the ability to deal with multiple scales is a distinct plus. Third, the above additive partition is based on the same principle as an analysis of variance — it is even formally equivalent to an analysis of variance when the diversity index Δ1 based on Simpson’s concentration index is used as the measure of diversity (Lande 1996). Note that an equivalent multiplicative partition can easily be obtained when Rényi’s entropy is used to measure diversity as in the above equation. Rényi’s entropy essentially measures diversity on a logarithmic scale. The equivalent measure of diversity on an arithmetic scale is Hill’s diversity index. Therefore the above equation can be rewritten on an arithmetic scale using Hill’s diversity index: (12) Saγ = 〈Saα〈 Saβ where Saγ , Saβ and Saα j are the exponentials of Haγ , Haβ and Haα j, respectively, and 〈Saα〈 is the geometric mean of Hill’s alpha diversity across the region. This equation shows that a multiplicative partition simply requires working with geometric means instead of the more traditional arithmetic means. Species richness, i.e., Hill’s diversity index of order zero, can also be used in an additive partition as above, but Hill’s diversity indices of higher orders generally cannot because they are not strictly concave, which means concretely that gamma diversity can be lower than the arithmetic average alpha diversity, thus resulting in a negative estimate of beta diversity (Lande 1996). The above equations show that a complete unification of measures of species diversity can be achieved using Rényi’s entropy or Hill’s diversity index. The two partitions of species diversity into its richness and evenness components and of gamma diversity into its alpha and beta components can be combined to yield a general, multi-component, multi-scale formal framework to measure species diversity (Kindt and Loreau, unpublished manuscript). There are two such frameworks: one, based on Rényi’s entropy, is additive, while the other, based on Hill’s diversity index, is multiplicative. Both frameworks allow complete diversity profiles to be studied for all the components of species diversity. Despite their power and generality, these frameworks may not always provide the most suitable description of spatial patterns of species diversity. Under some circumstances, one may be interested either in finer details or, on the contrary, in a broader picture. Pairwise similarity indices are measures of beta diversity that are usually standardised between zero and one and that provide detailed comparisons between the species compositions and species relative abundances of two local communities. There are many such indices, which I shall not discuss here. Equivalent aggregate measures of similarity between

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communities at the regional scale are provided by ratios between average alpha 33 diversity and gamma diversity, i.e., H aα /Haγ or 〈Saα 〈/Saγ , in the above frameworks based on Réyni’s entropy or Hill’s diversity index.

Species Diversity in Space: Species–Area Relationships At the other extreme, species–area relationships provide a global view of spatial patterns of species richness across a wide range of spatial scales. These relationships plot species richness against area sampled, generally on logarithmic scales. The relationship between the number of species, S, found in an area and the surface of this area, A, is approximately a power law of the form: S = cAz

(13)

which yields a straight line on a log–log graph. The slope of this line is the coefficient z of the power law. Again, there are several variants, but the main dichotomy separates nested sampling schemes in which small areas are subsets of large areas from non-nested ones in which samples of different sizes are independent from each other. Species–area relationships are continuous, and hence they do not distinguish between alpha diversity and gamma diversity, but their slope at each point, which measures the increase in species richness with area, provides a straightforward continuous measure of beta diversity. This slope is approximately constant at fairly large scales within biogeographical provinces, i.e., within regions that have distinct evolutionary histories and species pools. Empirical data and theoretical arguments have long suggested that it should be approximately equal to 0.25 (Preston 1962; May 1975). This property has been commonly used to predict the effects of habitat destruction and fragmentation on species diversity (Chapter 3). In actual fact, the slope of species–area relationships on a log-log scale varies widely depending on the spatial scale considered and the ecological and evolutionary processes that determine beta diversity at these scales. Empirical data from a nested sampling scheme of vascular plant species richness in Great Britain across a wide range of spatial scales, from small 10 × 10 cm quadrats to the whole of Britain, provide clear evidence for such a scale dependence (Crawley and Harral 2001). At small scales (less than 1000 m2), the species–area relationship has a shallow slope z = 0.16, indicating low beta diversity; at intermediate scales (from 1000 m2 to 16 km2) its slope is much steeper, with z = 0.45–0.50, indicating high beta diversity; and at the largest scales (25 km2 upward) its slope is shallower again, with z = 0.24 (Fig. 8). Ignoring this significant scale dependence, however, would yield an average slope of z = 0.27–0.30, which is statistically

32

WHAT IS BIODIVERSITY?

Fig. 8. Species–area relationships for vascular plants in Great Britain. (A) Logarithm of species richness in Silwood Park at scales from 0.01 m2 to 110 ha (mean ± SD). The smaller quadrats (up to 1000 m2) were located at random within habitats, whereas data from the contiguous 1-ha plots were aggregated to obtain species richness at larger scales. Piecewise linear regression highlights the strength of the scale dependence in the slope of the species–area relationship. The small-scale data (less than 1000 m2) have a much shallower slope than the larger-scale data (1000 m2 upward). (B) Logarithm of species richness in East Berkshire at scales from 1 km2 to 465 km2, in Berkshire as a whole (1876 km2), and in Great Britain (2.3 × 105 km2) (mean ± SD). For East Berkshire the data were accumulated from contiguous samples of 1 km2 to larger scales. Piecewise linear regression again highlights the strength of the scale dependence in the slope of the species–area relationship. Data from intermediate scales (1 to 16 km2) have a much steeper slope than data from the largest scales (25 km2 upward). Ignoring the significant scale dependence in the slope of the species–area relationship would yield an average slope of 0.302 in (A) and 0.267 in (B), none of which is significantly different from z = 0.25. Based on data from Crawley and Harral (2001)

BIODIVERSITY BEYOND SPECIES

33

indistinguishable from z = 0.25. Ecological interactions, such as competition and seed dispersal, presumably explain the low beta diversity at small scales; habitat heterogeneity created by geology, topography, hydrology, and management very likely explains the high beta diversity at intermediate scales; and evolutionary and historical constraints likely explain the relatively low beta diversity at the largest scales within Britain. At yet larger scales, when different biogeographical provinces are compared, species–area relationships are typically very steep (z ≈ 0.9–1), because the biodiversity of different biogeographical provinces is shaped by independent evolutionary processes (Rosenzweig 1999). Islands also usually have steeper slopes than continents because dispersal is limited and evolutionary processes play a more important role on islands. The more isolated they are, the steeper their species–area relationship (Rosenzweig 1999). Thus, although species–area relationships have so far provided one of the few available ways to predict the effects of habitat destruction on species diversity (Chapter 3), their predictive ability is strongly limited by the scale dependence, not only of their slope, but also of the ecological and evolutionary processes that govern it. These processes operate at very different time scales, which should be taken into account when making future predictions.

Biodiversity beyond Species I have focused on species diversity in the previous section because its measurement has been studied extensively in ecology. But the same approaches can be used to measure diversity at other levels of organisation, and even the disparity component of diversity. Although quantification of genetic diversity has received comparatively less attention than that of species diversity, it faces similar problems, with similar solutions. Genetic diversity is often assessed based on a simple count of the number of genetic variants in a population, and expressed in such forms as the proportion of polymorphic loci, the richness of allelic variants, and the average number of alleles per locus. These metrics are roughly equivalent to measures of species richness in ecology. The commonest measures of genetic diversity that incorporate the relative frequencies of genetic variants are identical to those traditionally used to measure species diversity in ecology. Diversity indices from the Rényi family have also been used in population genetics to partition genetic diversity within and between populations. For instance, Lewontin (1972) proposed an additive partition of genetic diversity in human populations based on Shannon’s diversity index which is essentially identical to the additive partition of gamma diversity into its alpha and beta components

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presented in the previous section. Nei (1973) proposed a similar approach based on Simpson’s diversity index. Other common spatial measures of genetic diversity include the standardised genetic variance among populations, FST, which measures the proportional reduction in the heterozygosity of a metapopulation due to differentiation among its components populations (Pannell and Charlesworth 2000). This is essentially a relative measure of beta genetic diversity. Ecological diversity above the species level has been studied surprisingly little quantitatively. Yet ecological diversity is potentially measurable using the unifying theoretical framework that has been developed to measure species diversity and genetic diversity. Tansky (1974) proposed a measure of ecological diversity that has the form of Shannon’s diversity index but in which the proportions used to compute the index are trophic coefficients describing the position of the various species in the food web, instead of mere relative abundances. Such attempts, however, are remarkably rare. As mentioned earlier, ecology has mostly focused on horizontal ecological diversity, while vertical ecological diversity has received much less attention, despite its obvious importance in the structure and functioning of ecosystems. Food chain length could provide a straightforward analogue of horizontal species richness in the vertical dimension (Duffy et al. 2007). Although the number of apparent trophic levels varies relatively little among ecosystems (typically between 3 and 5; Pimm 1982), more realistic continuous measures of trophic position (Vander Zanden and Rasmussen 1996) and inclusion of the many belowground organisms and parasites, ignored in traditional food web studies (Wardle 2002; Lafferty et al. 2008), are likely to significantly increase the range of variation of food chain length. More integrative metrics that include the horizontal and vertical dimensions of ecological diversity, such as the one suggested by Tansky, would be particularly valuable. Several approaches have been developed recently to measure the disparity component of organismal diversity. Traditional measures of species diversity treat all species as being equal in all respects except their abundance. But clearly all species are not equal. In particular, some species may be the last remnants of a very rare taxonomic group, while others may belong to a very common taxonomic group in which many species strongly resemble each other. A number of authors have devised measures of phylogenetic diversity or evolutionary distinctness to capture these differences in species evolutionary histories (Crozier 1997). These metrics use information contained in the phylogenetic tree that links the various species considered together, such as the total branch length of the tree (Faith 1992) or the average branch length between two randomly chosen organisms in the tree (Clarke and Warwick 1998). Some of the most recent metrics that incorporate information on species relative abun-

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dances resort to Shannon’s diversity index to measure the diversity of species evolutionary distinctness (Cadotte et al. 2009). Generalising these metrics to other members of the Rényi family would be straightforward. Similar metrics have also been proposed to measure phenotypic or functional diversity based on functional dendrograms (Petchey and Gaston 2002). A simple alternative is to use classic statistical moments of trait distributions, such as phenotypic variance (Norberg et al. 2001; Savage et al. 2007), in a community. Trait distributions, however, would again be easily amenable to a description using traditional diversity indices such as those from the Rényi family. Lastly, biodiversity has become an important societal issue. Therefore, while rigorous methods are needed to quantify biodiversity in the biological sciences, simpler, coarser and more integrative approaches are also needed to assess biodiversity for management and policy-making purposes. The index of biotic integrity, a now popular biological indicator of the quality of streams in the USA and elsewhere (Karr 1981), is an example of such approaches. The challenge today is to devise appropriate aggregate indices that are sufficiently reliable, representative, and sensitive to biodiversity changes to be used in national and international policy decisions, and that would play a role akin to price indices in economics. This is a social as much as a scientific challenge since no single number will ever capture the multidimensional nature of biodiversity and value issues will inevitably come into play. But the challenge should be taken up if biodiversity is to be given serious consideration in policy and economic decisions.

3 BIODIVERSITY IN CRISIS The Balance between Origination and Extinction Species diversity at any place and time is the result of a complex set of ecological and historical factors, including the current local biotic and abiotic environment, the current regional biotic and abiotic environment, the ability of organisms to disperse, and the past history of species assemblages at both local and regional scales. At large spatial and temporal scales, however, species diversity is ultimately maintained by a dynamical balance between origination of new species and extinction of existing species, two long-term evolutionary phenomena under the combined influence of genetic and ecological processes. The rate at which the number of species, S(t), changes through time t, is simply equal to the total rate of origination of new species, O(t), minus the total rate of extinction of existing species, E(t): dS (t ) = O (t ) − E (t ) dt

(14)

Similar balance equations hold for other taxonomic units, be they genera, families, or higher units. We still have limited understanding of the factors that determine the rates of origination and extinction over long time scales. The modern theory of evolution describes the mechanisms that underlie evolutionary dynamics, mostly at relatively short time scales. Random mutations and recombination of the genetic material during sexual reproduction or horizontal gene transfer create new genetic variability. Developmental and environmental constraints interact with the information coded in the genetic material to map genetic diversity onto phenotypic diversity. Natural selection then sifts phenotypic and genetic diversity based on the relative fitness of organisms to the prevailing environmental conditions. Although this theory provides a powerful framework to understand the mechanisms that operate in evolution, it says nothing about the long-term direction of evolution and the complex patterns it generates. These patterns are to a large extent governed by ecological processes, i.e., by the interactions between organisms and their changing biotic and abiotic environment. Ecology has made tremendous progress over the last century to disentangle the various factors that affect the interactions between organisms and their biotic and abiotic environment, again mainly on relatively short time scales. But it is still far from a comprehensive theory that integrates all levels of biological organisation at all spatial and temporal scales. Recent community and ecosystem evolution models suggest that ecological patterns are emergent

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properties of the combined ecological and evolutionary dynamics of interacting organisms, which we are but starting to explore (Loeuille and Loreau 2005, 2010; McKane 2004). It would be logical to expect the total rates of origination and extinction in the above equation to be functions of the number of species itself. The more species there are, the higher the chances that these species will give birth to new species. Total extinction rates should be even more sensitive to the number of species. If species extinctions were independent stochastic events, total extinction rates should be simply proportional to the number of species i.e., E(t) = e(t) S(t), where e(t) is an average extinction rate per species (or species extinction rate in short), and is independent of the number of species. But the extinction of a species is a complex process that culminates in the disappearance of the last individual of its last population. Therefore it is ultimately determined by chance demographic events in small populations, which are termed collectively demographic stochasticity. As a result, the probability of extinction of a population is a decreasing function of the number of individuals it contains (Lande et al. 2003). Since more species means fewer individuals per species on average when resources are kept constant, one should expect the species extinction rate, e(t), itself to be an increasing function of the number of species. Total extinction rates should then increase faster than total origination rates as species diversity increases, leading eventually to regulation of species diversity around some saturating level. There has been much debate and speculation on possible saturation of species diversity at various spatial and temporal scales in both ecology and paleontology, but the evidence for this hypothesis is still scant. One of the main problems with this hypothesis is that the dependence of species extinction rates on population size is strongly nonlinear, and mostly occurs when population size is very small. It is unclear whether the number of species is currently high enough to significantly affect species extinction rates at the global scale. Another major problem is that the physical environment undergoes considerable fluctuations on time scales smaller than those characteristic for the relatively slow origination processes, such that saturation of taxonomic diversity may never be achieved. Physical factors such as climate, available habitat area, and geographical barriers to dispersal are likely to be much more important in determining large-scale patterns of species diversity. Since neither current evolutionary theory nor current ecological theory is able to make more detailed predictions regarding the functional form and absolute value of long-term origination and extinction rates, most of the information currently available on the long-term dynamics of biodiversity in the past comes from the fossil record. Molecular approaches are also increasingly used to help reconstruct the past history of biodiversity. If mutations occur at random

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and natural selection is weak, the divergence of molecular sequences of nucleotides provides a “molecular clock” that allows estimation of the timing of the evolutionary divergence of different lineages. The fossil record provides the approximate times of the first appearance of a taxon and of its ultimate extinction, and hence, indirectly, an estimate of its lifespan. It also allows taxonomic diversity to be estimated at any point in time by adding the taxa known to be present at that time. Estimates of taxonomic diversity are more reliable for higher taxonomic units (e.g., families) because lower taxonomic units (e.g., species) are more likely to be missing in the documented fossil record. These estimates show that taxonomic diversity has increased through geological times, but with a few abrupt drops in the marine fossil record (Fig. 9). These abrupt drops correspond to episodes of mass extinctions during which a large proportion of species became extinct (Raup and Sepkoski 1982; Sepkoski 1993; Benton 1995). Five such episodes punctuate the history of life on Earth, each of which marks the end of a geological era (Fig. 9). The most cataclysmic of these episodes occurred at the end of the

Fig. 9. Changes in the number of families of marine vertebrates and invertebrates in the fossil record through the Earth’s evolutionary history. Five mass extinctions, indicated by numerals, are recognisable by abrupt drops in family diversity. Their relative magnitudes are given in parentheses in the box. After Raup and Sepkoski (1982)

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Fig. 10. Frequency distribution of species extinction rates per million years in the fossil record. Mass extinctions occupy the right-hand tail of the distribution. The mean species extinction rate, 25% per million years, is the reciprocal of the mean species duration, 4 million years. After Raup (1994)

Permian, when about 52% of all marine families and 96% of all marine species disappeared. The causes of mass extinctions are still debated, but the most commonly accepted hypotheses invoke major physical disturbances, including climate change, sea-level changes, extraterrestrial impacts, and changes in volcanic activity. Mass extinctions are fascinating because they illustrate the fragility of life on Earth. But they are but an extreme manifestation of the permanent phenomenon of species extinction. The frequency distribution of species extinction rates in the marine fossil record shows no sign of discontinuity (Fig. 10): mass extinctions are but the tail of a continuous distribution (Raup 1994). The average species extinction rate during the past 600 million years is about 25% per million years. The reciprocal of this rate, 4 million years, gives an estimate of the average lifespan of species in the geological past of the Earth. The average species lifespan, however, varies substantially among taxonomic groups, from about 1 million years in primates to about 25 million years in corals (McKinney 1997). Since taxonomic diversity has increased overall during the history of life, extinctions have been more than compensated, on average, by originations. The increase in taxonomic diversity, however, has occurred over such long periods of time that origination rates are virtually equal to extinction rates, on average (Benton 1995). But origination rates are just as variable in time as extinction

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rates. Mass extinctions are typically followed by longer periods of high origination rates, during which surviving taxa diversify relatively rapidly, probably as result of the new ecological opportunities offered by the disappearance of competing taxa. Rapid diversification on geological time scales, however, is still exceedingly slow on a human scale. Recoveries of taxonomic diversity from mass extinctions are predicted to take roughly 10 to 40 million years (Alroy 2008). In spite of the inevitably large uncertainty associated with estimates based on fossil data, the above figures provide invaluable orders of magnitude of the characteristic time scales of the evolutionary processes that govern the dynamics of global taxonomic diversity. These time scales are measured in millions of years, and are roughly five to six orders of magnitude slower than the characteristic time scales of the social and economic processes in current human societies.

Heading for a Sixth Mass Extinction The rise of the human species is creating a completely novel situation in the history of life on Earth. For the first time a single species has developed the ability to colonise, use and transform virtually the whole biosphere, and, as a result, is threatening global biodiversity just as surely as would a massive extraterrestrial body colliding with the Earth. Humans have had considerable impacts on biodiversity since prehistoric times. With perhaps the exception of Africa, colonisation of new continents and islands by humans seems to have been accompanied by mass extinctions of large-sized mammals and birds, probably due to hunting, deforestation, fire, and other changes in land use (Barnosky 2008). The colonisation of the numerous Oceanian islands alone may have driven thousands of bird species, many endemic, to extinction between 30000 and 1000 years BP (Steadman 1995). Humans seem to have also been involved in the mass extinction of all Australian land mammals, reptiles and birds weighing more than 100 kg in the late Quaternary (Miller et al. 1999; Roberts et al. 2001). Human impacts on biodiversity, however, have not been systematically negative. Moderate disturbances tend to enhance species diversity because they reduce the abundance of competitive dominants (Connell 1978; Huston 1979). Introduction of exotic plant species may also increase local plant diversity (Sax and Gaines 2008). Therefore moderate disturbances and species introductions generated by humans have probably increased the diversity of at least some taxonomic groups in some regions. For instance, the high plant diversity of the Mediterranean basin is thought to be partly attributable to the presence of humans, who have gradually modified landscapes for thousands of years, in particular through livestock grazing and fire,

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thereby enhancing the genetic diversity and species diversity of both domesticated and wild plants (Blondel 2006). But, in spite of such localised increases, the global trend has been clearly towards massive biodiversity loss. Unfortunately, there are few reliable quantitative data on species extinctions during human prehistory and much of human history. Species extinctions have started to be documented since about 1600. In its 2008 Red List, the International Union for Conservation of Nature (IUCN) reported 717 documented extinctions of animal species and 87 documented extinctions of plant species worldwide, many of which have taken place on islands during the last century. Islands have been particularly prone to species extinctions because they harbour endemic faunas that are poorly adapted to hunting by humans and to introduced predators and parasites (Sax and Gaines 2008). These figures unquestionably underestimate the number of real extinctions that have taken place during the past 400 years, because most extinctions go unnoticed. Many taxonomic groups are too poorly known, and documenting the extinction of a species with certainty is exceedingly difficult. Nevertheless, even these underestimates speak volumes. Take the bestdocumented taxonomic groups, i.e., mammals and birds. There have been 76 and 134 documented species extinctions in mammals and birds, respectively. Since there is an estimated 5488 described species of mammals and 9990 described species of birds, documented extinctions in these two groups correspond to 1.38 and 1.34%, respectively, of extant species. These percentages look like small numbers at first sight, but appearances are deceptive. Make the conservative assumption that these extinctions have been equally distributed over the past 400 years (which is incorrect since most of these extinctions have taken place in the last century, and extinctions are accelerating). Dividing these figures by 400 years then provides an estimate of the average percentage extinction per year, the reciprocal of which is the average species lifespan in years. This straightforward calculation yields estimated species lifespans of 29 000 years in mammals and 30 000 years in birds. These estimates are roughly 100 times lower than those typical for the fossil record, which means that recent species extinction rates have been roughly 100 times higher than background rates in the past history of life on Earth. These conservative estimates of species extinction rates based on documented extinctions in the past, however, give only a faint idea of current threats to biodiversity. The extinction of a species is a long process, which starts with declining population sizes, continues with the extinction of local populations, and ends with the disappearance of the last individual in the last population. Once a species is driven extinct locally, it can also generate a cascade of other extinctions in species with which it interacted. Not only is the final step of species extinction difficult to demonstrate, it also takes time — sometimes a lot

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of time. Unless their entire habitat is destroyed at once or they are killed massively through hunting or pollution, species may be committed to near-certain extinction and yet persist for many years, because individuals continue to survive or reproduce, but births are no longer able to compensate for deaths in the long run. Many extant species are effectively “living dead” (Janzen 1986): they are doomed to extinction but have not yet had the time to fulfil their destiny. To some extent, of course, all species are “living dead” since all species are ultimately doomed to extinction. But in a time when extinctions accelerate and are no longer balanced by speciations, as is the case today, not accounting for species that are already committed to extinction results in a massive underestimation of extinction rates. Therefore current generations are accumulating a huge extinction debt that will be paid by future generations (Tilman et al. 1994). The magnitude of the current acceleration of biodiversity loss and its longterm consequences cannot be overstated. If the current extinction debt is taken into account, estimates of species extinction rates increase by roughly an additional factor of 100. There are several methods to derive estimates of species extinction rates that incorporate the extinction debt. One traditional approach has been to use empirical species–area relationships. The relationship between the number of species, S, found in an area and the surface of this area, A, is approximately a power law of the form in Eq. (13) in Chapter 2. It follows from this law that if surface area is reduced by some small proportion ΔA, the proportional reduction in the number of species, ΔS, is approximately (May et al. 1995): ΔS = zΔA

(15)

The coefficient z of the power law varies depending on the spatial scale and isolation of the areas considered, but is roughly 0.25 on average (Chapter 2). It results from Eq. (15), then, that the proportion of species lost should be roughly one quarter of the proportion of area lost. Since species–area relationships describe statistical spatial patterns, they say nothing about the mechanisms that generate them or about the time it takes for the number of species to adjust to the area available. Therefore the above equation provides only rough estimates of committed species extinctions, including the extinction debt, following habitat destruction. It does not help predicting the timing of these extinctions, and hence does not provide an estimate of the number of extinctions that have already taken place. This simple approach predicts alarming rates of committed species extinctions. Deforestation of tropical forests, one of the richest reservoirs of terrestrial biodiversity on Earth, increases at a rate estimated somewhere between 0.8% and 2% per year. This yields an estimated proportion of species committed to extinction of 0.2–0.5% per year, corresponding to an average species lifespan

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of around 200–500 years (Wilson 1992; May et al. 1995). 200–500 years is a ridiculously short lifespan on evolutionary time scales, 8000–20 000 times shorter than the average species lifespan in the fossil record! These alarming figures do not seem to be an artifact of the method based on species–area relationships. Alternative approaches based on changes in the IUCN Red Lists through time also predict an average species lifespan of 200–400 years in mammals and birds currently (Mace 1995; May et al. 1995). Red List Indices measure the relative rate at which a particular set of species change in overall threat status, based on population and range size and trends (Butchart et al. 2005). Red List Indices for birds have declined systematically in recent years, particularly in the Indomalay biogeographic realm, indicating increasing extinction risks (Fig. 11). And projections of species distributions for future climate change scenarios predict that climate change alone, the first signs of which are only beginning to be apparent, might commit 15–37% of all terrestrial species to extinction by 2050 (Thomas et al. 2004). A comparison with the five episodes of mass extinction in the past history of life on Earth is appropriate here. These episodes were characterised by unusually high extinction rates in the fossil record, i.e., 75–95% of species becoming extinct in a million years. Since the background extinction rate is 25%, this means an additional 50–70% in a million years. At current rates of committed extinctions, we could easily lose 50% of extant species in the next 200 years or so, thus rivalling past mass extinctions in only a few centuries!

Fig. 11. Red List Indices (set to 100 in 1988) for birds for 1988–2004 in different biogeographic provinces. Decreasing indices mean increasing extinction risks. After Butchart et al. (2005)

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Projections of biodiversity loss for the more distant future are even more pessimistic. The above traditional projections based on species–area relationships assume a coefficient z of about 0.25, which is a reasonable average value within continents or across not too distant islands, i.e., in situations where organisms can maintain or re-establish populations through dispersal from source areas. But for most organisms that are not associated with humans the Earth is effectively a “shrinking island” (Rosenzweig 2003): its effective spatial extent is being reduced to that portion of the Earth’s surface which is not heavily transformed by humans. Ultimately this “shrinking island” should obey the rules typical for isolated biogeographical provinces, i.e., the number of species it harbours should be roughly proportional to area (z ≈ 0.9–1, see Chapter 2). Since protected areas designated explicitly for biodiversity protection currently span only 5.1% of Earth’s land area (Hoekstra et al. 2005), this means that species diversity should ultimately stabilise somewhere around 5% of its current level, and hence that about 95% of extant species should go extinct. How long it would take for diversity to stabilise at such a level — if it ever does — is, however, unknown. All these figures make our head spin. Are they reliable? Unfortunately, there is no way to know for sure, and by the time we get solid data it will be much too late. But, with hindsight, these figures make perfect sense. Humankind is reaching the point where it is impacting on the biosphere as a whole. As a result, not only is space shrinking for most species, but time is shrinking too. The characteristic times of the natural processes that are strongly impacted by humans, including species extinctions, are increasingly converging toward those of human social processes. Based on this straightforward observation, it is not difficult to predict that the lifespan of many species may well be decreased by yet another order of magnitude in the not too distant future if current trends are not halted or reversed, thus propelling us into a mass extinction of truly unprecedented magnitude in the history of life on Earth. It is important to keep in mind, however, that the above figures are projections of what is likely to happen if current trends continue, not predictions of what will actually happen. It is still possible for humankind to reverse these trends, at least partially, with appropriate conservation strategies and societal changes (Chapter 7). Also, these global estimates are based on fragmentary information. Inevitably, there will always be wide variations among geographic areas and taxonomic groups with regard to their degree of vulnerability. While taxonomic diversity allows illuminating comparisons with the past, it is but one component of biodiversity (Chapter 2). Current declines in species diversity are matched by similar declines in other components and levels of biodiversity. A wide range of natural or semi-natural ecosystems have been, and still are, converted into high-intensity monoculture croplands,

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reducing ecosystem and landscape diversity dramatically in many parts of the world. Genetic diversity has also declined globally. In particular, intensification of agriculture and economic globalisation have led to a substantial reduction in the in situ genetic diversity of domesticated plants and animals in agricultural systems. The FAO reported that 9% of the 7616 known breeds of domesticated animals have become extinct, and 20% are at risk of extinction (FAO 2007), but these figures are underestimates. Most of the extinct breeds have not been reported; as many as 5000 may have been lost over the past century (Hassan et al. 2005). Also, a large proportion of extant breeds have an unknown status in FAO statistics; among those breeds that have a known status, 24% of mammal breeds and 52% of bird breeds are threatened with extinction. Loss of genetic diversity also reduces the ability of wild populations to resist and recover from environmental fluctuations, thus further increasing the extinction risk of small populations.

The Causes of the Current Biodiversity Crisis The proximate causes of the current biodiversity crisis are relatively well known, at least qualitatively, and are all anthropogenic. They have been recently summarised by the Millennium Ecosystem Assessment (Fig. 12). The most important driver of biodiversity loss in terrestrial ecosystems has been historically, and still is, changes in land use resulting in habitat destruction for many wild species. Deforestation and the conversion of forests and grasslands into agricultural and urban systems have contributed to reduce biodiversity in temperate regions in the past, and are a still major threat to biodiversity in most drylands of the world. The biggest threat to global biodiversity currently, however, is deforestation of tropical forests. Tropical forests are biodiversity “hotspots”; they harbour a huge species diversity, most of which is still unknown. Their current destruction and fragmentation is likely to result in massive losses of species in South America, Africa, and especially Asia. Habitat change is also heavily impacting on biodiversity in freshwater ecosystems. Stream canalisation, dam construction, and water withdrawals from rivers and lakes for irrigation and urban or industrial use have increased dramatically during the past century, and have deeply altered the habitat of many freshwater species worldwide. These species are particularly vulnerable as they are often habitat specialists. Introduction of invasive exotic species has been a major cause of species extinction during the last few centuries, especially on islands. Although most species introductions go unnoticed and have few if any consequences, a small

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Fig. 12. The main proximate drivers of biodiversity loss in the various biomes of the world. After the Millennium Ecosystem Assessment (MEA 2005a)

proportion of them have considerable negative impacts on native species and ecosystems, altering nutrient flows, hydrology, fire regimes, water quality and access to resources, and eventually driving native species to extinction (Williamson 1996). The main cause of documented species extinctions on islands, however, has been the intentional or accidental introduction of exotic predators, parasites and diseases (Sax and Gaines 2008). Endemism is particularly high on remote islands due to their isolation. Species that evolved there often lacked appropriate defenses against the exotic predators, parasites and diseases introduced by Europeans when they colonised these islands, and accordingly, many of them were decimated. Biological invasions are now

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spreading rapidly worldwide with the fast-growing globalisation of capital and the ceaseless movements of goods and people it generates. Agriculture and aquaculture are moving many food species and their attendant pests and pathogens around the globe, with major impacts on native biodiversity (Naylor et al. 2001). Overexploitation of biological resources is another major driver of biodiversity loss. Hunting was probably the single most important cause of species extinctions in prehistoric times, and hunting and fishing are still a significant threat today, especially in oceans and in tropical savannas and forests. Decimation of large-sized terrestrial mammals and birds through hunting seems to have occurred throughout the history of the human species. By contrast, overfishing is a relatively recent phenomenon, which results from technological changes in fisheries that have increased fishing pressure considerably during the past 50 years. Global fisheries landings peaked in the late 1980s and are now stagnating or declining despite increasing fishing effort and power (FAO 2009; Worm et al. 2009) (Fig. 13A) because of declines in fish biomass (Fig. 13B). The collapses of the Peruvian anchoveta and the cod populations off New England and eastern Canada are but dramatic examples of a general trend toward population collapses in fisheries. Fourteen percent of all assessed fish stocks globally were collapsed in 2007, i.e., had declined to less than 10% of their initial biomass (Fig. 13C). In some areas, such as in eastern Canada, the proportion of collapsed fish stocks was as high as 60% (Worm et al. 2009). The main societal response until recently has been to “fish down marine food webs”, i.e., to increase the fishing pressure on other species lower in the food chain (Pauly et al. 1998), but there are currently growing efforts to rebuild fish stocks (Worm et al. 2009). Pollution has emerged as a significant driver of biodiversity changes in temperate grassland, freshwater and coastal ecosystems in the past decades, mainly in the form of excessive nutrient loading. The growing use of fertilisers in agriculture has boosted food production in the past 50 years, but it has also resulted in the accumulation of nitrogen and phosphorus in inland and coastal waters, where it is causing eutrophication and hypoxia, two phenomena that strongly reduce local species diversity. Nitrogen deposition is also reducing species diversity in many temperate grasslands. Anthropogenic climate change is emerging as a new major threat to global biodiversity. Recent minor changes in climate, notably warmer mean temperatures, have already had significant effects on species geographic distributions, the timing of reproduction and migration events, and the frequency of pest and disease outbreaks. Elevated sea surface temperatures are thought to be responsible for recent episodes of mass coral bleaching, and are likely to become a major threat to the survival of coral reefs (Donner et al. 2007). Projected

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Fig. 13. Global trends in (A) total catches, (B) biomass estimated from research trawl surveys, and (C) fraction of collapsed taxa in fished ecosystems during the last 60 years. Data in (A) and (B) are scaled relative to the time series maximum. Collapsed taxa are defined as those where biomass declined to