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Conservation Biogeography

Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3 C OMPAN

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Conservation Biogeography

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

C OMPAN I ON W EB S I T E This book has a companion website: www.wiley.com/go/ladle/biogeography with Figures and Tables from the book for downloading

Conservation Biogeography

Edited by Richard J. Ladle and Robert J. Whittaker

A John Wiley & Sons, Ltd., Publication

This edition first published 2011, © 2011 by Blackwell Publishing Ltd Blackwell Publishing was acquired by John Wiley & Sons in February 2007. Blackwell’s publishing program has been merged with Wiley’s global Scientific, Technical and Medical business to form Wiley-Blackwell. Registered office: John Wiley & Sons Ltd, The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK Editorial offices: 9600 Garsington Road, Oxford, OX4 2DQ, UK The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, UK 111 River Street, Hoboken, NJ 07030-5774, USA For details of our global editorial offices, for customer services and for information about how to apply for permission to reuse the copyright material in this book please see our website at www.wiley.com/wiley-blackwell The right of the author to be identified as the author of this work has been asserted in accordance with the Copyright, Designs and Patents Act 1988. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, except as permitted by the UK Copyright, Designs and Patents Act 1988, without the prior permission of the publisher. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print may not be available in electronic books. Designations used by companies to distinguish their products are often claimed as trademarks. All brand names and product names used in this book are trade names, service marks, trademarks or registered trademarks of their respective owners. The publisher is not associated with any product or vendor mentioned in this book. This publication is designed to provide accurate and authoritative information in regard to the subject matter covered. It is sold on the understanding that the publisher is not engaged in rendering professional services. If professional advice or other expert assistance is required, the services of a competent professional should be sought. Library of Congress Cataloguing-in-Publication Data Conservation biogeography / edited by R. J. Ladle and R. J. Whittaker. p. cm. Includes index. ISBN 978-1-4443-3503-3 (cloth) – ISBN 978-1-4443-3504-0 (pbk.) 1. Conservation biology. 2. Biodiversity conservation. 3. Protected areas. II. Whittaker, Robert J. QH75.C657 2011 333.95′16–dc22

4. Biogeography.

I. Ladle, Richard J.

2010037700 A catalogue record for this book is available from the British Library. This book is published in the following electronic formats: eBook ISBN: 9781444390018; Wiley Online Library ISBN: 9781444390001; ePub ISBN: 9781444390025 Set in 9/11pt Photina by Toppan Best-set Premedia Limited

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2011

Contents

Preface, ix Acknowledgements, xi Contributing authors, xii PART 1: ROOTS, RELEVANCE, AIMS AND VALUES, 1 1 THE ROOTS OF CONSERVATION BIOGEOGRAPHY, 3 1.1 What is conservation biogeography?, 3 1.2 The emergence of conservation biology and conservation biogeography, 4 1.3 The scope of conservation biogeography, 7 1.3.1 To what ends?, 7 1.4 Outline of the following chapters, 11 Suggested reading, 12 2 SOCIAL VALUES AND CONSERVATION BIOGEOGRAPHY, 13 2.1 Many values, many goals, 13 2.2 The origins and values of different protected area types, 14 2.2.1 Sacred sites, 16 2.2.2 Resource and game reserves, 17 2.2.3 State and country parks, 18 2.2.4 Nature monuments and nature reserves, 19 2.2.5 Wildlife sanctuaries and refuges, 19 2.2.6 Wilderness areas, 20 2.2.7 National parks, 21 2.2.8 Community conservation areas, 22 2.3 Reserve designations from international conventions, 22 2.4 An international system for categorizing protected areas, 23

2.5

Social values and conservation practice, 26 2.5.1 Attitudes to non-native species, 26 2.5.2 Restoration and rewilding, 28 2.6 Concluding remarks, 29 For discussion, 30 Suggested reading, 30 3 BASELINES, PATTERNS AND PROCESS, 31 3.1 Introduction, 31 3.2 Ecosystem composition and function, 31 3.3 Balance versus flux, 32 3.4 Understanding ecosystems in flux, 34 3.5 Defining and using baselines, 38 3.5.1 Baselines derived from relict pristine systems, 38 3.5.2 Baselines derived from long-term ecology, 39 3.5.3 Rewilding, 41 3.5.4 The challenge of rapid environmental change, 42 3.6 Adaptive ecosystem management, 42 For discussion, 44 Suggested reading, 44 PART 2: THE DISTRIBUTION OF DIVERSITY: CHALLENGES AND APPLICATIONS, 45 4 BASIC BIOGEOGRAPHY: ESTIMATING BIODIVERSITY AND MAPPING NATURE, 47 4.1 Introduction, 47 4.1.1 Our incomplete knowledge of biodiversity, 47 4.1.2 Why do we map?, 48 4.2 Three knowledge shortfalls, 49 4.2.1 The Linnean shortfall, 49 4.2.2 The Wallacean shortfall, 54 4.2.3 The extinction estimate shortfall, 58

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Contents

4.3

The fundamental taxonomic units of conservation biogeography, 62 4.3.1 Species versus other genetically-based conservation units, 62 4.3.2 Evolutionarily Significant Units (ESUs), 64 4.3.3 Other conservation units, 65 4.4 Spatial distributions: from genes to biogeographical regions, 65 4.4.1 Mapping species individually and collectively, 65 4.4.2 Phylogeography, 72 4.4.3 Endemism, 74 4.4.4 Biogeographical regions, 75 4.5 Mapping function, 76 4.5.1 Biomes, ecosystems and communities, 76 4.5.2 Ecoregions, 82 4.6 Natural units in the marine realm, 83 For discussion, 91 Suggested reading, 92 5 THE SHAPING OF THE GLOBAL PROTECTED AREA ESTATE, 93 5.1 Origins, 93 5.2 Typology of frameworks, 95 5.2.1 Spatial classification of approaches – contiguous areas, landscape units and habitat islands, 97 5.2.2 Biogeographical (compositional) versus Ecological (functional) approaches, 100 5.2.3 Strategic goals – composition, function, numbers and attributes, 102 5.3 Terrestrial protected area schemes, 104 5.3.1 IUCN Biogeographical Regions (Dasmann–Udvardy) scheme, 104 5.3.2 Endemic Bird Areas, 106 5.3.3 Conservation International’s hotspots, 109 5.3.4 The WWF Ecoregions scheme, 113 5.3.5 Important Bird Areas and Key Biodiversity Areas, 117 5.4 Marine protected areas, 121 5.4.1 Status of the marine realm, 121 5.4.2 Origins and expansion of the marine protected area estate, 122 5.4.3 A global representative system of marine protected areas, 123

5.4.4

Reefs at risk – hotspots/ threatspots, 126 5.4.5 Large Marine Ecosystems, 130 5.4.6 WWF Global 200 – the marine perspective, 131 5.4.7 Coastal Zone Management and critical seascapes, 132 5.4.8 High seas protected areas, 132 5.5 Current trends and future directions, 134 For discussion, 135 Suggested reading, 135 6 SYSTEMATIC CONSERVATION PLANNING: PAST, PRESENT AND FUTURE, 136 6.1 Introduction, 136 6.2 What is systematic conservation planning and why use it?, 138 6.3 Concepts and principles, 138 6.3.1 Representativeness, 138 6.3.2 Persistence (adequacy), 139 6.3.3 Efficiency, 139 6.3.4 Flexibility, 140 6.4 Developing a systematic conservation plan, 140 6.4.1 Achieving representation, 140 6.4.2 Achieving persistence, 146 6.4.3 Achieving efficiency, 151 6.4.4 Achieving flexibility, 152 6.5 Decision support tools to identify and prioritize new protected areas, 152 6.6 Consultation and implementation of systematic conservation plans, 155 6.7 What does the future of systematic conservation planning hold?, 156 6.7.1 Conservation planning is a dynamic problem, 158 6.7.2 Conservation assets change through time, 158 6.7.3 A mix of conservation actions could occur at any site, 158 6.7.4 Better economics and socio-economics, 158 6.7.5 Dealing with uncertainty, 158 6.7.6 Properly accounting for threats, 159 6.7.7 Persistence – attainable goal or impractical utopia?, 159 6.7.8 How much should we invest in improving a conservation plan?, 159 For discussion, 159 Suggested reading, 160

Contents

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PART 3: CONSERVATION PLANNING IN A CHANGING WORLD, 161

For discussion, 222 Suggested reading, 223

7 PLANNING FOR PERSISTENCE IN A CHANGING WORLD, 163 7.1 Introduction, 163 7.2 Using the past to understand the present and predict the future, 164 7.2.1 Predicting future ecosystem responses to changing conditions, 168 7.2.2 Interpreting recent trends in their historical context, 169 7.2.3 Geographical range collapse, 170 7.3 Predicting biodiversity change, 176 7.3.1 Modelling the current distributions of species, habitats and biomes, 177 7.3.2 Modelling range shifts, 180 7.4 What do we do about it? Dynamic conservation planning, 183 7.4.1 Incorporating dynamic biotic and abiotic processes into conservation plans, 183 7.4.2 Changes in socio-economic factors, 185 7.4.3 Climate change, conservation planning and assisted migration, 185 7.5 Closing remarks, 188 For discussion, 188 Suggested reading, 189

9 BIOLOGICAL INVASIONS AND THE HOMOGENIZATION OF FAUNAS AND FLORAS, 224 9.1 The biogeography of species invasions, 224 9.1.1 The invasion process, 224 9.1.2 Human-assisted versus prehistoric invasions, 226 9.1.3 Economic and ecological impacts of invasion, 227 9.2 Biotic homogenization, 229 9.2.1 The process of biotic homogenization, 230 9.2.2 Different manifestations of biotic homogenization, 230 9.3 Patterns of biotic homogenization, 232 9.3.1 Fishes, 232 9.3.2 Birds, 235 9.3.3 Plants, 237 9.3.4 Mammals, 237 9.4 Environmental and human drivers of biotic homogenization, 238 9.5 Biotic homogenization and conservation, 240 9.6 Novel assemblages, 241 For discussion, 242 Suggested reading, 243 PART 4: FUTURE DIRECTIONS, 245

8 APPLIED ISLAND BIOGEOGRAPHY, 190 8.1 Introduction, 190 8.2 Implications of habitat loss and fragmentation: from theory to evidence, 194 8.2.1 The use of species–area relationships in conservation, 194 8.2.2 Relaxation and the extinction debt, 199 8.2.3 Ecosystem collapse and threshold responses in habitat islands, 203 8.3 Species incidence, 208 8.3.1 Minimum viable populations, minimum areas and incidence functions, 208 8.3.2 Metapopulation dynamics, 211 8.4 Nestedness, 213 8.4.1 Edge effects, 216 8.4.2 Habitat corridors, 217 8.4.3 Landscape context – matrix effects, 218 8.5 Emergent guidelines for conservation, 219

10 PROSPECTS AND CHALLENGES, 247 10.1 Why we need conservation biogeography, 247 10.2 The challenges, 248 10.2.1 Filling the Wallacean and Linnean shortfalls, 248 10.2.2 Improving models, simulations and forecasts, 250 10.2.3 Turning theory into practice, 251 10.2.4 Education, communication and public engagement, 252 10.2.5 Reconciliation ecology and a biogeography of the countryside, 257 10.3 Looking to the future, 257 Glossary of terms, 259 References, 264 Index, 297 Colour plates follow the Index

This book has a companion website: www.wiley.com/go/ladle/biogeography

Preface

Most scientists would agree that life on Earth is currently experiencing a rapid and dramatic transformation and re-sorting, reminiscent of some of the most dramatic events in the planet’s history, such as the switches in and out of ice ages, biotic interchanges driven by the collision of continents, or the handful of massive and geologically abrupt past biodiversity collapses termed mass extinctions. All over the globe natural habitats are being transformed to suit the needs of the local human population or those of distant markets. Sometimes these changes are dramatic, such as the clearing of lowland rain forest to make way for pasture or crops. Other changes are more subtle but nonetheless have severe ramifications for the native ecology, such as the introduction of non-native species from widely distant land masses or water bodies. Moreover, these changes are taking place against a backdrop of global climate change, which has the potential to re-draw the geographic boundaries of many ecosystems. The full consequences of the contemporary humaninduced revolution in the Earth’s biota remain to be seen, although many aspects of anthropogenic impacts are already on record and many further responses to these drivers seem inevitable. Numerous species have already become extinct and, in the absence of concerted efforts to prevent this, many more seem destined to follow them into oblivion. Another seemingly inexorable process is the convergence of communities (especially where habitats have been disturbed) caused by the assisted transport of generalist species. Known as ‘biotic homogenization’, this process overwrites the local and particular species with, very often, the same sets of successful commensal species, many of which present very significant negative impacts on ecosystems and economies. Other consequences are harder to predict, such as the impact

of climate change on the make-up of communities, or even its influence on the geographic distribution of a particular species. Ultimately, the degree to which human transformation of landscapes and ecosystems impacts on the diversity and distribution of life on Earth will be determined by the response of societies, organizations and individuals. However, to make rational informed decisions about where to invest conservation’s limited resources (both taxonomically and geographically) requires an understanding of the principles, concepts and techniques of biogeography. Although the application of biogeographical theory to conservation problems has a long history, most notably in the design of island nature reserves, we have for some time felt that there was insufficient attention being paid to exploiting biogeographical information and knowledge in the practice of conservation and in the education and training of those intent on contributing to conservation policy and practice. The goal to develop teaching material with this focus was, therefore, an integral part of our plans for a new Master’s programme in Biodiversity, Conservation and Management, taught in the University of Oxford, and in which several of our colleagues – contributing authors to this book – have also been involved over the last six years or so. Encouraged by the interest in conservation biogeography evident within the mission statement (below) and membership of the recently formed International Biogeography Society, we set out to outline the aspirations, applications and limitations of the field in a prospective review paper published in 2005 in the journal Diversity and Distributions – A Journal of Conservation Biogeography. This paper, however, only gave the bare bones of a cohesive set of concepts and criticisms that define an important new perspective on global conservation.

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Preface

Our aim in this book is therefore to expand the scope and agenda of conservation biogeography, to identify critical gaps and weaknesses, and to provide an introduction to the toolbox of concepts and methods – and thereby to produce a broad-based text for university courses and programmes. To this end, we have tried to provide a strong pedagogic structure, starting with the values and imperatives underpinning conservation that determine where and how biogeography can be used. Subsequently we develop the key concepts and frameworks before concluding with the application of biogeography to real world conservation problems. To ensure adequate coverage of all of the most important areas of contemporary research and practice, we invited a number of colleagues, based at other universities around the world, to join with us in developing the first comprehensive textbook on conservation biogeography. Edited textbooks can often appear bitty and may lack cohesion, but it has been our great good fortune as editors that our colleagues in this venture have accepted with good grace – and, indeed, enthusiasm – our efforts to avoid such an outcome. Working within a pre-set structure, the author teams have developed, collaboratively, a text which we hope you, the reader, will find to be readable, coherent, insightful and comprehensive. Biogeography is entering a period of innovation, growth and expansion, and it has been reinvigorated by its fusion with the younger, crisis-driven agenda of conservation science. We hope this book will contribute in a small way to attracting more students and established scientists to work in this newly emerging field, and we look forward with great interest and expectation to tracking the practical and conceptual development of the discipline in the coming years. Finally, in furtherance of this goal, and together with our fellow contributing authors, we have pledged to

donate the royalties from the present work to the International Biogeography Society, which was founded as a non-profit organization in 2000 with the following mission: • Foster communication and collaboration between biogeographers in disparate academic fields – scientists who would otherwise have little opportunity for substantive interaction and collaboration. • Increase both the awareness and interests of the scientific community and the lay public in the contributions of biogeographers. • Promote the training and education of biogeographers so that they may develop sound strategies for studying and conserving the world’s biota. For further information on the IBS visit http://www. biogeography.org. Richard J. Ladle Viçosa, Brazil

Robert J. Whittaker Oxford, UK

July 2010 Richard Ladle was the founding course director of the MSc Biodiversity Conservation and Management, School of Geography and the Environment, University of Oxford. He is currently a Senior Research Associate of the School and a Visiting Professor in the Department of Agricultural Engineering, Federal University of Viçosa, Brazil. Robert Whittaker is the current Academic Director of the MSc Biodiversity Conservation and Management, and holds the title of Professor of Biogeography in the School of Geography and the Environment, University of Oxford. He was a founding member of the International Biogeography Society and is currently editor-in-chief of the Journal of Biogeography.

Acknowledgements

First and foremost we wish to thank our students (especially from the MSc Biodiversity, Conservation and Management course, University of Oxford) for discussion in class of many of the topics covered in this book. As in any such project, many individuals have contributed to shaping the content of this book by engaging with members of the author team in discussion, by supplying artwork and permission to use it, and by providing feedback on draft chapters. We particularly wish to thank in these regards: Natalie Ban, Peter Baxter, Josie Carwardine, Megan C. Evans, Mat Gilfedder, Carissa J. Klein, Vincent Devictor, Lincoln Fishpool, Helen Fox, Richard Grenyer, Steve Jennings, Liana N. Joseph, Mike Hopkins, Andrew T. Knight, Mark V. Lomolino, Jeffrey D. Lozier, Mikko Piirainen, Thomas K. Pool, Timoth Rayden, Carsten Rahbek, David M. Richardson, Robert J. Smith, Christopher Stewart, Jens-Christian Svenning, Sebastian Troeng, and Katherine J. Willis. Carissa J. Klein and Robert J. Smith kindly contributed to two boxes within Chapter 6. We thank Ailsa Allen for redrafting several of the figures. We thank the commissioning editor Ward Cooper, copy-editor Brian Asbury, and the production team at Wiley-Blackwell, in particular Kelvin Matthews and Camille Poire. We thank the authors, publishers, and other institutions who have kindly given us their permission to

reproduce or re-draw artwork originally published elsewhere. We apologize if any permissions requests have been overlooked in error. Contributing authors to the book wish to recognize support received, as follows: Miguel Araújo thanks Delta Cafés for supporting the Rui Nabeiro Biodiversity Chair at the University of Évora. His research is also funded through the EC FP6 ECOCHANGE project (Challenges in Assessing and Forecasting Biodiversity and Ecosystem Changes in Europe, contract no. 036866-GOCE). James Watson, Richard Fuller and Hedley Grantham are supported by the Applied Environmental Decision Analysis research hub, funded through the Commonwealth Environment Research Facilities programme, Australia. Lindsey Gillson is funded by the National Research Foundation (South Africa) and the African Climate and Development Initiative (UCT). Catherine Parr is supported by the Trapnell Fund and the Higgins-Trapnell Family Foundation. Hugh Possingham is an ARC Federation Fellow and his work is supported by The Australian Research Council and an Australian Commonwealth Environmental Research Facility grant. Kostas Triantis is supported by a FCT Fellowship (SFRH/BPD/44306/2008). Kerrie Wilson is an ARC Research Fellow and her work is supported by The Australian Research Council.

Contributing Authors

Miguel B. Araújo National Museum of Natural Sciences, CSIC, Madrid, Spain and University of Évora, CIBIO, Évora, Portugal Shonil A. Bhagwat School of Geography and the Environment, University of Oxford, Oxford, UK Richard A. Fuller School of Biological Sciences, The University of Queensland, Brisbane, Australia and CSIRO Climate Adaptation Flagship and CSIRO Sustainable Ecosystems, Brisbane, Australia Lindsey Gillson Plant Conservation Unit, Botany Department, University of Cape Town, South Africa Hedley S. Grantham School of Biological Sciences, The University of Queensland, Brisbane, Australia Paul Jepson School of Geography and the Environment, University of Oxford, Oxford, UK Richard J. Ladle School of Geography and the Environment, University of Oxford, Oxford, UK and Department of Agricultural Engineering, Federal University of Viçosa, Brazil Julie L. Lockwood Ecology, Evolution, and Natural Resources, Rutgers University, New Brunswick, NJ, USA

Sara A. Lourie Redpath Museum, McGill University, Montreal, Canada Julian D. Olden School of Aquatic and Fishery Sciences, University of Washington, Seattle, USA Catherine L. Parr Environmental Change Institute, School of Geography and the Environment, University of Oxford, Oxford, UK Hugh P. Possingham School of Biological Sciences, The University of Queensland, Brisbane, Australia Brett R. Riddle School of Life Sciences, University of Nevada, Las Vegas, USA Kostas A. Triantis Azorean Biodiversity Group, University of Azores, Terceira, Portugal and School of Geography and the Environment, University of Oxford, Oxford, UK James E.M. Watson School of Biological Sciences, The University of Queensland, Brisbane, Australia Robert J. Whittaker School of Geography and the Environment, University of Oxford, Oxford, UK Kerrie A. Wilson School of Biological Sciences, The University of Queensland, Brisbane, Australia

PART 1 ROOTS, RELEVANCE, AIMS AND VALUES

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

CHAPTER 1 The roots of conservation biogeography Robert J. Whittaker1 and Richard J. Ladle1,2 1

School of Geography and the Environment, University of Oxford, Oxford, UK Department of Agricultural Engineering, Federal University of Viçosa, Brazil

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1. 1 WHAT I S C ONSE R V AT I ON BI O G EO GR AP HY? For those poor souls trapped in narrow scientific disciplines there may be an excuse for introspection, but that is not the nature of biogeography. Biogeographers surely have a special responsibility to broaden the perceptions and awareness of policy makers and those coming new to the profession. (Koy Thomson, 1991, p. 477) As many others have argued before us, we live in an era, sometimes dubbed the Anthropocene, in which our species has increasingly shaped the world around us, influencing the physical and biological components of ecosystems at every scale from that of our own immediate surroundings up to the whole Earth system. We have developed, sometimes purposefully, but often haphazardly and accidentally, the habit of extinguishing species of plants and animals, domesticating them, assisting their spread to new territories, messing around to varying ends with their genetics and biology, and resorting them into novel communities embedded in socalled ‘cultural’ landscapes. In short, we have become the dominant force in altering the distribution, composition and diversity of life on Earth, with outcomes that are sometimes beneficial to the human condition and sometimes not, depending on which changes we are considering and the perspective of the observer.

Biologists have documented and modelled these changes in ecology and biogeography while articulating ever-increasing concern over the many perceived threats to biodiversity. The modern conservation movement has grown and evolved in response to these threats, with the most prominent national and international conservation organizations setting their sights on using the best possible scientific guidance to target resources on conserving whatever aspect(s) of biodiversity they value most highly. This book is written with the aim of providing a resource for those wanting to contribute to this endeavour. There are many books on conservation biology, so it is valid to ask why we need one on conservation biogeography and what is the operational remit of the field? We have previously defined ‘conservation biogeography’ in the following terms: ‘the application of biogeographical principles, theories, and analyses, being those concerned with the distributional dynamics of taxa individually and collectively, to problems concerning the conservation of biodiversity’ (Whittaker et al., 2005, p. 3). As shown schematically in Figure 1.1, this identifies conservation biogeography as a sub-set or sub-field of conservation biology. If it is a sub-field, then it is one with deep roots in the natural sciences. In broad terms, conservation biogeography is concerned with pattern and process over large extents of space (and time), so we have therefore mostly excluded from this book

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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The roots of conservation biogeography 1. 2 T H E EMER GEN CE OF CON S ER V AT I ON B I OLOGY AN D CON S ER V AT I ON B I OGEOGR APH Y Real-world biogeographers must balance their roles as citizens wanting to make the world a better place against their roles as scientists who are honest and skilful about their scientific limitations. Getting the science right is undoubtedly important for making policy decisions, but a wise approach to uncertainty is all the more so. (Thomson, 1991, p. 475)

Figure 1.1 Approximate chronology of the emergence of conservation biogeography, positioning the subject as a subset of conservation biology focused on pattern and process at coarser scales of analysis. The early conservation movement, labelled here wildlife conservation, in fact comprised a broader set of values, including the recognition/prioritization of wilderness and of natural monuments. In this period, much of the emphasis was on moral/aesthetic values. The emphasis on conservation of nature, embracing ecological/scientific grounds for conservation, was around from a comparatively early point but gained ground mostly during the latter half of the 20th century, giving rise to the distinct academic field of conservation biology in the final quarter of the century. The subject areas identified at each scale are illustrative not exhaustive (MVP, minimum viable population; ETIB, equilibrium theory of island biogeography; SLOSS, single large or several small reserves). Missing from this chronology is the emergence of the concept of biodiversity at the end of the 1980s (see Box 1.1).

material concerned with wildlife behaviour, population level processes, tracking and monitoring of wild animal and plant populations, and field ecology. These subject areas are important, but they are well covered in many conservation biology texts. When it comes to the coarser biogeographical scales of analysis, on the other hand, we feel that there has been something of a lag in translating updates of theory and practice into curricula. This is particularly important because many strategic conservation decisions of great significance to the ultimate effectiveness of conservation, such as how protected area networks are designed, require a deep understanding of this coarser-scale biogeographical science – the subject matter of this book.

As an applied and interdisciplinary science concerned with the conservation of nature, conservation biogeography can be seen as a product both of biogeography and of conservation biology. We briefly consider the origins of these related endeavours by the order of their emergence (Figure 1.1). Biogeography is the study at all scales of analysis of the distribution of life across space and how it has changed through time. In the broadest sense, biogeography could even be thought of as the ‘first science’, because the ability to understand and track the distribution of food and predators though time and space was arguably of even greater interest to our huntergatherer ancestors. Although the term biogeography appears to have been a 20th century innovation, the discipline has deep scientific roots. Many of the core principles and broadly known patterns of biogeography were established and debated before the end of the 19th century under the twin headings of zoogeography and phytogeography, by such towering figures as Alfred Russel Wallace (sometimes called the father of zoogeography), Charles Darwin, Philip Sclater, Georges-Louis Leclerc (Compte de Buffon), and Alexander von Humboldt (Lomolino et al., 2004, 2006). Indeed, some of the major themes were already established as areas of enquiry by the early 1800s: indicative of the foundational nature of the subject within the natural sciences (Lomolino et al., 2004, 2006). The study of biogeography thus developed in advance of the coalescence of theory and thinking that came to constitute the disciplines of ecology and evolution, with which the subject of biogeography is naturally intertwined. Biogeography has many facets, traditions, and schools of thought. They encompass deep time (historical biogeography), the recent past (palaeoecology) and contemporary pattern and process (ecological

Roots, relevance, aims and values biogeography). At the core of the discipline is an interest in describing, explaining and predicting patterns of distribution and diversity, whether at higher taxa level, species level, or most recently also sub-species levels of analysis. The modern conservation movement emerged in the late 19th century in response to fundamental changes in world views concerning the nature of the relationship between humans and the natural world, and it emanated largely from the elite society of the American East Coast and Western Europe (Jepson & Whittaker, 2002a). The movement was motivated both by a desire to preserve sites with special meaning for the intellectual and aesthetic contemplation of nature, and by acceptance that the human conquest of nature carries with it a moral responsibility to ensure the survival of threatened life forms. These early principles were later combined with a range of more utilitarian perspectives but, over the first half of the 20th century, the primary motivating forces were the conservation of wildlife, especially large game animals and birds, and the desire to preserve places of natural beauty and wonder (Figure 1.1, Box 1.1). Over time, the conservation movement diversified with increasing recognition of other (often deeply rooted) motivating ideas alongside the immediate imperative of saving particular types of species from extinction. Thus, nature conservation can be thought of as a social movement working to develop or reassert certain values in society concerning the human/ nature relationship (Jepson & Whittaker, 2002a). The movement gained new momentum in the second half of the 20th century, when science and environmentalism further expanded understandings of our relationship with nature (Frank et al., 1999; Adams 2004). Motivated by, but distinct from, the nature conservation movement, ‘conservation biology’ is the name given to applied research designed to inform management decisions concerning the conservation of biodiversity. As such, its roots lie largely within the mid 20th century. Conservation biology gained huge momentum during the 1970s and early 1980s, when it was formally identified as a sub-discipline, with dedicated journals and textbooks (e.g. Soulé 1986; Primack, 2002) and learned societies such as the Society for Conservation Biology, founded in 1986 and presently with over 10,000 members. Conservation biology can be defined narrowly as being concerned with the application of population biology, taxonomy and genetics to problems

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concerning the conservation of biodiversity. In a now classic paper, Graham Caughley (1994) pointed out that conservation biology research mainly operates within two overriding paradigms – studies that seek an understanding of the proximate causes of population decline (the declining population paradigm) and those that are concerned with the consequences of small population size (the small population paradigm). More recently there has been an acknowledgement that a biological understanding of rarity and endangerment is necessary, but not sufficient for policymaking designed to prevent biodiversity loss, and that conservation biology needs to incorporate a far broader range of disciplines. Indeed, conservation biology textbooks typically include a wider array of basic scientific and other academic disciplines, including such fields as anthropology, biogeography, environmental economics, environmental ethics, sociology and environmental law (e.g. see Primack, 2002). The incorporation of the social sciences under the umbrella of conservation (i.e. biological) science represents a recognition of the need to apply multiple forms of scholarship to address complex real-world problems. It may also reflect a general desire to bestow scholarly discourse and guidance with the additional gravitas associated with a ‘proper science’ given the status of scientific guidance within late 20th century society and politics (see, e.g. Knight & Cowling, 2007). Last to emerge within the framework shown in Figure 1.1 is conservation biogeography, a term that began to gain currency via a conference of the International Biogeography Society in Washington in January 2005 (Lomolino & Heaney, 2004), and that we ourselves promoted through a paper published the same month in the journal Diversity and Distributions, which simultaneously gained the sub-title ‘A Journal of Conservation Biogeography’. We subsequently realized that the term had been first coined at least twelve years earlier by John Grehan (1993) in the title of a paper in the first issue of the same journal (then called Biodiversity Letters), although he did not define in any precise way what he meant by the term. Whereas the formal use of the term may be recent, the use of biogeography within conservation biology has been going on for as long as scientific guidance to conservation has been offered, and biogeography indeed formed a central part of early theory within conservation biology (see, e.g. Primack, 2002). Classic foundational works combining biogeographical analysis with conservation guidance include the early papers

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The roots of conservation biogeography

Box 1.1 General characteristics of three broad themes and phases of the conservation movement The modern conservation movement can trace its roots back to various societies, clubs and social movements that formed in the late 19th century in response to a variety of changes in society and the natural world (reviewed in Chapter 2; and in Jepson & Ladle, 2010). Two themes were particularly prominent: 1 Wildlife conservation, in the sense of ‘[t]he controlled use and systematic protection of indigenous fauna’. (Matthews et al., 2001) This part of the early conservation movement was rooted in the natural history and hunting pastimes of elite society of the 19th century and is exemplified by the creation of the Boone and Crockett club (B&CC) in 1887 by future US President Theodore (Teddy) Roosevelt. The B&CC had two main goals: first, to create sanctuaries and refuges where wildlife could survive the onslaught of human expansion into frontier lands; and second, to change the culture of hunting to exalt the noble qualities of the chase above the number of animals killed (Jepson & Ladle, 2010). 2 Nature conservation, which stemmed from ideas of nature as being delicate and intricate system(s) sensitive to human interference. This idea manifested itself in the preservation of the status quo and of nature monuments – places for the contemplation of nature, antidotes to urban life. The concept of nature monuments was promoted particularly by the German forester Hugo Conwentz (discussed further in Chapter 2). Both nature conservation and wildlife conservation sometimes resulted in a ‘fortress conservation’ approach, where areas were fenced off from the local communities in order to save them – or at least save them for exclusive exploitation by Western interests. 3 Biodiversity conservation. The contemporary approach to conservation embraces both of the above sets of values and more besides, but can be crudely characterized as distinct from the above by its emphasis on attempts to conserve biodiversity. The term ‘biodiversity’ is simply a contraction of ‘biological diversity’ and may have first been used in a scientific study by Elliot Norse in a US government report in 1980. However, it is more commonly attributed to Walter Rosen around 1985 while planning a symposium; it was used in the title of the resulting 1988 symposium volume (Wilson, 1988a) and subsequently gained rapid uptake. Biodiversity has many definitions, one prominent one being ‘[t]he variability of life from all sources, including within species, between species, and of ecosystems’ (Matthews et al., 2001). It is, in its character, a scientific/technical term, although it is important to bear in mind that the study of biodiversity is not solely a branch of biology as it has an ethical/social dimension (Jeffries, 1997). Moreover, some commentators have noted that biodiversity definitions are often closer to subjective ‘value judgement’ concepts such as quality of life than an objective measure of an environmental property (Lambshead & Boucher, 2003). Although most definitions of biodiversity stress the complexity of life at multiple levels (e.g. genes, species, ecosystems), ‘biodiversity conservation activities are typically directed toward species’ (Matthews et al., 2001, p. 50).

applying island theory to the problem of habitat fragmentation (e.g. Diamond, 1975a) and Dasmann’s (1972) biogeographical regionalization approach to designing networks of protected areas (see Chapters 8 and 5, respectively). Although biogeographical science has played its part alongside other sub-fields of biology in the emergence

of current scientific guidance for biodiversity conservation, in our view it has done so as something of a poor relation – a Cinderella within conservation biology. We argue that biogeography can now cast aside its metaphorical rags as it emerges as a subject area of central importance to conservation planning. In part, this repositioning of biogeography is the result

Roots, relevance, aims and values of the recent and huge technological advances in biogeographical data collection, storage and analysis, which have enabled rapid progress in many areas of the field, both pure and applied; and, in part, it reflects theoretical and conceptual advances (e.g. Williams et al., 2000; Lomolino & Heaney, 2004). Yet we must also recognize that the underlying species distributional and other data often remain highly problematic, protocols for analysis are still in the early stages of development, and we have only recently begun the task of systematically analysing the sensitivity of our analyses to the starting assumptions and scale effects. There is an enormous degree of uncertainty in our science when it comes to predicting future distributions of taxa, diversity and biogeography (see Chapter 7). Accordingly, we argue that there is a need for more biogeographers to engage with the problems of conservation science, and for the injection of more biogeography into training for conservation scientists and practitioners.

1. 3 T H E S C OP E OF C ONS E R V A T I ON BI O G EO GR AP HY As we have indicated above, conservation biology is a large and all-embracing field. However, if it is subdivided by scale of application, we might recognize the following subdivisions of relevant theory (Figure 1.1): 1 Population scale: the development and evaluation of biological theory spanning population biological and genetic process. This is concerned with deterministic processes of population decline, population viability, genetic erosion from small populations, competitive influence of invasive species, behavioural ecology and so forth, i.e. concerned with processes in which biogeography is generally not prominent (e.g. see Caughley, 1994; Primack, 2002). 2 Landscape scale: theory concerning processes at the local–landscape scale, including the foundational influence of R.H. MacArthur and E.O. Wilson’s equilibrium theory of island biogeography, the derivative Single Large or Several Small reserves (SLOSS) debate, habitat corridors and matrix effects, metapopulation theory and nestedness (reviewed by Whittaker and Fernandéz-Palacios, 2007), i.e. issues clearly bridging ecology and biogeography. 3 Geographical scale: applications on a yet coarser scale in part are concerned with mapping and modelling biogeographical patterns, and they in part invoke

7

historical–biogeographical theory concerned with the distribution and explanation of geographical patterns in diversity. We see such coarser scale work on the geography of nature as being unambiguously within the heartland of biogeography (cf. Lomolino et al., 2004). Despite its undoubted importance within conservation science, we argue that it is here, in particular, that something of a ‘Cinderella’ tag applies to conservation biogeography; likewise, it is here where there is greatest need for critical attention to our science and for greater interaction between those involved in theory and application (see e.g. Lourie & Vincent, 2004). Conservation biogeography, the application of biogeography in conservation, is thus separable from the application of other areas of biology (i.e. community, population and behavioural ecology, macroecology, and genetics), most clearly at coarser scales of analysis. While the use of zoogeographic regions, areas of endemism, geographic patterns in species richness, or phylogeographic structure for conservation prioritization purposes are readily identifiable as conservation biogeography, applications at increasingly fine spatial scales, for example focused on habitat corridors or metapopulation dynamics, can be seen as simultaneously drawing from traditions in both ecology and biogeography. Similarly, macroecological analyses (referring to the analysis of the emergent statistical properties of ecological data sets (Brown, 1995)) may also be based on both ‘ecological’ traits (e.g. growth rates, propagule size, breeding system, body size) and ‘biogeographical’ traits (e.g. range size, region of origin). In illustration, efforts to develop explanatory and predictive models of invasiveness of non-native species have been made that use both sets of traits, frequently finding a biogeographical signal in the resulting models (DehnenSchmutz, 2004; Pyšek et al., 2004), which indicates that such analyses draw from both ecological and biogeographical traditions within conservation science to varying degrees. For further exploration of key scale and diversity concepts relevant to conservation biogeography, see Box 1.2.

1.3.1 To what ends? While our goal in this book is to provide students with a guide to the scientific underpinnings of conservation decision-making, it is important to recognize that such

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The roots of conservation biogeography

Box 1.2 Key diversity and scale concepts The first of the tables below is adapted from Whittaker et al. (2001) and highlights the varied meanings of the term species diversity, which itself is just one meaning of the term ‘biodiversity’ (see Box 1.1). Although the table refers to species as the unit of analysis, the terms can of course be applied at other taxonomic levels. The terms given here have been used in varied and inconsistent ways in the literature, leading to much confusion within diversity theory. Within species diversity terminology, perhaps the key distinction is between metrics and concepts recognizing differences in the number of species (which are forms of inventory diversity) and those that highlight the degree to which species are shared between or unique to (endemic to) the areas being compared (which are forms of differentiation diversity). This distinction is set out in Table B1.2b, which provides the foundational framework for the scale of application, designated by letters of the Greek alphabet, developed by the American ecologist Robert Harding Whittaker (e.g. 1960, 1977). R.H. Whittaker’s framework provided a nested sequence from inventory to differentiation forms reiteratively (Figure B1.2a). As his first inventory tier was termed alpha diversity and his first differentiation tier was termed beta diversity, he sometimes used the two terms interchangeably, a habit followed by many other authors, sometimes with confusing outcomes. In practice, diversity and other biogeographical patterns described at different focal scales of analysis may well be the outcome of different dominant processes, so it is of critical importance that the scale parameters of a study are explicitly taken into account when synthesizing information within biogeography, not least within conservation biogeography (see discussion in Whittaker et al., 2001, 2005). Table B1.2a Key diversity and scale concepts. Modified from Whittaker et al. (2001, Table 1). Diversity concepts

Meaning

species diversity

varied meaning: e.g. number of species, or indices weighted by abundance distributions of species (equitability); implying of itself no standardization of sampling

species richness

number of species, implying of itself no standardization of sampling

species density

number of species in a standardized sample, e.g. per unit area; more precise than the above but less widely adopted

species turnover, i.e. differentiation diversity

in the present context meaning compositional turnover in space between two inventory (typically alpha-scale) samples, expressed by a variety of indices or multivariate analyses, and thus qualitatively different from species richness or density

endemism

an endemic is simply a species (or other taxonomic entity) confined to a particular geographical area; a focus on areas of high numbers of endemics implies an interest in biogeographical distinctiveness (whether at species or other taxonomic level)

Scale concepts

Meaning

spatial scale

should refer to the size of the base unit used in sampling and analysis, but in practice usage of this term varies such that it may mean either or both of ‘extent’ and ‘focus’; moreover, size of sample unit is very often not held a constant (as it should be) but is allowed to vary within a study

(geographical) extent

the geographical space (distance) over which comparisons are made, whether they be using e.g. 1 m2 or 10,000 km2 sample units; i.e. of itself implying nothing about spatial scale in the strict sense

focal scale

the spatial scale at which data are analysed, being either the size of the sampling unit (also called the ‘grain’) or the unit to which these data are aggregated for analysis (e.g. local or field scale to regional scale); this concept, unlike ‘extent’, can be synonymous with spatial scale

Roots, relevance, aims and values Table B1.2b Terminology used in describing diversity patterns at different scales of analysis. This table has been compiled and modified from various sources, notably R.H. Whittaker (1975, 1977); Stoms & Estes (1993); R.J. Whittaker et al. (2001). R.H.W. tiers

Spatial scale

Description

Nature of diversity metric

point

species found at a precise point within a local community, e.g. contact of grassland plant species with a pin

inventory

Alpha

local

species richness within local communities/ patches

inventory

Beta

landscape

turnover of species between local communities within a landscape

differentiation

Gamma

landscape

species richness of whole landscape

inventory

Delta

regional

turnover of species between landscapes along major gradients of climate and/or physiography

differentiation

Epsilon

regional

the species diversity of a broad region of differing landscapes

inventory

inter-regional/ inter-provincial

replacement of higher taxa, e.g. placental mammals by marsupials

differentiation

Figure B1.2a An illustration of Robert Harding Whittaker’s diversity scale framework, showing how each scale of analysis nests within the next and highlighting the distinction between inventory diversity and differentiation diversity concepts. Re-drawn from Stoms and Estes (1993).

9

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The roots of conservation biogeography

scientific guidance, and the language in which it is couched, is value-laden (see, for example, the passionate polemic by Stott, 1998) and that there is still a debate to be had concerning which properties of nature we wish as a society to foster (Trudgill, 2001). Much of the scientific guidance and of current conservation practice assumes that this debate will have a particular and almost preordained outcome, without paying much attention to the possible validity of alternative value systems (but see: Redford et al., 2003). For instance, we might wish to emphasize saving species from extinction as the prime goal, without paying great attention to the assemblages and landscapes they occur in. Or, we may wish to emphasize the importance of intact megafaunal assemblages, aesthetic and cultural significance of landscapes, ecosystem health, or biotic integrity (cf. Callicott et al., 1999; Redford et al., 2003; Adams, 2004). These ideas are linked to a similar diversity of social values motivating conservation action in many nations, especially at local scales, but also globally (Chapter 2; Jepson & Canney, 2001, 2003; Trudgill, 2001). The decision to adopt a particular set of values is not within the bounds of science and, although conservation scientists are well placed to contribute to this debate, there is an important distinction between the processes leading to the adoption of a set of values and the process of deriving the scientific guidelines to implement these values. In our view, conservation biogeographers should be in the business of providing alternative scenarios that address differing end goals (cf. Williams et al., 2000; Dimitrakopoulos et al., 2004) if they are to place their science at the service of society so as to best inform decision-making processes. To expand on this a little, any system of conservation prioritization, even if based on the application of numerical algorithms to comprehensive data sets, ultimately reflects value judgements about which features are important and how to weigh them up (Knight & Cowling, 2007). Applying funding or protection to areas ranked highly by the chosen protocols may, in the end, diminish the opportunity for conservation elsewhere, perhaps including other areas of pressing conservation concern. On the scale of landscapes, regions and states, biogeography is well placed to inform such choices. Some, reading this introduction, may wish to contest the notion that values are separable from science at all, a view with which we have some sympathy. Indeed, as pointed out by Trudgill (2001), many of the terms in

use in biogeography and conservation biology (e.g. ‘equilibrium’, ‘alien species’, ‘native species’, ‘climax community’ and ‘natural’) are deeply value-laden and defy easy objective definition. Although not included in the chronology of ideas in Figure 1.1, crucial to the recent progression of the conservation movement has been the emergence during the late 1980s of the concept of ‘biodiversity’, a term of technical and scientific resonance but one that defies precise scientific definition (Takacs, 1996). Indeed, as noted in Box 1.1, it has been argued that biodiversity definitions are closer to being subjective ‘value judgement’ concepts (such as quality of life) than they are to being an objective measure of an environmental property. Most commonly used definitions imply in some way that biodiversity is a ‘good’ thing per se and that, conversely, biodiversity loss through human action is ‘bad’ and should be prevented or minimized. Another difficulty implicit in many of the definitions of biodiversity, including that adopted by the 1992 Convention on Biological Diversity (CBD), is that biodiversity can, and should, be both conserved and used. The extent to which we regard these goals of conservation and development as compatible or in conflict describes, to a large degree, where we position ourselves as members of our society. How might this influence our work as scientists? Similarly, others have observed that natural scientists working on conservation science problems have traditionally worked within rather static equilibrial frameworks that portray nature as unchanging in the face of abundant evidence of inherent variability and flux in many natural systems (Pickett et al., 1992; Wu & Loucks, 1995). The language used in the ecological and conservation literature, according to Stott (1998), frequently reveals a desire for ‘stability’ and ‘safety’ (the so-called ‘precautionary principle’), whereas in reality we live in a world in which change takes place all the time, in all sorts of directions and at all sorts of scales; everything is in flux (summarized from Stott, 1998, p. 1). Stott calls for biogeographers and ecologists to wake up to the non-equilibrium nature of the world around us and to re-examine the assumptions and the language we use in discussing environmental problems/ opportunities and conservation. Although the socalled ‘balance of nature’ paradigm is rapidly being superseded in scientific circles by more dynamic conceptions of nature, as a handy metaphor it still has

Roots, relevance, aims and values considerable traction in society and is commonly used in popular discourse on conservation issues (Ladle & Gillson, 2009). How much do such frames of reference continue to influence the science we conduct and how we interpret our data? While our focus in this book is very much on the way that biogeographical science can contribute to conservation, we recognize that these and other critiques of the objectivity of our science require that we pause to consider the interaction between social values, the conservation movement and biogeographical science. Hence we devote chapters in this opening part of the book both to a consideration of values motivating conservation action (Chapter 2), and a consideration of the concept of alternative ecological (scientific) baselines and how they may inform conservation (Chapter 3). From this brief outline we wish to highlight three points. First, the conceptual origins of biogeography as an academic discipline substantially predate the emergence of conservation biology. Second, conservation biogeography forms an important and distinctive (but not entirely distinct) subset of conservation biology. Third, the motivating force for these scientific endeavours is a diverse and dynamic social movement, representing varied values and world views.

1. 4 O U T L I NE OF T HE F OL L OW IN G CH A PT E R S Edited books can sometimes be a bit fragmented. In Conservation Biogeography we have tried to avoid this by imposing a strong structure, prescriptive content and strong editorial guidelines throughout the process of developing the text. It is our hope, therefore, that the book can be read either as a single cohesive narrative or as ‘stand-alone’ chapters. The book is divided into four major parts. The first part, Roots, relevance, aims, and values (Chapters 1–3) has the aim of providing the historical and philosophical context of conservation biogeography, as well as introducing key terminology and frames of reference. Chapter 2, Social Values and Conservation Biogeography, focuses on the frequently neglected subject of the values underlying decisions to prioritize or protect certain geographic areas for conservation, and how to manage those areas once they have been designated. The key point is that decisions about where, what and how to conserve may be based on hard data

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and scientific principles, but are ultimately a reflection of different values within society or the global conservation community. Chapter 3, Baselines, patterns and process, examines the two main conceptual approaches underpinning modern conservation practice – compositionalism and functionalism – and how they have profound influences on conservation objectives. It also introduces the concept of ecological baselines and how these have become important targets for conservation and restoration (although in fact there may be multiple alternative baselines, states or dynamic frames of reference for the same region or landscape). The second part of the book, The distribution of diversity: challenges and applications (Chapters 4–7), provides an overview of the current state of global biogeographical knowledge and how this knowledge has been used to better focus conservation efforts. Chapter 4, Basic biogeography: estimating biodiversity and mapping nature, focuses on what we know and what we don’t know about the distribution of biodiversity and the varying phenomena we may want to map (e.g. biogeographical regions, biomes, ecoregions, areas of endemism, evolutionary significant units, etc.). It also covers the varying approaches to species mapping and how to deal with challenges such as scale issues, representations of species ranges, and bioclimate envelope modelling and mapping. Chapter 5, The shaping of the global protected area estate, gives an overview of the history and development of protected area planning frameworks at global to regional geographical scales. The chapter splits these frameworks into two main approaches: zonal, involving the mapping of attributes of nature into a suite of broadly climatically or historically determined non-overlapping areas (e.g. ecoregions); and azonal, involving the application of biogeographical principles to identify a particular set of disconnected places across the world (e.g. hotspots, important areas). Schemes based on these contrasting approaches have become key determinants of global funding and conservation action. The final chapter of this part is Chapter 6, Systematic conservation planning: past, present and future. Here, the principles and applications of computer-based and data-intensive approaches to protected area network design are reviewed and discussed, covering important network design concepts such as complementarity, irreplaceability and redundancy and the development and application of reserve selection algorithms. The

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The roots of conservation biogeography

chapter also contains extensive examples of how to conduct systematic conservation planning in marine and terrestrial ecosystems. The third part of the book, Conservation planning in a changing world (Chapters 7–9), addresses some of the key challenges to undertaking effective conservation during a time of unprecedented ecological change. Chapter 7, Planning for persistence in a changing world, starts by comparing current ecological trends to past changes in the biota. It follows this by an extensive analysis of methods used to predict biodiversity change (e.g. species distribution models, range shift models). Using the insights from these studies, the chapter concludes with a discussion of the need for dynamic conservation planning approaches that are flexible and responsive to changing predictions about the potential fate of the natural environment. One of the key biogeographical questions in conservation is how biodiversity will respond to the continuing and widespread loss and fragmentation of natural and semi-natural habitats. Chapter 8, Applied island biogeography, explores the implications of this insularization of formally contiguous ecosystems and reviews the pervasive influence of the equilibrium theory of island biogeography on the conceptual development and practice of conservation. The chapter covers important applied issues such as the use of the species– area relationship, efforts to model metapopulation dynamics, analyses of nestedness, edge effects and the efficacy of habitat corridors and the influence of varying types of matrix habitats on the abilities of threatened species to disperse between high quality habitat. It ends with a discussion of how conceptual advances can be converted into practical solutions and guidelines for conservation. Part 3 concludes with Chapter 9, Biological invasions and the homogenization of faunas and floras, which tackles one of the greatest challenges in modern conservation: how to understand, control and manage introduced species. The chapter starts by reviewing the biogeography of invasion and the process by which regionally distinct, native communities are gradually replaced by locally expanding, cosmopolitan, nonnative communities (homogenization). After brief review of patterns of homogenization by taxon, it

addresses the human causes of this process and concludes with a discussion of what these novel assemblages might mean for the future of conservation. The final part of the book, Future directions, contains a single chapter entitled Prospects and challenges, which discusses the future of conservation biogeography and focuses on the global challenge of filling the Wallacean and Linnean shortfalls, the rapid evolution of predictive models and the necessity to develop tools and applications that fulfil the needs of society to develop sustainably. The chapter casts an eye over new technological developments, such as the latest generation of biodiversity information systems that have the potential to radically alter the amount and quality of data available to biogeographers. The book concludes with some reflections on the role of conservation biogeographers in shaping the future of conservation and how to engage more fully with society and real-world conservation issues. Each main chapter contains a selection of suggested key readings in addition to the extensive literature cited within the text. The purpose of these readings is to guide students to core texts, seminal papers or stimulating contributions that expand upon topics within the chapter. Each main chapter also concludes by raising a number of questions that could form the basis of a class discussion or may be used to test understanding of the material.

S U GGES T ED R EADI N G Grehan, J.R. (1993) Conservation biogeography and the biodiversity crisis: a global problem in space/time. Biodiversity Letters, 1, 134–140. [Of historic interest as the first use of the term ‘conservation biogeography’ of which we are aware] Lomolino, M.V. & Heaney, L.R. (eds.) (2004) Frontiers of biogeography: new directions in the geography of nature, Sinauer Associates, Sunderland, MA. Primack, R.B. (2002) Essentials of conservation biology, 3rd edn, Sinauer Press, Inc., Sunderland, MA. Whittaker, R.J., Araújo, M.B., Jepson, P., Ladle, R.J., Watson, J.E.M. & Willis, K.J. (2005) Conservation biogeography: assessment and prospect. Diversity and Distributions, 11, 3–23.

CHAPTER 2 Social values and conservation biogeography Richard J. Ladle1,2, Paul Jepson1 and Lindsey Gillson3 1

School of Geography and the Environment, University of Oxford, Oxford, UK Department of Agricultural Engineering, Federal University of Viçosa, Brazil 3 Plant Conservation Unit, Botany Department, University of Cape Town, South Africa 2

2. 1 MANY V AL UE S, M ANY GOALS The problems involved in saving species from extinction and maintaining biodiversity are typically presented to the public in technical or scientific terms, often alongside proposed technical scientific solutions, such as captive breeding and re-release or the creation of a new protected area. However, while conservation clearly draws upon science and technology to a large degree, the desire to conserve nature and the behaviours associated with this desire are expressions of underlying human values. Values are here taken to mean beliefs and ideas that inform assessments of worth and which are, by definition, socially constructed. The practice of conservation of the natural world in this context is clearly a social phenomenon, and one that often becomes strongly political when conservation ideals conflict with other societal aspirations such as poverty alleviation or economic development. In the broadest sense, conservation is about asserting (or reasserting) certain values in society concerning the human/nature relationship (Jepson & Canney, 2001). Common conservation practices such as the establishment and management of protected areas, ecosystem restoration, reintroduction of large predators or the eradication of invasive species are simply the outward expression of these values. It has been argued that all humans have an innate emotional affiliation to other forms of life and are

therefore predisposed to value life and living systems (Wilson, 1984). In the broadest possible sense this may be true, but it is also clear that the specific characteristics of the natural world that are valued by people, cultures and organizations can also vary considerably. Furthermore, these values can clash dramatically, especially where there are conflicts between local people and an animal that is immensely valued by the global conservation movement. Good examples include the problems caused by elephants in Asia and Africa, the numerous conflicts around the world caused by depredation of cattle by large felids and canids, and the social barriers to reintroducing large predators such as bears and wolves into human-dominated landscapes in Europe and North America. Effective conservation requires the support of key stakeholders, scientific evidence to support a particular conservation strategy and, equally importantly, awareness on the part of the conservation organization of differing social values. Put another way, the world can be seen as a geographic patchwork of different values that reflect cultural/societal differences, the varying frames of reference of different conservation organizations, and governmental concerns and imperatives. Moreover, contrasts in values and strategies also typically exist not just between different regions but within communities, organizations and government departments for any given area. These layers of values and the relative power of different conservation

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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Social values and conservation biogeography

stakeholders to enforce change arguably determine the chances of success of any given science-based conservation intervention far more than the quality of the science involved (Ladle & Jepson, 2008). Therefore, examining the social values that underlie different types of conservation interventions and the organizations that endorse/promote them is vital for understanding the modern landscape of conservation and for effective conservation practice. In the context of conservation biogeography, social values have been of fundamental importance in shaping the geographic foci of action (e.g. western attitudes towards African savannas or tropical rain forests) and determining the abundance, distribution and relative success of different forms of protected areas. In a more general sense, conservation will only succeed if people, willingly and voluntarily, give the natural world the space and resources it needs to thrive. Social values are what promote the self-regulation and individual restraint required for this endeavour. Such values also legitimize the conservation laws, policies and actions and engender the willing support (or at least compliance) of citizens (Cotgrove & Duff, 1991). This chapter will examine how conservation values have been expressed in the creation of different types of protected areas and in the development of overarching international conservation treaties and conventions. We will then explore how protected areas have been formalized into an international categorization system. We will also briefly consider the values inherent in certain common forms of conservation management. Finally, we will discuss the dynamic and ever-changing nature of the conservation values ‘landscape’ and consider the ways in which conservation is adjusting to a global polity that is increasingly focused on other agendas such as climate change and poverty alleviation.

2. 2 T HE OR I GI NS AND V AL UE S OF DI F F ER E NT P R OT E C T E D AR E A T Y PES Conservation values have often been split into ‘intrinsic’ and ‘instrumental’, reflecting a perceived dichotomy between the desire to protect nature for its own sake and protecting nature because of its value to humanity. ‘Intrinsic’ value arguments typically concern the aesthetic and intellectual appreciation of nature and human compassion or reverence towards other life forms and have roots in deep-seated psychological tendencies (biophilia), cultural world views and

an eclectic variety of religious perspectives. In contrast, instrumental arguments are those concerned with ensuring human survival, well-being and the potential to develop materially (Ehrlich & Ehrlich, 1992). One of the latest and most powerful expressions of instrumental values has been the rapid rise to prominence of the concept of ecosystem services – natural processes and products such as pollination and watershed protection – which play an important role in the economy. The monetary value of such services can be crudely calculated (e.g. Costanza et al., 1997), thereby facilitating the entry of conservation into the realms of economic policy and political debate. The physical expression of conservation values, whether instrumental or intrinsic, has frequently been through the gazetting of land primarily or in part for conservation – commonly referred to as ‘protected areas’. The number and area of protected areas increased dramatically from the 1970s onwards, but in the terrestrial realm this growth has now levelled off (see Figure 2.1). In 1962 there were about 1,000 terrestrial protected areas worldwide. In 2004 the count was around 104,000, covering an area of just over 20 million km2, mostly of terrestrial habitats and amounting to 12.2 per cent of the Earth’s land surface (Chape et al., 2003, 2005). Growth in marine protected areas (MPAs) is a relatively recent phenomenon and the global MPA estate of around 2.59 million km2 (0.7 per cent of the ocean surface) remains comparatively small (Figure 2.1). It is important to note that the term ‘protected area’ is something of a catch-all category that may be applied to areas that have been allocated by states or held by private interests with the primary function of conserving attributes of nature that are valued. Existing protected areas have been designated on the basis of both instrumental and intrinsic values and often through a combination of both. Indeed, it is probably better to view the terms ‘intrinsic’ and ‘instrumental’ as umbrella concepts that incorporate a wide variety of closely related values, often with strong cultural and historical roots. It should also be noted that while social values play an important role in promoting, shaping and defining the characteristics of protected areas, there are many other legal, socio-political and scientific factors that play a role (Ladle & Malhado, 2007; Figure 2.2). Given the complexity of the drivers of protected area formation, it is perhaps unsurprising that they are difficult to divide up neatly into clear thematic categories. Indeed, there is a vast array of names and terms which are applied to specific sites, e.g. ‘national park’, ‘wildlife

Roots, relevance, aims and values

15

Figure 2.1 Growth in nationally designated protected areas (1872–2008). NB. This excludes 52,932 sites with no year of establishment, hence the total area shown is rather less than the figure of around 20 million km2 cited in the text. Re-drawn from World Database on Protected Areas (www.wdpa.org).

Figure 2.2 Factors influencing the establishment of new protected areas. The creation of a major new protected area is primarily achieved through national legislation but is heavily influenced by socio-political factors and is informed by science. Re-drawn from Ladle & Malhado (2007).

refuge’ and ‘wilderness area’, to name but a few. To add further confusion, many places carry multiple names or designations. This is has arisen because the purpose, legal definition and degree of protection of a given protected area can change over time and can vary considerably between countries. The complexity and overlapping nature of designations and terminology associated with contemporary protected areas is built upon a smaller and simpler set of foundational values of the modern-day conservation movement with roots in the late 18th and early 19th centuries. These core values (Table 2.1) arose largely out of a seismic shift in western society’s view of the relationship between humans and the natural world. A series of discoveries, events and circumstances, culminating in the formulation of Darwin’s famous theory of evolution, prompted vigorous debate that transformed understandings of the place of humans within nature. Some of the most influential of these events were the sudden and well-publicized rash of extinctions (e.g. the passenger pigeon, the great auk and Steller’s sea cow), the rapid demise of the vast forests in the American Great Lakes region and the discovery of the great apes (the first scientific description of the gorilla was published in 1847), whose anatomy was remarkably similar to our own.

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Social values and conservation biogeography

Table 2.1 Foundational values of the modern conservation movement and approximate location and date of emergence and social groups involved. Social values

Movement & origins

Location & date

Access to nature and countryside is necessary for the health and well-being of urban-dwellers

Social reformers/openspaces movement

London, mid 19th century

Free enjoyment of nature is one of humanity’s most precious privileges and access to nature should not to be abridged by private right for greed or gain

Social reformers, urban planners

New York, early 20th century

Healthy ecosystems are necessary to safeguard economic growth, high quality livelihoods and social stability

Colonial scientists

Colonies, late 18th century

Natural resources should be managed for the greatest good for the greatest number in the long term

US Foresters

US, early 20th century

Landscapes evoking wilderness should be preserved as a benchmark from which to assess urban/industrial modernity and as places for spiritual, aesthetic and physical exploration and rejuvenation

Writers & artists, Wilderness movement

US, late 19th century

Human conquest of nature carries with it a moral responsibility to ensure the survival of threatened life-forms

Big game hunters, Wildlife movement

New York & London, turn of 20th century

Wanton and unnecessary slaughter of wildlife is cruel and barbaric

Big game hunters, politicians and entrepreneurs

London, early 20th century

Aesthetic and intellectual contemplation of nature are integral to the biological and cultural inheritance of many peoples and monuments of nature should be guarded from ruin

Prominent citizens

Western European cities, early 20th century

In the early days of the global conservation movement, the purpose of protected areas was far more explicitly reflected in the terminology used to describe them. For example, proponents of the late 19th century wildlife movement advocated the establishment of wildlife sanctuaries and refuges; natural historians argued for the establishment of nature monuments and nature reserves; activists of the open spaces movement pushed for country parks (UK) and state parks (US); adherents of the wilderness movement for the creation of wilderness areas; and the ‘wise-use’ resource mangers established game reserves, forest reserves and watershed protection forests (reviewed in Jepson & Ladle, 2010). These terms, along with national parks, can be viewed as forming the basis of modern protected area types and designations.

In the following examples, we describe a variety of types of protected areas and the values that have influenced their formation and management, and which, by extension, underpin the contemporary protected area system. The list of protected area types is by no means exhaustive, but rather has been chosen to reflect types of protected area that are both numerically important for modern conservation and which reflect a distinctive suite of values.

2.2.1 Sacred sites Despite the deeply rooted and universal emotional human attachment to life and living processes (Wilson, 1984, 1993), the history of human societies has been

Roots, relevance, aims and values littered with well-documented examples of environmental destruction and unsustainable exploitation (Penn, 2003; Diamond, 2004). Indeed, the (relatively) lower ecological impact of people in many traditional societies has been argued to be primarily a function of their low population density and the limited availability of technology that would allow greater levels of exploitation (Ruttan & Borgerhoff-Mulder, 1999; Penn, 2003). Interestingly, there is also no obvious association between a society holding strong beliefs about the sacredness of nature and lower levels of environmental destruction (Low, 1996). Even so, the practice of restricting how resources are exploited or accessed on religious or spiritual grounds is widespread across human societies, and there are many cultures that refrain from exploiting particular areas or species (Colding & Folke, 2001). These beliefs are being increasingly considered as potential instruments for conservation practice, either on their own or nested in more formal arrangements (Barrett et al., 2001). The motivations for prohibitions or taboos (a word derived from Polynesian languages and culture to mean something that is forbidden) are as varied as the belief systems themselves. For example, in Northern Madagascar there is a complex system of taboos (known locally as ‘fady’) to which many Malagasy people adhere and which suffuses every element of their life (Ruud, 1960). Fady covers both habitats and particular resources and plays an important role in regulating a number of natural resource uses, including the exploitation (and protection) of a particular species or habitat (Mannle & Ladle, in press). In southern Madagascar, Lingard et al. (2003) reported that local people considered radiated tortoises as fady, a cultural association which may have saved them from extinction. Perhaps the most important type of taboo for conservation, and the basis of many informal, culturally prescribed protected areas, are habitat taboos. These form the basis of protection (from local communities) for a large number of natural sites around the globe dedicated to ancestors or deities. Bhagwat & Rutte (2006) refer to these informal protected areas as ‘natural sacred sites’ and argue that they are an important and often overlooked addition to the global network of protected areas. Indeed, natural sacred sites may be especially important because they are typically found in remote areas that are high in biodiversity but which have low levels of formal protection.

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Although the origins of many of these sites are impossible to uncover, there is some evidence to suggest that sacred forest groves in India existed before the advent of agriculture, and that this was the primary reason for these patches to be spared even though the surrounding land was cleared (Hughes & Chandran, 1998). As outlined above, sacred groves and other natural sites protected for spiritual and historical reasons probably have origins in pre-agricultural societies and animistic belief systems. Many sacred sites around the world are currently threatened as mainstream religions and secular values replace traditional beliefs. This process is by no means a recent phenomenon. In ancient Europe (4th–5th century AD), the sacred groves of the druids were destroyed with the arrival of Christianity (Matthews & Matthews, 2002). More recently, in India, local folk deities have been (and continue to be) replaced with Hindu deities – a process referred to as ‘Sanskritization’ (Kalam, 1996). The last fifty years have seen a new threat to the ancient belief systems that support sacred sites as local customs are being increasingly challenged by globalized western consumerist culture. The globalization of culture, although not a direct threat, is partly responsible for eroding the cultural importance of sacred sites for younger generations of local people (Bhagwat & Rutte, 2006).

2.2.2 Resource and game reserves Managing natural resources in a planned and rational manner has a long history. For instance, Schama (1995) describes how the strategic and political importance of timber for ship-building for the navy and for domestic use and iron smelting created powerful conservation voices within the courts of Tudor England (1550–1650), which prompted the creation of systems of forest reserves, regulations, penalties and guards. Protecting sites in order to maintain the quality of the hunting also has an ancient history. In medieval Europe, many kings and aristocrats maintained private hunting reserves with strict control of poaching. Grove (1996) traces the roots of western environmentalism to scientific societies formed in the 18th century by individuals employed by European companies to investigate the economic potential of unfamiliar flora, fauna and geologies in new territories. Through observing and debating the environmental changes

18

Social values and conservation biogeography

wrought by imperialism, these scientists significantly influenced the attitudes and policies wrought by the administrative elite of the European colonial powers. In particular they provided the first evidence of the association between deforestation and famine, and hence also of agrarian economic failure and social unrest. Between 1764 and 1791, colonial powers established forest reserves and supporting legislation in Mauritius, Tobago and St Vincent, which in turn can be regarded as the antecedents of the forest planning and protection systems that developed in the Indian sub-continent during the 19th century. This powerful movement of rational resource planning developed further in America under the leadership of Gifford Pinchot, who was from 1905–1910 the head of the US Forest Service (established in 1905 under President Theodore Roosevelt’s patronage). Pinchot understood the importance of public discussion in shaping public attitudes and social values, and he conducted massive publicity campaigns (including debates with opponents such as John Muir, who were deeply opposed to the commercialization of nature) to promote and direct a national discussion on natural resource management issues (Balogh, 2002). These debates helped embed in American society the value that natural resources should be managed for the greatest good for the greatest number in the long run; an idea also expressed as the development of natural resources for the many rather than the few (Pinchot, 1910). Pinchot was a key promoter of these ideas but, as outlined above, he was not the first to express them, and the uptake of such values by governments led to the designation of watershed protection forests and forest reserves in Europe, America and many colonial territories from the mid-19th century onwards. Additionally, colonial administrations in Africa established game reserves for the utilitarian purposes of managing ivory (elephant) resources and protecting source populations of game species that could spill out into the wider landscape and provide a source of meat for new settlers (MacKenzie, 1988). The most recent incarnation of this idea can be seen in sub-Saharan Africa, where there is a growing number of private game reserves, some in excess of 10,000 hectares. Most, if not all, are run as commercial operations. These private reserves typically offer an exclusive African experience involving luxury lodges and camps, private guides and trackers to help the guest get the perfect photograph or memorable kill. In some areas,

managing game reserves for hunting represents a viable alternative to cattle ranching that can benefit wildlife, although this option is not one that chimes well with the values of many in the conservation movement. The formation of the United Nations in the aftermath of World War II, and the rise of science-based rational resource management in international development, promoted a re-formulation of these earlier social values, which can be expressed as the notion that healthy ecosystems are necessary to safeguard economic growth, high quality livelihoods and social stability (Ehrlich & Ehrlich, 1992). Subsequently, the concept of biodiversity served to propel this value further up the political agenda, culminating in the 1992 Earth Summit and the associated Convention on Biological Diversity. The function of many older reserves, originally established in response to other conservation values, were often re-stated to align with new conservation agendas relating to the maintenance of genetic reservoirs, protection of ecosystem services and sustainable utilization linked to livelihood development.

2.2.3 State and country parks State and Country Parks were first implemented in response to the open spaces movement (also termed the amenity movement) which arose amongst social reformers concerned with urban poverty and health in industrial cities during the mid-19th century. The catalyst for action in the UK was plans to sell off Hampstead Heath in London for housing in the 1860s. The Heath, a ridge with fine views, was popular with day-trippers from London’s increasingly crowded suburbs and provided them with a vital respite from the dust, smog, fumes and waste that created the notoriously unhealthy living conditions of 19th century industrial cities. Activists in this movement recognized the health and social benefits of countryside recreation and the need to preserve natural areas both inside and on the fringes of cities and make them accessible to all classes of people. These activists were adept at influencing the parliamentary processes and, through organizations such as the Commons Preservation Society, they secured legislation empowering Metropolitan authorities to acquire open spaces to act as an informal ‘countryside’. The social value they articulated was that access to nature and countryside is necessary for

Roots, relevance, aims and values the health and well-being of urban dwellers. This value soon took root in urban planning in west European and east coast American cities, leading to the creation of urban parks (notably Central Park in New York) and prompting the acquisition of land in or close to cities for designation as country parks (UK) or state parks (US), managed primarily for informal outdoor recreation.

2.2.4 Nature monuments and nature reserves A further site-based agenda for protecting natural sites emerged from the older natural history and philosophical societies. By the mid-19th century, natural history had reached craze proportions as a popular pastime and scientific endeavour among certain sections of European society (Allen, 1994). It was considered the ideal form of self-improvement, providing exercise, education and rational amusement. In combination with romanticist sentiments, it made the field excursion and field club an important focal point of European middle class social life. Perhaps not surprisingly, these field clubs mobilized to protect favourite field sites and other natural sites with special cultural, aesthetic and scientific appeal from urban and agricultural threats (and in Germany from clear-felling policies). A call for action was made between 1900 and 1908 by a prominent German forester, Hugo Conwentz, who conducted a series of high-profile lectures in European cities, where he articulated the threats to natural sites and promoted the concept and vision of Naturdenkmal. This consisted of three inter-connected ideas: 1 that the idea of memorial – Denkmal in German and usually applied to anything in commemoration (e.g. eminent persons, works of literature and art, and ancient buildings) – could also be applied to nature; 2 that Naturdenkmal, like great works of art, should be guarded against ruin; 3 that such action had patriotic value, because, ‘by these undertakings, parts of the country at home become better known and more fully appreciated’ (Conwentz, 1909). The concept of Naturdenkmal captured the value that aesthetic and intellectual contemplation of nature is integral to the biological and cultural inheritance of many peoples, and thus that monuments of nature should be guarded from ruin. The idea of protecting places where people could develop a greater

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appreciation of their natural surroundings was simple and powerful, and found a ready support base amongst the middle class citizenry and some governments. Preservationist groups devoted to identifying and acquiring nature monuments were established in France (1901), Holland (1904), Switzerland (1909) and England (1912), while the Prussian and Swedish states appointed government commissioners for this task in 1906 and 1912 respectively. In the US, the 1906 Antiquities Act provided for the creation of ‘national monuments’, including natural sites (Sellars, 1997). The idea also spread to the Netherlands’ Indies (now Indonesia), where the government designated 24 Natuurmonumenten between 1914 and 1924 for reasons as varied as protection of botanical, faunal and geological features, beautiful panoramas, the Javan rhino, a memorial for an 18th century naturalist and a sacred fig tree (Dammerman, 1929; Jepson & Whittaker, 2002a). Unfortunately, from a conservation perspective, the nationalistic values promoted by Naturdenkmal were appropriated to a degree by right-wing and nationalist politicians in the first half of the 20th century, leading to a distinct ‘down-playing’ of this value in modern conservation propaganda. This trend of politicization was less evident in the UK, where the designation ‘nature reserve’ was preferred to the grander term ‘nature monument’.

2.2.5 Wildlife sanctuaries and refuges Hunting and natural history were two of the great passions of the Victorian age which brought the metropolitan elite into contact with nature at home, and with the frontier landscapes of Africa and the American West. These passions cut across society and led to politically influential citizens witnessing at first hand the disastrous impacts of cultivation, market hunting and resource extraction on species and landscapes (Jepson & Whittaker, 2002a). Arguably the most influential conservationist of all time was the 26th President of the United States of America, Theodore (Teddy) Roosevelt – an avid hunter whose vision for conservation was largely moulded by the excesses he perceived in many of his fellow big game and market hunters. Roosevelt is personally responsible for articulating two of the core values of the modern conservation movement (Table 2.1): first, that humanity has a moral responsibility to save

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Social values and conservation biogeography

threatened life forms; and second, that needlessly slaughtering wildlife is cruel and barbaric. Roosevelt was also a founder member in 1887 of the elite Boone and Crockett Club (B&CC), the goals of which were to create sanctuaries and refuges where wildlife could survive the onslaught of human expansion into frontier lands and to change the culture of hunting to value the noble qualities of the chase rather than the number of animals killed. The activities of Roosevelt and his colleagues in the B&CC were mirrored by European big game hunters, most notably in the formation of the Society for Preservation of the Wild Fauna of the Empire (SPWFE) founded in London in 1903 – still in existence as Fauna & Flora International (FFI). Like the B&CC, the members of the SPWFE were men of great eminence, including the Secretary of State for the Colonies, colonial administrators, hunters and other experts on game in Africa (Prendergast & Adams, 2003). The SPWFE shared the same concerns as the B&CC, but its focus was on declining game populations in the colonies, and in particular Africa. In the early decades of the 20th century, the two organizations joined forces to create the first international conservation treaty – the 1933 London Convention on African Wildlife – which created the legal basis and political clout to establish a network of parks to protect Africa’s game species. These networks established wildlife sanctuaries through a combination of informal lobbying and highlevel representations, and members of each organization held the executive power to create wildlife refuges. For example, as president, Roosevelt designated 59 wildlife refuges between 1909 and 1914, and European colonial governors designated sanctuaries in West Africa, India, Malaya and Indonesia (Jepson & Whittaker, 2002a).

2.2.6 Wilderness areas In the aftermath of the American War of Independence (1775–1783) US politicians sought to create a distinct culture as a mark of ‘true’ nationhood that would both complete and justify the American Revolution. Intellectuals of the time searched for something uniquely ‘American’ to ‘transform embarrassed provincials into proud and confident citizens’ (Nash, 1982, p. 69). A key source and inspiration in this quest for cultural autonomy was the American

landscape. Painters and poets of the Hudson River School (ca. 1825–1890) and the transcendentalist writers (ca. 1830–1870) were inspired by European romanticism to create an intellectual movement that identified wilderness as a basic ingredient of American culture. The philosophical roots of wilderness reserves can be traced back to Henry David Thoreau, who, in 1862, declared that ‘In wildness is the preservation of the world’ (cited in Cronon, 1996, p. 7). Thoreau was a leading light of the transcendentalist movement and rejected the doctrines of established religions in the belief that life was about the search for an ideal spiritual state that ‘transcends’ the physical. Thoreau believed that such a state could be more easily realized in the majestic and awe-inspiring monuments of nature that abounded in the New World. These natural places, in essence, reflected a higher transcendent truth. The motivation to preserve wilderness in parks and reserves probably emerged among the urban political and business elites of New York and other East coast cities and was motivated as much by commercial as spiritual or cultural interests. Since the 1820s, an American ‘grand tour’, visiting picturesque landscapes and supported by imagery and writings exploring their sublime qualities, had been popular among educated urban people. The first officially designated wilderness area was probably Yosemite, deeded by the US government to the state of California in 1864 as a ‘wildland park’ (Cronon, 1996). This was shortly followed by the designation of Yellowstone National Park in 1872, a move enthusiastically supported by railroad interests that saw the area’s potential as a profitable vacation resort and were keen for a nationally sanctioned reservation that would keep speculators and squatters out of the area (Nash, 1982). It took John Muir, the Scottish-born naturalist, to translate the attitudes of the transcendentalists towards the American wilderness into a dynamic social movement with the power to wield real political influence. Muir founded the Sierra Club in 1892, with the goal of establishing Glacier and Mount Rainier national parks, and saving California’s last fragments of coastal redwoods. He gained national prominence through the campaign to stop the damming of the Tuolumne River in the Hetch Hetchy valley of the Yosemite National Park (designated in 1890). The campaign, which ran

Roots, relevance, aims and values from 1908 to 1913, pitted Muir and his followers against Pinchot and his wise-use philosophy. Muir and the ‘preservationists’ adopted a deeply religious rhetoric and imagery in support of their cause. This reinforced the ethos of the 19th century cultural movement and gave shape and popular meaning to a core social value, which can be summarized by the notion that landscapes evoking wilderness should be preserved as a benchmark from which to assess urban/ industrial modernity and as places for spiritual, aesthetic and physical exploration and rejuvenation. The Sierra Club, together with Aldo Leopold’s Wilderness Society (founded in 1935), helped define wilderness areas as a significant part of the American heritage. Their efforts culminated with the signing of the US Wilderness Act in 1964 by President Lyndon Johnson. At a stroke this led to the designation of 37,000 km2 of national forest as wilderness areas, with a wilderness defined in the Act as ‘… an area where the earth and its community of life are untrammelled by man, where man himself is a visitor who does not remain’. As America’s overseas influence increased during the latter half of the 20th century, notions of wilderness strongly influenced the identification and designation of protected areas around the world. Strict nature reserves have some similarities with wilderness areas, although their origins are somewhat different. This category of reserve was invented by colonial powers negotiating the 1933 convention on African wildlife to accommodate Belgian and French views that tourism (as promoted by the British and US) was incompatible with preserving nature in a natural state. As such, this type of reserve is clearly closely related to the underlying philosophy of wilderness reserves as they were originally conceived.

2.2.7 National parks The term ‘national park’ is probably the most widely recognized category of protected area, yet the underlying purpose is not clearly evident from the term itself. This may be because national parks often appear to represent the merger of conservation goals with efforts to create or reinvigorate a sense of national identity (Jepson & Ladle, 2010). Many national parks started off as wildlife sanctuaries or nature monuments and were part of a conscious strategy of ‘nation building’

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adopted by many governments, especially in newly democratic nations. This helps explain why national parks vary so much between countries in size, scope, the nature(s) they conserve and their value for biodiversity conservation. For example, in the USA the goal was to use national parks to capture the grandeur and sense of a pristine nature present in landscapes such as the Rocky Mountains. The creation of Yosemite, Glacier and other national parks, after intense lobbying by the Sierra Club and similar organizations, was instrumental in building national pride and cultural identity in a country still trying to throw off the yoke of British colonialism. In contrast, the ‘political’ purpose of national parks in the UK might be regarded as part of a project to help redefine and renew national identity after the collapse of empire. Accordingly, these parks were initially created in cultural landscapes close to urban centres, such as the English Lake District and Yorkshire Dales. The national parks model has also been exported to the developing world. However, once again the manifestation of the parks has been altered to fit with the political ideals of countries struggling to cement their position in an increasingly globalized world. For example, in countries such as Indonesia and Madagascar, national parks have been used as important symbols of governmental commitment to the international agenda of reducing biodiversity loss, promoting sustainable development and empowering local communities (Jepson & Ladle, 2010). Conservation NGOs have strongly endorsed and promoted the national parks concept in such developing countries because these typically cover larger areas than other reserve types, and also because governments may feel a greater duty to protect and manage them. Furthermore, the governments of the developing countries gain a tangible benefit from designating new national parks, because such acts project a favourable international image, attract international funding and may have the added bonus of strengthening central state control over remote areas. National parks certainly appear to have an enduring appeal to ambitious politicians and they are still growing in number, especially across the developing world. As recently as 2002, President Omar Bongo Ondimba of Gabon pledged to create a network of 13 national parks comprising ten per cent of the country’s land cover.

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Social values and conservation biogeography

2.2.8 Community conservation areas It is now widely recognized that many animal- and plant-rich landscapes are the result of traditional human interactions with their environments. Often, however, traditional practices and forms of landscape management have been eroded either by authoritarian states, markets, environmental degradation or a combination of all three. Since the late 1980s, there has been a concerted effort from the international community towards the rebuilding or reinvigorating of community-level institutions with the authority and capacity to manage their natural resources in a sustainable manner, under the name of Community-Based Natural Resource Management (CBNRM). Tsing et al. (1999) identify three premises of CBNRM programs: first, that local populations have a greater interest in the sustainable use of resources than the state or other potentially interested parties; second, that local communities are more knowledgeable about local ecological processes and practices than are other interest groups; and finally, that such communities are better placed to effectively manage their own resources. One consequence of CBNRM and similar initiatives has been the creation of Community Conserved Areas (CCAs) where local people ‘own’ and manage wildlife rich sites and landscapes. One of the best examples of this is the ‘Communal Conservancy’ model developed in Namibia and now being adopted in other African countries. CCAs are founded on the belief that if communities have exclusive use rights of wildlife resources they will manage the ‘resource’ sustainably. In 1996, Namibia introduced the Nature Conservation Amendment Act (GRN, 1996), giving conditional use rights over wildlife to communities in communal areas. Under the terms of the Act, communities are required to form a local management institution (the conservancy) comprising a committee drawn from the community, draw up a constitution and management goals, develop approaches and monitoring protocols and create a local mandate for conservancy management staff to operate (Massyn, 2007). To help in the process, external conservation groups such as the World Wildlife Fund (WWF) and the independent policy research Institute (IIED) provide technical input, training and general support. The revenue in these areas is almost exclusively generated through safari-tourism and trophy hunting, so the community has a strong vested interest in

maintaining natural resources and key species at healthy levels. This processes of devolving authority over wildlife and tourism to local communities seems to be working well, at least in Namibia. Populations of cheetah, leopard, wild dog, springbok and zebra have been increasing on private land since the 1960s (Barnes & de Jager, 1996), and the communal conservancy model has created hundreds of jobs as field officers, community game guards, community resource monitors and as office staff. In 2006 it was estimated that approximately US$2.4 million ‘new money’ had flowed into conservancy areas in Namibia since their creation (Jones & Barnes, 2006).

2. 3 R ES ER V E DES I GN AT I ON S FR OM I N T ER N AT I ON AL CON V EN T I ON S Perhaps the main driver of the global increase in protected area numbers (and the associated proliferation of terminology) has been the creation of legal frameworks for conservation at the level of individual states or through supra-national bodies such as the United Nations or the European Commission. A good example of the former is the National Parks and Access to the Countryside Act passed by the UK government in 1949, which resulted in the creation of a protected area category known as Sites of Special Scientific Interest (SSSIs) and the formation of a government agency (The Nature Conservancy Council) responsible for their designation. Land could be designated as an SSSI irrespective of whether it was in an existing protected area or not, thereby creating a legislative building block for a more extensive and ‘joined-up’ site-based conservation strategy. N.B.: The SSSI designation is not solely concerned with ‘biodiversity’ attributes and, for example, also covers sites of importance for geological features of interest. National legislation has been vitally important in the creation of new protected areas, but in the last three decades has often been created in response to countries and multinational bodies (e.g. The European Commission) signing up to big global agreements and conservation conventions. It is these international agreements that have done the most to promote the recent growth of protected areas (Figure 2.1) and to provide a framework for their designation. Perhaps the most important such agreement to date has been the Convention on Biological Diversity,

Roots, relevance, aims and values or CBD, which was signed at the Earth Summit in Rio de Janeiro in 1992. One of the key treaty commitments of states that signed up to the CBD was to select, establish and manage a network of protected areas as outlined in Articles 8a and 8b of the convention: • CBD Article 8a. Establish a system of protected areas or areas where special measures need to be taken to conserve biological diversity. • CBD Article 8b. Develop, where necessary, guidelines for the selection, establishment and management of protected areas or areas where special measures need to be taken to conserve biological diversity. Like most international conventions, the CBD is not legally binding in a strict sense, and each individual state needed to draft or redraft legislation in order for the convention to take effect. As might be expected, the translation, transposition, and integration of international law into national law is complex and timeconsuming and has allowed for considerable flexibility in how protected areas networks have been developed and designated in different countries. The CBD is just one of a number of international conventions that require states to identify, designate and create protected areas. For example, the Convention Concerning the Protection of World Cultural and Natural Heritage adopted by UNESCO in 1972 allows member countries to propose sites for addition to the World Heritage list. Many countries proposed their most spectacular national parks, e.g. the Serengeti (Kenya), Ayers Rock or Uluru (Australia) and Yosemite (US). In such cases, the designation merely overlays the national park status. In other cases, world heritage status proposals have been developed as a tool to conserve a heavily used landscape by mobilizing the local pride and sense of international responsibility attached to the designation to strengthen existing planning laws and prompt heritage-friendly development. An example of this approach is the Jurassic Coast World Heritage site in Dorset, England, which covers a scenic coastal landscape and which has been designated largely for its landforms and geology but which has many small nature reserves ‘nested’ within it. Similarly, the European Union’s Birds Directive (1979) and Habitats Directive (1992) require that member states identify and designate Special Protection Areas (SPAs) and Special Areas of Conservation that together form a European network of protected sites called Natura 2000. These may be sites managed (or at least legally designated) as reserves, those holding other

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designations (e.g. SSSI), or sites that have not previously been designated under any legal framework.

2. 4 AN I N T ER N AT I ON AL S Y S T EM FOR CAT EGOR I Z I N G PR OT ECT ED AR EAS Protected areas are generally no longer thought of as embodying specific social values. This is possibly because they have become an integral part of the global system of environmental book-keeping that is an essential component of international conservation conventions such as the CBD. However, in order to fit comfortably into a global accounting system, a standardized classification system based on scientific principles is required. The first attempt at such a classification scheme was developed in 1978 in a joint project between the International Union for the Conservation of Nature and Natural Resources (IUCN) Commission on National Parks and the World Commission on Protected Areas) (IUCN, 2003). By 1994, and after several iterations, this framework had stabilized into six categories of protected areas distinguished by their primary management objectives (Table 2.2; IUCN, 1994). Under the IUCN system, a protected area is defined as: ‘An area of land or sea especially dedicated to the protection and maintenance of biological diversity, and of natural and associated cultural resources, and managed through legal or other effective means’ (IUCN, 1994). Significantly, under this scheme, a protected area is designated to the IUCN classification which best reflects its management aim without reference to its legal title (Fitzsimmons & Wescott, 2004). Of course, grouping the entire world’s protected areas into six management categories (whatever their origins and motivating objectives) inevitably brings with it some limitations, but this has proved useful not only for monitoring purposes but also in providing a framework for reserve planners in developing or adding to protected area systems. For example, the system was adopted and applied in the development of a complete protected area system for the Canary Islands, leading to the designation of approximately 40 per cent of the land surface area of the archipelago and providing a framework not only for strict protection but also for integrated conservation and development projects (see Martín-Esquivel et al., 1995; Whittaker & Fernández-Palacios, 2007, Chapter 12).

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Social values and conservation biogeography

Table 2.2 IUCN protected area categories as defined in the revised Guidelines for Applying Protected Area Management Categories (Dudley, 2008), which are essentially the same as, but reworded from those previously in use (see IUCN, 1994). The revised guidelines also offer a redefinition for protected areas in general, as follows: ‘A clearly defined geographical space, recognized, dedicated and managed, through legal or other effective means, to achieve the long-term conservation of nature with associated ecosystem services and cultural values’ (Dudley, 2008; and compare with IUCN 1994 definition given in text). Designation

Name

Definition

Category 1a

Strict nature reserve

Strictly protected areas set aside to protect biodiversity and also possibly geological/geomorphological features, where human visitation, use and impacts are strictly controlled and limited to ensure protection of the conservation values. Such protected areas can serve as indispensable reference areas for scientific research and monitoring.

Category 1b

Wilderness area

Usually large unmodified or slightly modified areas, retaining their natural character and influence without permanent or significant human habitation, which are protected and managed so as to preserve their natural condition.

Category 2

National park

Large natural or near-natural areas set aside to protect large-scale ecological processes, along with the complement of species and ecosystems characteristic of the area, which also provide a foundation for environmentally and culturally compatible spiritual, scientific, educational, recreational and visitor opportunities.

Category 3

Natural monument

Areas set aside to protect a specific natural monument, which can be a landform, sea mount, submarine cavern, a geological feature such as a cave or even a living feature such as an ancient grove. They are generally quite small protected areas and often have high visitor value.

Category 4

Habitat/species management area

Areas that aim to protect particular species or habitats and management reflects this priority. Many Category 4 protected areas will need regular active interventions to address the requirements of particular species or to maintain habitats, but this is not a requirement of the category.

Category 5

Protected landscape/ seascape

A protected area where the interaction of people and nature over time has produced an area of distinct character with significant ecological, biological, cultural and scenic value; and where safeguarding the integrity of this interaction is vital to protecting and sustaining the area and its associated nature conservation and other values.

Category 6

Protected area with sustainable use of natural resources (previously termed ‘Managed resource protected area’)

Areas set aside to conserve ecosystems and habitats, together with associated cultural values and traditional natural resource management systems. They are generally large, with most of the area in a natural condition, where a proportion is under sustainable natural resource management and where low-level non-industrial use of natural resources compatible with nature conservation is seen as one of the main aims of the area.

Roots, relevance, aims and values As mentioned above, one of the strengths of the IUCN system is that it appears to provide a clear and unambiguous measure of conservation progress that can be easily accounted at regional, country and global levels. However, the system has had several critics. Some commentators consider using management goals to be a rather ‘unidimensional’ indicator of progress towards biodiversity targets (Chape et al., 2005). In response to such accusations, Boitani et al. (2008) have recently suggested revising the IUCN system to make the conservation outcomes explicit. Under this new system, biodiversity outcomes would be measured through a range of biophysical metrics of species and

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ecosystems such as phylogenetic uniqueness, vulnerability, irreplaceability, richness, and ecological integrity. This new proposal may have strengths, but it still downplays the fact (as we have demonstrated above) that protected areas are not simply conservation tools – they are also value-laden institutions that simultaneously reflect, emphasize and embed specific human values about nature. Interestingly, because the IUCN classification system reclassified many protected area types that were originally created in the cause of specific values, most of the IUCN categories broadly equate to different and discrete sets of foundational conservation values (see Table 2.3).

Table 2.3 A social purpose classification of protected areas based on original protected area values. Original protected area type

Equivalent IUCN (1994) category

Places evoking wilderness should be preserved as benchmarks to assess urban/industrial modernity and for spiritual, aesthetic and physical exploration and rejuvenation

Wilderness area

Ib

Humanity has a moral responsibility to ensure that its actions do not knowingly cause the extinction of species

Wildlife sanctuary/ refuge

IV

Aesthetic and intellectual contemplation of nature is integral to the cultural and scientific inheritance of many peoples and monuments of nature should be protected

Naturdenkmal/ nature monument

III

Benchmark/representative sites are required for the study of natural systems

Nature reserve

Ia

Access to nature and countryside is necessary for the health and well-being of urban-dwellers

Urban/country/ state park

Not included, but see National Park

Natural resources should be managed to support livelihoods of settlers/local people; natural resources should be managed for the greatest good for the greatest number in the long run

Forest reserves & game reserves

VI

Healthy ecosystems are necessary to safeguard economic growth, high-quality livelihoods and social stability

Watershed protection forest

Not included, but see National park

Places that symbolize the above conservation values and their associated social practices can help create or reassert national identities

National park

III

Sensitive management of cultural landscapes evoking beauty and heritage will bring cultural, economic, and conservation benefits

Landscape protection area (various national terms)

V

Conservation values

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2. 5 S O C I AL V AL UE S AND CO N S E R V AT I ON P R AC T I C E In addition to underpinning protected area designation and classification, social values also strongly influence management within reserves and the wider landscape matrix. Perhaps the most fundamental level at which values can be seen to influence management is the decision of what is a justifiable target for conservation. Callicott et al. (1999) suggest that underlying the diverse goals and management objectives of conservation are two divergent philosophies: first, to manage for function/process (the functionalist perspective); or, second, to manage for a particular composition of species and communities (the compositionalist perspective). This theme is developed in more detail in Chapter 3, but it should be noted that the two rationales are not always separable in practice. For example, small protected areas may be incapable of functioning as fully independent ecological systems and may require intensive management to maintain particular species or assemblages and to suppress natural fluctuations in population size or community composition. Another fundamental issue in conservation management is the notion of ‘naturalness’ and to what extent it dictates conservation goals and management. Conservation, as the name implies, has always had a strong emphasis on preserving nature in its ‘original’ form. However, the word ‘original’ can mean radically different things, depending on the temporal perspective that is adopted. Clearly, the ecological baseline that is chosen will strongly influence how a reserve is perceived and managed. In Europe, and especially in the UK, a preindustrial baseline with its associated intensive management is the norm, in North America and the developing world a pre-human ecology is more frequently sought (Sutherland, 2002). These issues are fundamental to conservation and are explored in detail in the next chapter. Within this chapter, we now move on to a brief consideration of the issue of non-indigenous species. Attitudes to non-natives are likely to vary depending on whether a functionalist or compositionalist perspective and value set is adopted. From a compositionalist perspective, non-natives may be viewed as ‘unnatural’ elements that should be eradicated. From a (more pragmatic) functionalist perspective, non-native species may have become an essential component of ecosystem function by taking on the role of a species that is

no longer present. The biogeography of invasive species is explored in detail in Chapter 9, so here we restrict ourselves to a discussion of the divergent attitudes that non-native and invasive species provoke in the conservation community.

2.5.1 Attitudes to non-native species It is well known that some exotic species can have a negative impact on biodiversity and ecosystem function, and that they are a major cause of extinction as well as the driver of huge economic losses (Perrings et al., 2005). It is generally accepted that the magnitude and rate of today’s biological invasions is unprecedented and that present-day anthropogenic introductions differ from ‘natural’ invasions in the increased spatial scale over which organisms are being moved and the greatly increased frequency with which such events occur (Ricciardi, 2007). Many biogeographers are concerned that area-specific distinctiveness will be lost in a process of biotic homogenization (Olden et al., 2004; Rooney et al., 2007; Chapter 9). Nevertheless, the debate over alien species and their management has been critiqued on the grounds that not all exotic species are harmful; many are useful to humans and some, at least, have ‘positive’ biodiversity impacts (Kendle & Rose, 2000). Furthermore, several authors object to the ideology suggested by the emotive and value-laden terms that are commonly used in the literature of non-native species (Theodoropoulos, 2003; Brown & Sax, 2004). Finally, some authors have questioned whether invasions are a cause of extinction and ecological ‘harm’ (Sagoff, 2005) or merely a consequence of biodiversity loss generated principally by coincidental processes (Gurevitch & Padilla, 2004). Some authors have noted that there are similarities between terms used in invasion biology and those used in relation to human immigration, and this has raised concerns about the underlying ideology of invasion biology. Theodoropoulos (2003, p. 120), for example, states that ‘the foundational concepts and logical structure of invasion biology are identical to those of the discredited ideologies of xenophobia, racism, nationalism, and fascism.’ Similarly, but perhaps less polemically, Brown and Sax (2004) suggest that ‘[t]here seems to be something deep in our biological nature, related to xenophobia toward other humans, which colours our view of alien plants and animals. There is a tendency to treat

Roots, relevance, aims and values foreigners differently from natives: with distrust, dislike even with loathing’ (Brown & Sax, 2004, p. 530). While this direct transposition of biological terms with social critique/ideology may surprise many invasion biologists, the confluence of language is evident. In order to circumvent issues of confusing, ambiguous or emotive terminology, Colautti and MacIssac (2004) suggested that the terminology surrounding ‘invasive’ species could be made more neutral. They proposed a five-stage model in which each stage, labelled numerically, describes an operational filter that allows some species to persist, reproduce and spread while others do not. However, this scheme has not been widely adopted, perhaps because the reader needs anyway to return to a description of each filter and therefore the parsimony and neutrality of a series of numbered stages is lost. Richardson et al. (2000) prefer to distinguish the introduction (transport by humans) and naturalization (survival and reproduction) from invasion, which requires that the introduced species has spread to and reproduced in areas distant from the sites of introduction (Richardson et al., 2000). This distinction is helpful because the process of naturalization is not necessarily seen as ‘bad’. Many favoured exotic species in gardens and farms are naturalized, but do not spread independently from the locations in which they are desired (Kendle & Rose, 2000; Sagoff, 2005). In short, not all exotic species are invasive. Conversely, not all invasive species are exotic, as exemplified by bracken, Pteridium aquilinum which, although native in the UK, is an often unwelcome invader of heathlands and moorlands. Furthermore, many non-native species hold particular utilitarian, cultural or aesthetic roles. For example, the bird of paradise flower (Strelitzia) is native to South Africa, but is the official flower of the city of Los Angeles, California, USA. Nonetheless, the negative connotation of the word ‘invasion’ still remains and is problematic when one considers that ongoing climatic change, ecosystem interactions and anthropogenic management all cause variations in the distribution and abundance of species over time, and in this sense all species’ distributions may be considered the result of a past invasion (Keitt et al., 2004). For example, ice sheets covered large areas of the northern hemisphere land mass during the last glacial maximum, c. 21,000 years ago, and these areas were recolonized or invaded by species that persisted in more southerly refugia. Similarly, volcanic

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islands and other areas that experience large infrequent disturbances recover when species recolonize from surrounding areas or the mainland. A long-term perspective raises the issue that definitions of ‘native’ and ‘non-native’ are ‘based on value judgements associated with a selective time frame, and a selective categorization of which types of humans can legitimately act as modes of dispersal’ (Kendle & Rose, 2000, p. 22). Moreover, as Gurevitch & Padilla (2004) point out, the correlation between extinctions and invasions does not necessarily imply causation, and care needs to be taken to distinguish whether invasive species directly displace native ones, causing local extinction, or whether both invasion and biodiversity loss are a consequence of habitat degradation. Clearly then, it is not the process of invasion itself that is problematic, but the invasion by a species that would not ‘naturally’ be in the area and, furthermore, whose arrival to an area reduces biodiversity and threatens ecological processes, and/or economic resources or livelihoods. Returning to the concept of naturalness, Brown & Sax (2005) caution against what they term the ‘naturalist fallacy’, meaning the assumption that ecosystems prior to invasion must be natural and pristine, and that this is the ecosystem state that should be preserved. As Kendle and Rose (2000, p. 20) point out, this is problematic because ‘it commits us to supporting a flora that reflected a particular environmental and climatic state that cannot continue forever and has probably already changed’. Stability of species composition is especially unlikely if the projections of rapid 21st century climate change made by the global climate science community are borne out, as these changes will generate reshuffling and redistribution of species as they track (or in some cases fail to track) suitable climate space (Chapter 7; Araújo et al., 2004a; Williams & Jackson, 2007). As mentioned above, although the narratives of invasion scientists are not meant to apply to people, they may nevertheless have negative overtones for immigrant communities (Kendle & Rose, 2000). The potential for confusion is exacerbated by the use of the same words in both a societal and a biological context. While biologists may have an unambiguous understanding of the term ‘alien’ as referring to an animal, plant or pathogen, the same terminology is also used with regard to human immigration – specifically the use of the term ‘illegal aliens’ in the USA and other countries. To many people, this issue may seem like

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another manifestation of contemporary society being overanxious about ‘political correctness’. However, particularly at the interface of science and society, scientists need to reflect on how and with what consequences their technical use of language translates into the public domain. One option is to develop distinct terminologies for animals, plants and pathogens and a separate vocabulary for issues relating to immigration, but even so there would probably be little room in such a scheme to consider the notion of exotic plant and animal species having positive effects. An alternative approach, suggested by Kendle and Rose (2000), is to acknowledge the aesthetic, recreational and cultural role of non-indigenous species in cultural landscapes. This attitude would be congruent with those espoused regarding modern cosmopolitan cities which celebrate the diversity of human inhabitants, and the enrichment of society that comes from engaging with multiple cultures. Application of this approach would require the use of neutral terms for non-native species, while only those capable of invading and causing actual economic or ecological losses would be labelled pejoratively. As this line of argument implies, alongside or alternative to changing the terms used there is perhaps a need to re-think individual and institutional attitudes to non-native species. Consider the response among the birdwatching sub-community of ‘twitchers’ in the UK to the arrival of a straggler vagrant individual only very rarely seen in the UK; typically, many enthusiasts travel to the spot as rapidly as possible to gain a sight of the stray before the bird expires or disappears again. Such behaviour suggests that our attitudes to the exotic can, in the right circumstances, be remarkably positive. Also, given sufficient time following their introduction, many exotic species are effectively culturally assimilated to the point that they are generally regarded as native, notwithstanding that they have been introduced by people from perhaps a very distant and different biogeographical region. This again suggests that notions of ‘naturalness’ are in fact somewhat more fluid and less reactionary than some conservation commentators suggest may be the case. In summary, non-indigenous species that are invasive, such as rats, zebra mussels, and water hyacinth, to name a few, have had adverse effects on ecosystem structure and function, and are a legitimate target for conservation intervention in areas set aside for wild

nature. In protected areas, indigenous species will be favoured over exotics, while elsewhere our cultural and agricultural landscapes are often enriched by, and may even depend upon, exotic species. Focusing management interventions more specifically on invasive species in protected areas would allow a more effective and efficient targeting of conservation funds.

2.5.2 Restoration and rewilding The goal of restoration ecology can be broadly described as an attempt to return an ecological system to some historical state. Most conservationists realize the impossibility of ever reaching such a singular goal, so a better, if fuzzier, definition might take the form, ‘an attempt to move a damaged system to an ecological state that is within some acceptable limits relative to a less disturbed system’ (Falk et al., 2006). Once again there is a clear values component to these goals, since conservation managers seeking to ‘restore’ a site may need to make a choice between a range of potential historical states (the last ecological survey, the first ecological survey, pre-industrial, pre-human) and/or the key ecological processes that they seek to influence. This latter issue may be beyond the scope of many restoration projects due to the highly scale-dependent nature of many ecosystem processes and the associated ‘services’ that they provide for humanity. Issues of restoring communities and ecological processes have recently come together under proposals to recreate long-absent assemblages of large mammals that existed before humans first spread into Europe and North America. This conservation strategy, known as ‘rewilding’, is based on research that suggests that these ‘lost’ large mammals may have played essential roles in determining a range of ecosystem processes. The best example of applying the values of rewilding to a real protected area, albeit a small one, is arguably the 5,600 hectare Oostvaardersplassen reserve in the Netherlands. Conservationists have successfully introduced red, fallow and roe deer, Heck cattle and konic ponies, the latter two species as replicates of the extinct auroch and tarpan (Sutherland, 2002). Konics are considered an ancient breed, thought to be very close to Europe’s extinct horses, whilst the Heck cattle were developed in Germany in the 1920s through crossbreeding old breeds of cattle. Wild boar, lynx and wolf

Roots, relevance, aims and values have, so far, not been introduced and probably never will be due to societal resistance. The intriguing idea of reconstituting past large mammal assemblages is not confined to Holland. The Pleistocene Park project underway in Russia is transforming taiga ‘back’ into mammoth tundra steppe through creation of grasslands and the introduction of bison, musk ox, Yakutian horse, hares and marmots (Zimov, 2005). The introduction of predators is planned once populations have established. A number of private land-holders are also engaging with similar ideas; for example, Paul Lister, heir to the MFI retail empire, is pursuing a dream to rewild a Scottish estate with wild boar, elk, bears and wolves (Sidway, 2006). A third variant is the ‘Pleistocene rewilding of North America’ proposal, which involves the introduction of non-native analogues (elephant and camel) for species that became extinct coincidental with human colonization of the Americas 13,000 years ago (Donlan et al., 2005; Donlan, 2007). This suggestion has aroused intense debate, with critics arguing that introducing elephants, camels, cheetahs and lions would be both ecologically and socially unsound (e.g. Smith 2005; Rubenstein et al., 2006). In a less high profile, but no less ambitious project, the World Wildlife Fund (WWF) and the American Prairie Foundation are buying up properties in north central Montana that they eventually hope to combine with adjacent public lands to provide a habitat for nearly the entire suite of Pleistocene North American grassland species (Dinerstein & Irvin, 2005). Rewilding also raises a number of scientific, practical and social concerns that are yet to be fully resolved. First, scientific knowledge of past assemblages is often incomplete, thus making it difficult to establish the baseline conditions for restoration with accuracy (see Chapter 3, Section 3.5.3). Second, without the social consent required to reintroduce large predators back into rewilded systems, it is unlikely that the objective of recreating evolutionary process can be fully realized (Schlaepfer et al., 2005). Third, the poor record of reintroductions, especially of carnivores, suggests that the success of more ambitious rewilding projects is by no means assured. Fourth, many of the rewilded megaherbivores are closely related to domestic forms, significantly increasing risk of disease transmission between wild and domestic stock. Fifth, restoring ecosystem process and viable populations may require reserves that are too large to be realistically accommodated in many parts of the world because of competing land

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uses. Sixth, rewilding may run up against strong social barriers, especially with respect to the reintroduction of predators such as wolves and bears. Notwithstanding these serious issues, the idea is one that engages a great deal of popular interest, capturing the hostility of some but also the imagination of many. Rewilding solutions lying between the relatively enclosed nature of Oostvaardersplassen and the revolutionary ideas of Pleistocene rewilding would seem to offer considerable potential for successful ecosystem restoration. One such example is the recent introduction, as an ecological analogue for extinct Mascarene Island giant tortoise, of captive-bred populations of Aldabran tortoise to Round Island, a small island (255 ha) with noxious weed problems (Griffiths et al., 2009). The idea is that the introduced species may act as an ecosystem engineer, restoring the functionality lost due to the extinction of the indigenous species and thus contributing to efficient biodiversity conservation in the Mascarene Islands. If based on careful case-bycase planning, and experimental testing and trialling, rewilding has some role to play in 21st century conservation.

2. 6 CON CLU DI N G R EMAR K S In this chapter, we have argued that human values are both the motivation for conservation and have always influenced the science and practice of conservation. Values are always changing and it is possible that one day the foundational values of the modern conservation movement will seem outdated and out of tune with global opinion. Moreover, there is also a danger that values will be increasingly overlooked as conservation develops increasingly sophisticated ways to measure and protect nature. For example, the introduction and promotion of the term and concept of ‘biodiversity’ has undoubtedly been hugely successful as a means of raising funds and focusing resources, but is poorly understood by the public (Christie et al., 2005) and is not closely associated with any of the historically important conservation values (but see Section 2.2.2). If the underlying values of nature reserves and conservation initiatives are not clearly stated, there is a danger that the public may lose its appetite for conservation – a process that would see politicians quickly follow suit. As you read through the rest of this book, we would like you to bear in mind that although

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conservation biogeography draws on a base of hard evidence and scientific principles, its overall remit is intrinsically linked to the fundamental values of the conservation movement. Moreover, as human impacts on natural systems continue to broaden and intensify, patterns of species distribution and abundance are increasingly ‘biocultural’ – which is to say that they find explanation in the geography of social values towards nature. Decisions need to be made on what to conserve, where to spend money, and who will ultimately benefit (or lose out) from conservation. In other words, conservation is social and political as much as it is scientific and rational.

F O R DI SC USS I ON 1 How do modern conservation values differ from those apparent at the beginnings of the global conservation movement? 2 Are conservation values universal and does this matter?

3 Is rewilding a useful conservation tool or an unnecessary distraction? 4 How can an understanding of social values be used to make protected areas more relevant to the public? 5 Is ‘wilderness’ a meaningful concept in the 21st century?

S U GGES T ED R EADI N G Bhagwat, S.A. & Rutte, C. (2006) Sacred groves: potential for biodiversity management. Frontiers in Ecology and the Environment, 4, 519–524. Ehrlich, P.R. & Ehrlich, A.H. (1992) The value of biodiversity. Ambio, 21, 219–226. Jepson, P. & Whittaker, R.J. (2002) Histories of protected areas: internationalisation of conservationist values and their adoption in the Netherlands Indies (Indonesia). Environment and History, 8, 129–172. Jepson, P. & Ladle, R. (2010) Conservation: a beginner’s guide. Oneworld, Oxford. Vera, F.W.M. (2000) Grazing ecology and forest history. CABI Publishing, Oxford. Wilson, E.O. (1984) Biophilia. Harvard University Press, Cambridge, MA.

CHAPTER 3 Baselines, patterns and process Lindsey Gillson1, Richard J. Ladle2,3 and Miguel B. Araújo3,4 1

Plant Conservation Unit, Botany Department, University of Cape Town, South Africa School of Geography and the Environment, University of Oxford, Oxford, UK 3 Department of Agricultural Engineering, Federal University of Viçosa, Brazil 4 National Museum of Natural Sciences, CSIC, Madrid, Spain and University of Évora, CIBIO, Évora, Portugal 2

3. 1 I N T R OD UC T I ON If modern conservation practice is characterized by anything in particular, it is the diversity of goals that conservationists have for protected areas (see Chapter 2). Typically, site management may be directed towards one or more of the following: • maximizing biological diversity (biodiversity) or biological integrity; • protecting a particular species or subset of species; • restoring a past ecological community; • maintaining or enhancing ecological (ecosystem) services; • providing a platform for sustainable development (e.g. enhanced tourism resources, sustainable extraction, etc.); • increasing ecosystem health. On-the-ground conservation decisions depend, to a large extent, on the underlying objectives and values of the organizations or individuals involved. These values and objectives are, in their turn, shaped by perceptions of ecosystems on the one hand, as stable balanced entities or, alternatively, as dynamic systems in flux. The former perception is linked to the so-called balance of nature paradigm in ecology, and within conservation it is aligned with a compositionalist perspective, while the latter, flux of nature, paradigm is strongly associated with a greater emphasis on functionalism. In this chapter we examine the debate between these two contrasting viewpoints, with particular focus on

the baselines or benchmarks against which conservation goals are measured.

3. 2 ECOS Y S T EM COMPOS I T I ON AN D FU N CT I ON In groundbreaking papers, Noss (1990) and Callicott et al. (1999) argue that normative concepts within conservation can be grouped into two philosophies/ approaches that they term ‘functionalism’ and ‘compositionalism’. According to this classification, functionalists can be distinguished by the tendency to focus on ecological processes such as nutrient cycling and thermodynamics. Compositionalists, by contrast, derive their world view from ecological biogeography and community ecology, and they view ecosystems as interacting hierarchies of individuals, populations and communities. The importance of the functionalist/compositionalist dichotomy to conservation practice is that these differing perspectives strongly influence the desired end point of a conservation intervention. Seen through the eyes of a functionalist, the goal of conservation is to restore and maintain ecosystem processes. Conversely, a compositionalist approach emphasizes re-creating or maintaining species assemblages that closely resemble past communities, usually those that existed in preindustrial or prehistoric times. A good example is the conservation of heathlands in southern England,

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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which are maintained by mimicking medieval farming practices (Ladle & Jepson, 2010). In reality, most conservationists have both compositionalist and functionalist tendencies and it may be better to think of individuals and organizations as occupying part of a functionalist/compositionalist continuum (Callicott et al., 1999). Neither compositionalist nor functionalist approaches are without drawbacks. A pure functionalist approach may lead to many rare or endangered species being replaced by ‘weedy’ species that occupy the same ‘role’ within the ecosystem with equal or superior facility. In addition, assessments of ecosystem processes are highly scale dependent, and a functionalist approach may simply be ineffective for smaller reserves such as many of those found in abundance in some industrialized European nations. Conversely, a compositionalist approach, by largely neglecting process, may necessitate huge and continuing investment of resources to maintain the desired assemblages. Furthermore, a compositionalist approach is likely to have far less traction with the polity in countries where populations are heavily reliant on natural resource extraction or ecosystem services or both. Williams and Araújo (2002) have argued that the two approaches must be linked for conservation to be successful. This makes sense, because the persistence of species and the sustainability of ecological processes are largely dependent upon one another, even if the two are not wholly interchangeable. Logically, any such linkage should begin from the compositionalist perspective, because the maintenance of ecological processes (including the politically and economically important ‘ecosystem services’) depends upon the conservation of the ecosystem components. A linkage between compositionalist and functionalist approaches will necessarily have an explicit spatial component, with smaller, more intensively managed areas of compositionalist interest being embedded within larger areas or landscapes that are managed and monitored from a functionalist perspective. In Western European countries, this typically takes the form of a semi-agricultural landscape containing a variety of small nature reserves, or semi-natural areas that are managed for a variety of conservation goals, including conservation of rare species and recreation. A further challenge for 21st century conservationists will be maintaining ecological pattern and process in the face of anthropogenic climate change. The predictions of most climate envelope models (e.g. Thomas

et al., 2004), even if highly uncertain (Chapters 4, 7; Araújo et al., 2005b), suggest that it will be difficult to retain the existing composition of species in reserves (especially small ones) and the wider landscape (e.g. Araújo et al., 2004a; Hannah et al., 2007) and that conservationists will have no choice other than to focus on restoring and strengthening ecological process so that natural systems can respond effectively to climate change. Responses may include facilitating movement of species between reserves or, more passively, relying on the evolutionary responses of species in situ. In some respects, this is merely an extreme example of a more general problem for compositionalist approaches to conservation – how best to conserve systems that are intrinsically dynamic?

3. 3 B ALAN CE V ER S U S FLU X In recent decades, there has been something of a paradigm shift in ecology, from an equilibrium or ‘balance of nature’ world view, to one of nature in flux, or not at equilibrium. This shift in perspective has profound implications for the way ecosystems are understood and managed, although whether it truly constitutes a paradigm shift is moot. Such dualities have typically existed in parallel as binary opposites (sensu Eagleton, 1983; cited in Crisci & Katinas, 2009), such that for a while one perspective is central to the discipline, and is privileged, while the opposite member of the pair is marginalized. This representation perhaps better captures the character of such debates than the notion that one view was universally held and then universally discarded in a once-and-for-all paradigm shift. Along these lines, it is almost certainly a mistake to conceptualize ecological systems as either wholly equilibrial or entirely non-equilibrial in nature (e.g. see discussion in Whittaker & Fernández-Palacios, 2007). The equilibrium paradigm dominated ecology for most of the 20th century, but its origins can be traced far back in time to ancient Greek and Judaeo-Christian traditions (Egerton, 1973; Wu & Loucks, 1995). Nature was conceptualized as a stable and unchanging entity, a view that underpinned influential ecological ideas such as the climatic climax, the logistic growth equation and ideas of carrying capacity (Bartels & Norton, 1993). These apparently diverse and unrelated ideas describe vegetation assemblages and populations as

Roots, relevance, aims and values homeostatic systems that respond to disturbance by returning to a pre-determined state, through a predictable series of changes. In the climatic climax, for example, vegetation types can be predicted according to climatic factors such as rainfall and temperature – a view later modified to incorporate finer scale patterns as a function of soil type and geology (Clements, 1916; Tansley, 1939; Whittaker, 1953). Following disturbance, ecosystems would progress through several stages to a defined end point, the climax community. Similarly, the logistic curve describes a population increasing exponentially in response to a constant supply of resources, until a point of inflection where organisms compete for resources (Pearl & Reed, 1920). Competition increases as population size grows, until resources are consumed at the same rate as they are supplied; at this point, birth rate and death rate become equal and the population stabilizes at ecological carrying capacity (Bartels & Norton 1993). The rate of population growth increases to a maximum when population is at half carrying capacity, at which point it begins to decline. At ecological carrying capacity, the rate of change is zero (see Figure 3.1). Carrying capacity and the logistic curve dominated stock management and resource harvesting for much of the 20th century. The way stocking rates were determined was based on carrying capacities of different range types, and wild populations were harvested with the aim of maintaining maximum population growth. Maximum sustainable yield was predicted at half of the ecological carrying capacity, known as the economic carrying capacity (Figure 3.1). The strength of the climax theory is that it captures the idea that climate is indeed a major determinant of vegetation type. At least at the biome scale, climate determines the distribution of deserts, rain forests, savannas and other vegetation types. Similarly, carrying capacities for stocking rates and harvest levels can be partially effective, because resources and reproductive rates are finite, and an upper limit for livestock density or the harvesting of wild populations has sometimes proved valuable in preventing overexploitation and degradation of rangelands. At finer spatial scales, however, the influence of climate is modified at landscape and local scales by topography, hydrology, fire, herbivory, anthropogenic management, other forms of disturbance and interactions and feedbacks between these factors. Furthermore, climate varies on timescales from seasonal and interannual to geological, altering primary productivity,

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Figure 3.1 (a) A graphical representation of the logistic population growth curve and its relationship with maximum sustainable yield (MSY). (b) The inflection point, shown as BMSY (the biomass at which MSY occurs) is the point at which the population replaces itself at the maximum rate. This is the MSY for exploited populations such as fish stocks, or the population density of those pests that are hardest to control. B0 = the average unexploited biomass of the stock (the average ‘carrying capacity’).

and vegetation composition. This, in turn, alters the supply and quality of resources available to animals. Populations respond to this variability through increases and decreases in their rate of population growth as well as changes in distribution. Population size also changes in response to demographic stochasticity, biotic interactions, disease and disturbance events. Feedbacks occur between these many interacting biological and environmental variables; for example, the frequency and intensity of fire will depend on seasonal rainfall, which influences standing biomass and likelihood of ignition (Sousa, 1984). For these reasons, ecological equilibria are thought to be transient (temporally unstable) and scale-specific (spatially constrained) (cf. Whittaker et al., 2001;

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Whittaker & Fernández-Palacios, 2007). Equilibriumbased models in ecology are therefore useful in describing the feedbacks that could theoretically lead to stability, but in reality this stability is elusive, or at least scale-specific; a patch dynamic landscape, for example, may retain the characteristics of stability over large spatial extents or short temporal scales, but will be highly dynamic at fine spatial scales or over long time periods. In parallel to the dominant ideas of balance and equilibrium, some ecologists pursued ideas of ecosystem change and landscape dynamics. As early as 1930, Charles Elton asserted that ‘the balance of nature does not exist and perhaps never has existed’ (Elton, 1930). Elton’s belief arose from his understanding of the complexity and dynamism of biological and environmental variables, leading him to believe that change, rather than stability, was the norm for natural systems. In the following decade, Alexander Watt’s prescient address to the British Ecological Society described dynamic landscapes in southern England in which patches of vegetation underwent cyclical changes of ‘pioneer’, ‘building’, ‘mature’ and ‘degenerate’ phases (Watt, 1947). From his own observations of landscapes, he described seven ecosystems, including dwarf heather (Calluna), grasslands and beech (Fagus) woodlands in which he had observed this pattern, leading him to believe it was of general significance to many ecosystems. In later years, Watt’s ideas proved central to the development of our understanding of many ecosystems. The concepts of minimum dynamic area (Pickett & Thompson, 1978), mosaic cycles (Remmert, 1991), forest gap dynamics (Shugart, 1984) and Holling’s adaptive cycles (Holling et al., 2001) rested on Watt’s thesis of patch dynamics, and his work informed many of the influential papers that heralded the paradigm shift to the ‘new ecology’ of nature in flux (Levin, 1992; Pickett et al., 1992; Wu & Loucks, 1995).

3. 4 U N D E R ST AND I NG E C OS YS TEMS IN FLUX Under the ‘flux of nature’ paradigm, ecosystems are understood to be heterogeneous (patchy) and dynamic (variable over time). They are often influenced by multiple variables, feedbacks and non-linear responses. Furthermore, they are prone to stochastic variations,

or ‘ecological surprises’, that may be driven by environmental or biological processes (Pickett & Ostfeld, 1995). The need to understanding complex, dynamic ecosystems has stimulated a wealth of new ecological theory, and two of the most influential themes discussed here are the Hierarchical Patch Dynamics Paradigm (HPDP) and the literature centred on ideas of resilience and ecological thresholds. The HPDP combines hierarchy theory (which proposes that scale-dependent levels of organization exist in nature) and the patch dynamics perspective, thereby providing a framework for structuring knowledge about complex systems (Wu & Loucks, 1995). Three main concepts are central to the HPDP (Wu, 1999; Wu & David, 2002): First, ecosystems may be considered to be complex systems, because they are spatially and temporally heterogeneous and are composed of many interacting components. Moreover, the interactions between ecosystem components are characterized by various feedbacks, non-linearity and threshold responses. The complex properties and dynamics that scientists observe in ecosystems systems emerge from these interactions and from the exchange of energy and materials from outside the system. Second, complex systems can be considered as a nested arrangement of interacting sub-systems, i.e. the system is composed of discrete but interacting subsystems that are themselves composed of sub-systems. Hierarchy theory provides scientists with a means of organizing information about these complex systems by identifying the systems and sub-systems and their corresponding hierarchical levels. Sub-systems at different levels in the hierarchy are dominated by different processes. Higher levels are larger and are characterized by slower processes; these higher level processes impose constraints on lower levels, whereas lower level processes provide the mechanism by which higher levels emerge. Third, ecosystems vary considerably in time and space. HPDP provides a framework for identifying and describing the constituent systems and sub-systems which generate heterogeneity over time and space. Thus, heterogeneous landscapes can be described by identifying patches – spatially discrete entities whose internal structure or function is significantly different from those of their surroundings. For example, in his seminal paper, Watt described vegetation assemblages in terms of dynamic mosaics of patches at different successional stages (Watt, 1947). Crucially, Watt’s

Roots, relevance, aims and values ideas about patch dynamics were some of the first to capture the link between pattern and process in ecology, because they made explicit the functional link between the pattern of vegetation in a landscape and the ongoing process of plant succession. To summarize, the HPDP is a powerful framework for describing and understanding ecosystems because it links pattern, process and scale in a spatial and temporal hierarchy (O’Neill et al., 1986; Pickett et al., 1987, 1989; Urban et al., 1987). A hierarchical (multi-scale) structure has been described for riverine systems (e.g. Frissell et al. 1986), and savannas (du Toit et al., 2003). Further, spatial hierarchies are being used extensively as frameworks for modelling ecological complexity (e.g. Wu & David, 2002), landscape analysis (e.g. Burnett & Blaschke, 2003) and the effects of climate change and land-cover change on species distribution (e.g. Pearson & Dawson, 2003). In savannas, Coughenour and Ellis (1993) proposed a hierarchical structure for ecological processes that nested small-scale patterns of disturbance (e.g. by fire and herbivory) within a broader spatial framework of climate, topography, geology and hydrology. Palaeoecological evidence from the savannas of Kenya provides strong support for this perspective (Gillson, 2004a; see Figure 3.2). Over the past few decades, rapid progress has been made in the development of theories that describe ecosystems in terms of their resilience, defined as their capacity to absorb disturbance, and their ability to reorganize when a critical threshold is exceeded (Holling, 1973). Ideas of resilience and thresholds provide a framework around which the complex, non-linear behaviour of ecosystems can be explained. Furthermore, resilience theory integrates human influences on ecosystems with the feedbacks between linked environmental and social systems (Berkes et al., 2003). Many ecosystems exhibit threshold behaviour, in which a critical environmental or biological threshold is crossed, causing reorganization and transition to a new quasi-stable state or phase. Such phase transitions (reviewed in Folke et al., 2004) have been observed in coral reefs, where overfishing, eutrophication and bleaching can cause a switch to an algal dominated reef; in freshwater lakes, which can switch from a clear oligotrophic to a turbid eutrophic state because of agricultural run-off; in savannas, which switch between woodland and grassland phases, depending on changes in disturbance by fire and herbivores; and in forests that can switch from evergreen needle-leaved

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to deciduous broad-leaved, depending on climatic variation linked to timing of disturbances. The principle of self-organization is one of the key features of systems that exhibit phase and transition. Transition to a new phase may be precipitated by extreme or external environmental factors, but maintenance of the new state is sustained by processes internal to this phase. Thresholds are crossed when ecological resilience is exceeded, either because environmental or biological change reaches a critical point or because of an unusually intense disturbance such as a hurricane, flood or severe fire. Alternatively, or in combination with environmental or biological change, ecological resilience may be modified because of anthropogenic activities that can affect the ability of systems to absorb disturbance. Sometimes, humans have been shown to increase resilience of favoured landscape elements, while in other cases over-exploitation or mismanagement has led to loss of resilience (Berkes & Folke, 1998; Adger, 2000; Dearing, 2008). Even before the widespread recognition of ecological thresholds, many ecosystems had been described in terms of transitions between two or more quasi-stable ‘phases’. Phase and transition describes the dynamic process by which ecosystems transform between alternative states of organization. Rather than a linear process of successional change, ecosystem behaviour clusters around regions of higher probability space, or domains of attraction. These domains are not necessarily equilibrium points, which is why ‘phase’ is a preferable term to ‘stable state’ or ‘equilibrium’. In rangelands, for example, Westoby et al. (1989) described alternate states (phases), with transitions between these states being driven by combinations of climatic factors, and management actions such as fire or changes in grazing pressure. Similarly, in savanna ecology, two apparently stable phases – woodland and grassland – are known. Relatively rapid transitions occur between these phases, and have been observed at a range of spatial scales. In east Africa, Dublin (Dublin et al., 1990) hypothesized a regional scale transition from open grassland to woodland in response to the dramatic reduction in herbivory that occurred due to the rinderpest pandemic at the end of the 19th century, and a later transition back to a more open savanna in response to growing elephant populations and fire, caused by increasing biomass build-up (for comparable landscape-scale dynamics in Australia, see, e.g. Sharp & Whittaker,

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Macro–scale

Desert Savanna Forest

The limits of the savanna biome are defined by broad-scale climatic patterns Climatic change, fire and arboriculture can all modify the extent of the savanna biome

Regional–Landscape Scale Within the savanna biome, the type of savanna depends on: Topography, hydrology, geology and rainfall Regional changes in herbivore abundance (e.g. because of Rinderpest, hunting for ivory)

Local–scale Within savanna landscapes, patchiness is determined by: Local variations in soil and hydrology Disturbance e.g. fire, herbivores

Micro–scale Within patches tree density depends on micro-climate, selective herbivory and micro-disturbance Plant–plant interactions such as competition and facilitation (represented by arrows) also occur at this scale

Figure 3.2 A hierarchy of processes determine tree density in savannas through their effects on tree dispersal, germination, recruitment and mortality. Lower tiers in the hierarchy emerge from the processes which dominate at each spatial scale. Higher-level systems constrain lower-level processes, while lower-level sub-systems provide the mechanisms and constituent parts from which higher-level systems emerge. Modified from Gillson (2004b).

Roots, relevance, aims and values 2003; Sharp & Bowman, 2004). At smaller spatial scales, grassland/woodland transitions have been observed in east and southern Africa at the sites of abandoned livestock enclosures, where confinement of animals led to enrichment of tree seeds and nutrients (Blackmore et al., 1990). A major conceptual advance on the phase and transition concept, as well as the movement to incorporate the effects of anthropogenic influence on ecosystem dynamics, was developed by a group of ecologists and economists working together, led by Crawford (Buzz) Holling and Lance Gunderson (Holling et al., 2001). Their ‘adaptive cycles’ described ecosystem change as developing through a cycle of four different phases: conservation, release, reorganization/renewal and growth/exploitation. According to this conceptual framework, an ecosystem or socio-ecological system in a relatively stable phase, maintained by internal feedbacks, becomes brittle or over-connected. As a result, internal mechanisms like senescence, or external disturbances like environmental change or a variation in anthropogenic management, will lead to a release phase, where previously stabilizing interactions break down. New interactions form in the reorganization phase, and a new organizational state emerges during the growth phase (Figure 3.3). The importance of the adaptive cycle framework is that it considers not only what happens before and after a transition, but the whole process of building, collapse and reorganization. It can also be used as a framework for integrating anthropogenic and environmental effects, and it is equally relevant to social and socio-ecological systems. Furthermore, like the HPDP, adaptive cycles can be nested hierarchically, thereby providing a framework for understanding how processes interact at different spatial and temporal scales. Adaptive cycles often take decades, centuries or millennia, and it is only by considering the long-term pattern and process of change that they can be identified. A recent paper by Dearing (2008) mapped the millennial-scale patterns of land use, erosion and monsoonal intensity (reflected in speleothem, pollen, magnetic susceptibility and sand content data from lake and alluvial fan sediments) in Yunnan, south-west China, onto the adaptive cycle of conservation, collapse, reorganization and rapid growth (Holling et al., 2001) (Figure 3.4).

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Figure 3.3 Adaptive Cycles. Four distinct phases have been identified: 1 growth or exploitation (r); 2 conservation (K); 3 collapse or release (omega); 4 reorganization (alpha). The adaptive cycle exhibits two major phases (or transitions). The first, often referred to as the foreloop, from r to K, is the slow incremental phase of growth and accumulation. The second, referred to as the backloop, from omega to alpha, is the rapid phase of reorganization leading to renewal. From the Resilience Alliance (www.resalliance.org/570.php).

There were two distinct phases of surface erosion. The first was largely resilient to monsoon intensity and corresponded to landscapes undisturbed by people between 2960 and 1430 cal yr BP. The second period of erosion, from 800 cal yr BP, was strikingly different. In this case, erosional intensity had a positive correlation with monsoonal intensity, indicating a loss of resilience of more open human-dominated landscapes. Interestingly, the loss of resilience was not associated with the initiation of intensive agriculture, but occurred during periods of social upheaval when agricultural lands were abandoned. This allowed rapid erosion from the sides of hills that were now neither covered in vegetation nor buffered by a well-maintained terrace system. Another significant aspect of this process was that the loss of ecosystem resilience appeared to be hysteretic (irreversible), even with reforestation, because the hills became criss-crossed with steep erosional gullies. The study thereby elegantly demonstrated how societal changes and corresponding changes in land use can interact with environmental variables to drive an ecosystem across a threshold of reorganization and into a new phase, itself maintained by emergent properties (Dearing, 2008).

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Baselines, patterns and process

Figure 3.4 Landscape stability in alternative steady states. Superimposed lines show possible non-linear change from a non-degraded ‘steady state’ before 1430 cal yr BP, through a 600-year transition period leading to the modern degraded ‘steady state’ after 800 cal yr BP. T1 and T2 represent likely positions of major thresholds in the system. The dashed arrows from T2 show possible future trajectories of landscape recovery. From Dearing (2008). (See Plate 3.4 for a colour version of this image.)

3. 5 DEF I NI NG AND US I NG B AS ELI N ES From a conservation perspective, ecological baselines are static points in time and space from which ecological information, typically on species composition and abundance (relative or absolute), can be compared with contemporary sites for the purposes of assessing anthropogenic impact/environmental change and informing management decisions, or both. Baselines are frequently derived from historical literature (e.g. published studies or reports) or from palaeoecological investigations, but may also be taken from ecologically similar sites that are not subject to significant levels of human impact (Arcese & Sinclair, 1997). The choice of baseline, and the way in which it is used to inform management, can have a profound influence on the ecology of a protected area. For example, choice of a baseline from the 1970s, preindustrial or pre-human colonization could profoundly change biodiversity targets, plans for reintroduction and the level of resources required to achieve the conservation objectives. Generally speaking, European conservation has tended to favour the use of preindustrial baselines that reflect a deeply rooted yearning for a lost pastoral idyll (see Chapter 2). Conservationists in North America and most of the

rest of the world, perhaps because of the continued existence of large wilderness areas, tend to use prehistoric baselines. Behind these generalizations lies enormous variation in how baselines are chosen and used, as well as an increasingly sophisticated understanding of their power and limitations. Most importantly, conservationists have begun to appreciate the ramifications of the shifting world view of the dynamism of natural ecosystems. 3.5.1 Baselines derived from relict pristine systems Clements (1934, p. 42) defined an ecological relict as a ‘community or fragment of one that has survived some important change, often to become in appearance an integral part of the existing vegetation’. Relict communities can often be found in places that are very isolated or difficult to access, such as cliffs, mountaintops and steep-sided valleys – typically places where livestock grazing is difficult or impossible – and sometimes in refugia that were buffered from past climate changes due to micro-climatic stability. These sites have been extensively used both to assess the impacts of direct or indirect human disturbance and, in some

Roots, relevance, aims and values situations, to offer inspiration for ecological restoration when historical data are missing. Relict sites are also frequently the focus of conservation themselves, as they often contain a high proportion of endemic or endangered fauna and flora. More generally, the analysis of relict or natural sites through phytosociology has played a major role in identifying sites for regional conservation, especially in Europe. Phytosociology is a sub-discipline of plant community ecology that seeks to describe and understand plant species co-occurrences – or, in the words of Ewald (2003, p. 291), it deals with the ‘compositional patterns and gradients at the “grain” of the plant community’. The tools of modern phytosociology are gradient analysis, classification and other multivariate methods used to identify characteristic plant assemblages that can be used as baselines for conservation. In the UK, phytosociological data have been used extensively to identify and assess natural and seminatural sites that qualify as Sites of Special Scientific Interest (SSSIs), one of the key national protected area designations in the UK (NCC, 1989) and a constituent part of the larger Natura 2000 network of protected areas in Europe. The desire to preserve ‘semi-natural’ habitats is interesting and could be interpreted as a desire to protect and restore pre-industrial baselines (as opposed to ‘natural’ habitats that are presumably prehuman), although the UK’s national agency (now agencies) was by no means explicit about this (e.g. NCC, 1989). A good example is the identification and prioritization of woodlands in the UK. Only ancient ‘seminatural woodland’ is considered for conservation (NCC, 1989, p. 73) and this is identified using the National Vegetation Classification (NVC) scheme, a standardized classification protocol for all the vegetation types in the UK. Under the NVC, each broad classification (e.g. woodlands) is divided into communities, sub-communities and sometimes variants based solely on the presence or absence of species. UK woodlands are divided into 25 communities which are further divided into 73 sub-communities (Hall et al., 2001). These act as references against which new sites can be assessed and existing protected sites can be monitored. For instance, Upland Oak Woodland, defined as woodland within the ‘upland region’ of England generally with at least 80 per cent oak or birch in the potential canopy, is classified as W11 or W17, depending on the nature of the field layer (see further discussion of these approaches in Chapter 4, Section 4.5.1).

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Relict communities have also been used as potential reference sites for ecosystem and habitat restoration projects. One example can be seen in attempts to restore some of the ‘lost’ forests of Tenerife (Canary Islands), where scientists are drawing upon data from both palaeoecology and contemporary studies of relict communities to provide a basis for restoration efforts. Before human occupation, the island of Tenerife contained all of the main vegetation types of the Canarian archipelago, from the hot, dry semi-desert scrub of the coastal lowlands to the lush, green laurel forest that is swathed in cloud for long periods every day (FernándezPalacios et al., 2004). These vegetation types occurred in distinct climatic zones or bands stretching from sea level to the top of Mount Teide, the 3,700 m dormant volcano that dominates the Tenerife landscape. As happened on many oceanic islands after occupation, agriculture began to expand from the coast to higher elevations and, in the process, devastated or completely removed several ‘bands’ of typical vegetation. In particular, the thermophilous forest was almost completely destroyed (less than 1 per cent remains), with only a few pockets of juniper (Juniperus spp.) forest clinging on to existence on exposed cliff sites. The European Community has recently funded a pilot conservation project to restore a 53 ha patch of thermophilous woodland on the Teno peninsula (www.tenerife.es/life/). Scientists have based the replanting scheme on phytosociological analysis of tiny remnant fragments of juniper forest in steep cliffs and remote gullies on Tenerife and nearby La Gomera. Amazingly, a recent analysis of pollen from sediments in a dried-up lake in the city of La Laguna suggests that there may also have been a native broad-leaved forest of hornbeam (Carpinus) and oak (Quercus), species now considered as non-native on the archipelago, which was most likely sandwiched somewhere between the laurel forest and the pine zone on Tenerife. These findings challenge prior conceptions of indigenous ecological baselines on the Canary Islands (de Nascimento et al., 2009).

3.5.2 Baselines derived from long-term ecology Current ecological understanding recognizes that most ecosystems are dynamic, being subject to ongoing processes of changing climate and other environmental disturbances, and that many landscapes have also

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been shaped by humans for millennia (Botkin, 1990; Pickett & Ostfeld, 1995; Wu & Loucks, 1995; Knapp, 2003). According to these contemporary ecological ideas, change is inevitable and an understanding of the past can help in predicting future community change. Environmental proxies (including fossil pollen, stable isotopes, macro-fossils and charcoal), as well as historical records, photographs and archaeological data, can reveal how communities have responded to environmental change over thousands of years (e.g. Hunter et al., 1988; Birks, 1996; Landres et al., 1999; Swetnam et al., 1999). Such information can be used to interpret current trends and predict how ecosystems might respond to changing climate and land use in the future (e.g. Delcourt & Delcourt, 1991; Hannah et al., 2002a). Although palaeoecology is traditionally seen as a mainly descriptive science that seeks to understand Quaternary change, the discipline is currently undergoing a renaissance and is becoming increasingly relevant to our understanding of ecosystem dynamics, environmental change and ecosystem management. The synergy between the ecology of flux and the ecology of temporal change is yielding many exciting opportunities for the interpretation – and reinterpretation – of palaeoecological data in terms of thresholds, resilience, phase and transition and adaptive cycles (Carrión et al., 2001; Gillson, 2004b; Dearing, 2008). The palaeoenvironmental record shows that some species and even communities can persist, despite climate change, until an ecological or climatic threshold is crossed (e.g. Theurillat & Guisan, 2001; Von Holle et al., 2003; Nogués-Bravo et al., 2008). Palaeoecological studies can help in evaluating the resilience and inertia of ecosystems and in determining critical thresholds at which dramatic ecological changes occur. Furthermore, palaeoecological evidence also indicates that extinctions are common at times of high climate variability, but that rates of loss vary geographically (Botkin et al., 2007). These findings are of critical interest to policy makers, conservation planners and subsistence users, because rapid change at an ecological threshold provides little time and opportunity for adaptation (Chapin et al., 2004). Comparison of past changes in climate and biome extent can help in predicting the ecological consequences of future climate change. Specifically, the outputs of coupled climate and biome models can be compared against the known distribution of biomes from the palaeo-record, enabling the accuracy of model outputs to be evaluated (Harrison et al., 2002).

Figure 3.5 A diagram illustrating how non-analogue communities can develop over time with changing climate. The light grey oval indicates climatic conditions in a particular area. Dark grey shapes indicate the occurrence of a species in that area. Dotted lines indicate the fundamental niche of three species in terms of climate variables 1 and 2. (a) Species 1 and 2 co-exist in the present climate space but species 3 is absent. (b) Species 1 and 3 co-exist in the future climate space – a new species association. Species 2 would need to evolve new climate tolerance to persist under the future climate of the area. After Williams & Jackson (2007).

However, the palaeoecological record provides multiple illustrations from the past, indicating that species, to a large degree, respond individualistically to climate change. This applies particularly to periods of rapid and dramatic climate change, which are notable in the records for the emergence of communities with no modern analogue (e.g. Bush et al., 2004; Williams & Jackson, 2007). Non-analogue communities arise for three main reasons (Parmesan et al., 2005; and see Figure 3.5). The first is that species differ in their ability to keep pace with climate change, so different communities may arise due to some species rapidly moving into available climate space while some persist for a time in areas that are no longer climatically suitable. Second, species may be able to extend their range individualistically when different combinations of climatic factors arise, corresponding to different dimensions of the fundamental niche (Williams & Jackson, 2007). Third, if competition, dispersal and other biotic factors restrict ranges at the outer latitudinal and elevational limits of a species range, then the fundamental niches may be wider than present or past distribution suggest (Parmesan et al., 2005). Climatic change will therefore affect different members of the same community in differing ways – a property that is consistent with the impermanence of plant and animal communities.

Roots, relevance, aims and values Changes in land use and ‘natural’ disturbance events such as hurricanes, floods, volcanic eruptions, severe fires or other Large Infrequent Disturbances (LIDs), can interact with or even override the ecological consequences of prevailing climatic trends (Turner & Dale, 1998). A good example of how palaeoecology and archaeology have elucidated feedbacks between disturbance, landscape patterns and human behaviour is from Garua and Numundo Islands, Papua New Guinea, where the wide-scale effects of catastrophic volcanic events were overlain by a patchwork of different vegetation responses and human activity at finer spatial scales. Boyd et al. (2005) used a combination of archaeology and phytolith (silica bodies that occur in plants) analysis to compare vegetation responses to tephra deposition at local, sub-regional and regional scales. They discovered that the effects of major eruptions (c. 5900, 3600, 1700 and 1400 cal yr BP) varied spatially, partly because of topographic control on deposition patterns of tephra, and this pattern was overlain in turn by human partitioning of the landscape. When tephra accumulation continued after the initial volcanic eruption, landscapes were abandoned by people and the impacts on the vegetation were severe. In contrast, after a low-impact eruption with little tephra deposition, the forest persisted and human occupation continued. The story was somewhat different on the mainland, where the spatially patchy recovery of the coastal lowland reflected the modification of the landscapes by humans prior to eruptions (Boyd et al., 2005). For example, habitats dominated by grasses and other pioneer species (e.g. gardens and forest clearings) recovered more slowly and took much longer to be recolonized by the local population. This example illustrates how a combination of data derived from long-term ecology and strong conceptual structures can generate real insights into the feedbacks and interactions between human activity and landscape pattern, process and scale.

3.5.3 Rewilding Rewilding has been defined as ‘action at the landscape level with a goal of reducing human control and allowing ecological and evolutionary processes to reassert themselves’ (Klyza, 2001, p. 285). In this context, rewilding projects can be grouped in with

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other functionalist approaches to conservation, wherein priority is given to managing and restoring ecological processes. However, it is functionalism with strong compositionalist tendencies; inspiration for rewilding projects is frequently taken from the palaeoecological literature, and evolutionary and ecological processes are typically restored through the reestablishment of ancient assemblages (or their nearest functional equivalent if the original inhabitants are extinct). For example, one of the architects of Dutch rewilding (described in Chapter 2, Section 2.5.2), Frans Vera, drew heavily on the palaeoecological record to argue that the primeval landscape of lowland Europe was much more open than previously thought, primarily due to the actions of large herbivores that are now extinct or greatly reduced in abundance (Vera, 2000). The Dutch government allowed Vera and his colleagues to produce a public demonstration of this hypothesis at the Oostvaardersplassen reserve, a 5,600 ha polder 50 km from Amsterdam, which was created in 1974– 78. This was land that had been reclaimed, but not cultivated, and where elder and willow scrub had rapidly colonized (Sutherland, 2002). They introduced a dynamic grazing management that involved the introduction of red, fallow and roe deer and Heck cattle and Konic ponies as substitutes for the extinct Auroch and Tarpan (Sutherland, 2002; Jepson & Ladle, 2010). The experiment has been a remarkable success, and by 2005 the cattle and horses were exhibiting ‘natural’ herd social structures and had mostly synchronized breeding. The habitat mosaic of pasture, woodland and wetland created by high herbivore densities has produced some novel and often unexpected conservation outcomes. The most high profile of these are the 60,000 greylag geese (Anser anser) that appear on the site each autumn (Sutherland, 2002) and the first ever breeding in Holland of whitetailed eagle (Haliaeetus albicilla) in 2006 (Jepson & Ladle, 2010). Rewilding with large herbivores may also be an effective strategy for mitigating the negative ecological effects of rapid environmental change. Zimov (2005) has suggested that large herbivores may act as ecological buffers between the changing climate and the constituent ecosystems that they help create. Introducing wild-acting herbivores, therefore, has the potential to create a more flexible form of management – one that is more responsive to climate change than the typically crude human management interventions. However,

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the success of such reintroductions might depend on the successful establishment of wild populations of predators or, alternatively, on maintaining at least some element of management of herbivore populations by humans (van Wieren, 1991). Rewilding clearly offers a potential alternative to the current target-driven, intensively managed approach to conservation that is common in many parts of the world and especially Europe. Its combination of functionalist objectives achieved through recreating past assemblages also offers a flexible approach that has potential to be scaled up to encompass entire landscapes (see also Chapter 2, Section 2.5.2).

3.5.4 The challenge of rapid environmental change One of the potential problems of using baselines derived from palaeoecology, historical records or relict populations as guidelines for restoration and conservation initiatives is that the environment may have changed to such a degree that maintaining historical species assemblages is no longer a viable option. This issue is gaining increasing attention due to the widespread concern with anthropogenic climate change. The most recent prediction of the Intergovernmental Panel on Climate Change (IPCC) is that global temperatures could rise by between 1.1 and 6.4°C (2.0 and 11.5°F) during the 21st century, accompanied by changes in precipitation regimes, extreme conditions and seasonality (IPCC, 2007). Such changes, when combined with ongoing destruction, fragmentation and modification of natural habitats and ecosystems, and the anthropogenic transportation of non-native species, will generate unique suites of environmental conditions, novel ecosystems and novel communities (see e.g. Willis & Bhagwat, 2009). In such circumstances, identifying current or past community compositions as target baselines will become increasingly problematic and impractical (Hannah et al., 2002b). Bush (2002) suggests that the solution to this problem is to concentrate on conserving animal and plant niches rather than identifiable communities, but the scientific challenges involved in determining niche requirements at the species level on the scale required are, of course, not trivial. Further discussion of these challenges follows in Chapter 6 and especially in Chapter 7.

3. 6 ADAPT I V E ECOS Y S T EM MAN AGEMEN T The complexity of ecological systems and the uncertainty with which ecosystem changes can be predicted raises dilemmas for ecosystems managers, who must make decisions when knowledge is imperfect and stakes are high (Funtowicz & Ravetz, 1994; Ravetz & Funtowicz, 1999). The management of ecosystems requires recognition that, for any given ecosystem, there may be a range of possible ecosystem states and an equally wide variety of societal responses to these states (Ravetz & Funtowicz, 1999). This uncertainty, complexity and plurality requires an adaptive approach to ecosystem management, i.e. one that continually monitors and evaluates the outcomes of management interventions and adjusts conservation and management goals in the light of new scientific understanding, inputs from stakeholders or ecological surprises (Grumbine, 1994, 1997; Sabine et al., 2004). Critically, ecosystem managers need to know the position of ecological thresholds so that they can maintain desired states or facilitate beneficial changes through management interventions. In Australian rangelands, Westoby et al. (1989) described an opportunistic management system for rangelands, based on the idea of state and transition (see Section 3.4.1). This management approach required understanding of the processes that drive transitions between phases, along with a classification of known vegetation phases according to whether they are favourable or unfavourable – primarily for livestock owners, but the principle could equally be extended to biodiversity conservation. Favourable transitions, such as a change from saltbush to grass-dominated vegetation with scattered woody plants, could be facilitated by de-stocking, whereas maintaining grazing pressure would be more likely to cause a transition to shrub cover, a phase less favourable for livestock (Westoby et al., 1989). In the Kruger National Park (KNP), South Africa, strategic adaptive ecosystem management (Figure 3.6) is used, with the aim of maintaining natural ecological dynamics (Biggs & Rogers, 2003). KNP ecologists and international collaborators have developed processorientated management goals, based on ecosystem properties, known as Thresholds of Potential Concern (TPC). These thresholds are points along a continuum of ecological or environmental change, at which managers either intervene to guide ecosystem change, or at

Roots, relevance, aims and values

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Figure 3.6 The process of Strategic Adaptive Management, as applied within Kruger National Park by South African National Parks (SANParks – the government department responsible for managing national parks in South Africa), illustrating the inter-linkages between describing the desired state, developing a management plan, implementation, monitoring and review. Source: www.sanparks.org/parks/kruger/conservation/scientific/key_issues/plans/adaptive/pdfs/chapter_03.pdf)

which they decide whether to alter their management objectives or TPCs (Biggs & Rogers, 2003). For example, concern for the impact of elephants on trees and shrubs led to the development of a TPC for elephant management. In essence, if woody vegetation cover drops below 20 per cent of its ‘highest ever value’, managers will control elephant numbers by culling or translocation, or will adjust the TPC. Using

such a TPC raises the question of how dense woody cover was in the past, and how far present-day tree cover deviates from the ‘highest ever’ value as judged over varying reference time frames. Using fossil pollen data and landscape modelling software, Gillson and Duffin (2007) estimated that, at two study sites, woody vegetation cover had fluctuated between 25 per cent and 55 per cent over a 1,400 year

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period, compared to 20–40 per cent in the present day, and between 13 and 19 per cent over a 4,900 year period, compared to 4–36 per cent in the present day. They also found that recent increases in elephant abundance did not appear to have exceeded the resilience of the woody vegetation, since no dramatic decreases in woody vegetation had occurred. Their data demonstrate that the ‘highest ever’ woody vegetation cover is indeed dependent on the length of the temporal record used. A more useful TPC might therefore be developed, based on changes in mean tree abundance over time or changes in the patchiness of tree cover. The study also revealed that the variability and abundance of tree cover was different at the two study sites, highlighting the importance of developing TPCs that are site-specific and sensitive to local conditions (Gillson & Duffin, 2007).

F O R DI SC USS I ON 1 What is the significance of a ‘natural’ habitat to the modern conservation movement? 2 In the light of present-day levels and rates of environmental change, how useful is the information derived from long term ecology for contemporary conservation practice?

3 Are baselines important for conservation approaches based on ecosystem services? 4 Does the widespread occurrence of non-native species mean that compositionalist approaches to conservation that focus on restoring past baselines are no longer viable? 5 How does the size of a protected area influence the type of conservation management (e.g. for function or for composition) that is adopted? S U GGES T ED R EADI N G Agnoletti, M. (2007) The degradation of traditional landscape in a mountain area of Tuscany during the 19th and 20th centuries: implications for biodiversity and sustainable management. Forest Ecology and Management, 249, 5–17. Bayliss-Smith, T., Hviding, E. & Whitmore, T. (2003) Rainforest composition and histories of human disturbance in Solomon Islands. Ambio, 32, 346–352. Bush, M.B. (2002) Distributional change and conservation on the Andean flank: a palaeoecological perspective. Global Ecology and Biogeography, 11, 463–473. Callicott, J.B., Crowder, L.B. & Mumford, K. (1999) Normative concepts in conservation. Conservation Biology, 13, 22–35. Sabine, E., Schreiber, G., Bearlin, A.R., Nicol, S.J. & Todd, C.R. (2004) Adaptive management: a synthesis of current understanding and effective application. Ecological Management and Restoration, 5, 177–182.

PART 2 THE DISTRIBUTION OF DIVERSITY: CHALLENGES AND APPLICATIONS

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

CHAPTER 4 Basic Biogeography: Estimating Biodiversity and Mapping Nature Brett R. Riddle1, Richard J. Ladle2,3, Sara A. Lourie4 and Robert J. Whittaker2 1

School of Life Sciences, University of Nevada, Las Vegas, USA School of Geography and the Environment, University of Oxford, Oxford, UK 3 Department of Agricultural Engineering, Federal University of Viçosa, Brazil 4 Redpath Museum, McGill University, Montreal, Canada 2

4. 1 I N T R OD UC T I ON Are very closely allied species ever separated by a wide interval of country? What physical features determine the boundaries of species and of genera? Do the isothermal lines ever accurately bound the range of species, or are they altogether independent of them? What are the circumstances which render certain rivers and certain mountain ranges the limits of numerous species, while others are not? None of these questions can be satisfactorily answered till we have the range of numerous species accurately determined. (Alfred Russel Wallace, 1852, p. 110)

4.1.1 Our incomplete knowledge of biodiversity We may look at the natural world through many different lenses. In Chapter 3 we divided these perspectives into two broad classes: compositionalist and functionalist. These might also be respectively labelled biogeographical versus ecological or ecosystem approaches. When they are applied for conservation

purposes, both require the identification of nature’s units, which, as will already be clear from Chapter 3 represent human impositions. In the present chapter we tackle both these themes, beginning with the compositionalist/biogeographical approaches – those concerned with describing biodiversity variation geographically. Here our prime focus is typically with the species unit, but we can also be concerned with higher or lower levels in the taxonomic hierarchy. We give most attention to the terrestrial realm, but go on to illustrate how compositionalist and functionalist approaches are applied also in the marine realm. The chapter embraces both foundational approaches within biogeography, such as the endeavour of identifying the world’s major biogeographical regions (a project that was well under way in the 19th century), as well as some of the latest approaches to analysing species interrelationships and present and future distributions. When we are contemplating conservation action, we are often concerned with saving locally or globally unique taxonomic elements, typically species which we perceive to be threatened by anthropogenic extinction. We may also be motivated by the notion of saving valued ecosystem types, some of which may be

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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Basic biogeography: estimating biodiversity and mapping nature

essentially cultural in origin or maintenance (in the UK, chalk grassland, lowland heaths and hay meadows come to mind) or of conserving valuable ecosystem function (e.g. estuarine habitat for wintering wildfowl or habitat connectivity for migratory terrestrial species) (see Chapters 2 and 3). Many of the most fundamental questions in conservation biogeography require knowledge of the geographical distributions (and ecological niche requirements) of individual species. Such knowledge is, of course, essential before we can assess the threats to the viability of their populations in a rapidly changing world (see Chapter 7). Unfortunately, we all too often lack this knowledge – a deficit that has, for reasons outlined below, been called the Wallacean shortfall (Lomolino, 2004). Also, before we can even begin to understand the distributions of organisms, we need to know that they actually exist, and unfortunately our knowledge gap between formally described and yet-to-be-discovered species, referred to as the Linnean shortfall (Raven & Wilson, 1992), is vast. These two knowledge deficits contribute significantly to a third key knowledge shortfall, which is that our grasp of the magnitude of anthropogenic extinctions (past, present and especially future) is also marked by a high degree of uncertainty. There is a strong consensus that the rate of loss is already significantly above background levels of extinction derived from the fossil record, but no one can be quite sure by how much. We therefore begin the present chapter with consideration of these three themes, before moving on to describe other forms and levels of biological organization used in conservation biogeography. In terms of functionalist approaches to mapping nature, we focus particularly on two types: biomes and ecoregions. In essence, these are closely related concepts concerned with delineating and mapping Major Ecosystem Types (METs). These METs, in turn, represent a key way in which humans perceive nature in the terrestrial realm based largely on physiognomic features of the vegetation (e.g. temperate grasslands or temperate deciduous woodland). Species are also constrained to varying degrees in their distribution to particular METs, so there is a broad, albeit imperfect, correspondence between functionalist and compositionalist approaches. Both are foundational to the efforts of those involved in conservation planning, as will be discussed further in Chapters 5, 6 and 7, and in practice many fully developed multi-scalar approaches to the problem of mapping nature involve a mixture of the two.

4.1.2 Why do we map? I soon found that the Amazon, the Rio Negro and the Madeira formed the limits beyond which certain species never passed. Thus there are four districts, the Guiana, the Ecuador, the Peru and the Brazil districts, whose boundaries on one side are determined by the rivers I have mentioned. (Alfred Russel Wallace, 1852, p. 110) As a young naturalist and scientific collector in the tropical forests of the Amazon, Alfred Russel Wallace had already developed a remarkable understanding of the importance of accurate distributional mapping as a basis for whatever could subsequently be learned about biogeographical patterns and processes. For example, carefully recording the geographical distributions of species of monkeys, birds and insects gave him the insight which led him to speculate on the relationship between geographical features and attributes of species affinities and distributions to a degree not realized amongst his contemporaries. Indeed, as the above quote indicates, 24 years before reinforcing and expanding upon Sclater’s (1858) scheme of six great terrestrial zoogeographical regions of the Earth (Wallace, 1876), he was already dividing Amazonia into geographic units according to species’ distributions and features of the Earth associated with limits to their distributions. Of course, Wallace and Sclater also had predecessors who had begun to understand and summarize the non-random nature of species’ distributions on grand scales (Ebach & Goujet, 2006; Lomolino et al., 2010). Commencing largely in the 18th century, and more fully developed by Wallace and his contemporaries, these studies form the foundations for efforts to map the distribution of biodiversity across the Earth. Several persistent themes in biogeography developed in conjunction with distributional mapping at taxonomic scales, ranging from intraspecific to higher taxa, using aggregations from single taxa to entire biotas, and at geographical scales ranging from local to global (Lomolino et al., 2010, their Chapter 2). In illustration, we have picked out three deep-rooted themes foundational to modern conservation biogeography which provide the framework for predicting effects of climate change, invasive species, habitat fragmentation and loss, and other anthropogenicallymediated influences on populations, species and biotas.

The distribution of diversity: challenges and applications 1 Classifying geographical regions based on their biotas. Sclater (1858) and Wallace (1876) would have been unable to construct the six great terrestrial biogeographical regions of the world that are still used today with modifications (Cox, 2001; Kreft & Jetz, 2010) without recognition and analysis of non-random global and regional distributions of birds, mammals, and other groups. Within conservation biogeography, these early representations of global biodiversity became constituent building blocks of the Dasmann (1973) and Udvardy (1975) IUCN Biogeographical Regions framework discussed in Chapter 5. 2 Reconstructing the historical development of lineages and biotas, including their origin, spread, and diversification. Both Darwin (1859) and Wallace (below) argued strongly in favour of a ‘natural’ system of taxonomy that formed the informational basis required to study the geographical distribution of animals and plants: A little consideration will convince us, that no inquiry into the causes and laws which determine the geographical distribution of animals or plants can lead to satisfactory results, unless we have a tolerably accurate knowledge of the affinities of the several species, genera, and families to each other; in other words, we require a natural classification to work upon. (Wallace, 1876, p. 83) Most modern methods of reconstructing biogeographical history rely on either phylogenetic methods of reconstructing ‘natural classifications’ of taxa based on ‘descent with modification’ (Hennig, 1966) or some form of genetically-based population similarity analysis, with the mapping of these relationships onto geography (Avise, 2000; Riddle et al., 2008). The use of genetic data to investigate the geographical patterns of historical and ongoing connections between populations within a single species or several closely-related species is called phylogeography (Avise, 2000, 2009; Riddle & Hafner, 2006), and this emerging sub-field is generally recognized as one of the most important recent advances in biogeography. Phylogeography has also spun off several newer approaches, including landscape genetics and phylochronology (the idea of focusing on change in genetic diversity of populations in given localities over time; Hadly et al., 2004). Applications of phylogenetic and population genetic

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methods in conservation biogeography include those that incorporate a phylogenetic diversity metric (Faith et al., 2004b), and a variety of genetic approaches to evaluate the distinctiveness of populations and species (Avise & Hamrick, 1996; Frankham et al., 2002), patterns of endemism and biodiversity hotspots (Verboom et al., 2009). These metrics have considerable potential for incorporation into protected area planning processes (Chapters 5 and 6). 3 Explaining the differences in numbers as well as types of species among geographical areas and along geographical gradients, including patterns related to area, isolation, latitude, elevation and depth. One of the most pervasive patterns in nature, the species–area relationship, was recognized rather soon after geographically representative natural history specimens began to accumulate from global explorations (Forster, 1778) and regional collections (De Candolle, 1855; Watson, 1859). Formal mathematical analyses were initiated early in the 20th century (Arrhenius, 1921; Gleason, 1922). A long history of theory and modelling in conservation biogeography has developed, based on the predicted ‘meltdown’ in species diversity following habitat fragmentation and loss and the consequent impact on species–area relationships (Chapter 8). Several other themes in conservation biogeography have relied heavily on knowledge of non-random distributions of species richness and species types across geography; examples include the grand clines of diversity from poles to the equator and the concentration of endemic species in biodiversity hotspots (Chapter 5; Myers et al., 2000). While the value of mapping biodiversity for strategic conservation planning purposes is clear, the reality is that there are still many gaps in our knowledge of what is out there, how it is distributed and the extent to which it is threatened with extinction. These are the topics covered in the following section.

4. 2 T H R EE K N OW LEDGE S H OR T FALLS 4.2.1 The Linnean shortfall The Linnean shortfall is named after the Swedish naturalist Karl von Linne (1707–1778), better known through his ‘Latinized’ sobriquet of Carolus Linnaeus. He was the creator of the system of Latin binomials and is widely regarded as the father of modern

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taxonomy. The shortfall refers to the enormous discrepancy between the number of species that have been formally described by taxonomists (around 1.7 million at the last count), and the number of species that are thought to exist – somewhere between 3 and 100 million (excluding bacteria and viruses). In other words, it is the disparity between the number of described species and the total number of species in existence (Raven & Wilson, 1992; Lomolino, 2004; this definition from Lomolino et al., 2010). The shortfall is a problematic issue for conservation biogeography, because it can add considerable ‘noise’ to any attempt to map and compare biodiversity. The depth and breadth of the Linnean shortfall is clearly illustrated by the number of new species that are still being discovered – almost at a weekly rate in some parts of the world. These discoveries are not just restricted to insects and other small and inconspicuous inhabitants of tropical rain forests, although these fauna do represent the bulk of the Linnean shortfall. For example, 11 of 80 extant species of cetaceans (whales and porpoises) were discovered only in the 20th century, one as recently as 1991 (Raven & Wilson, 1992). Higher taxonomic groups are also still being discovered. For example, Raven and Wilson, in their 1992 paper, note that three new families of flowering plants were discovered in Central America and southern Mexico within the previous decade, while two new phyla were described in the last two decades of the 20th century (Lomolino et al., 2010). However, it is fair to state that most big and well-studied groups, such as birds and mammals, have been reasonably well described, while the shortfall is greatest in smaller, less charismatic taxa such as fungi and, especially, many types of arthropods. Additions to the global list of species come about both by new collections of voucher material and by re-inspection of material previously gathered and residing in museums around the world. In addition to the traditional systematic taxonomy (whereby specimens are classified based on morphology into different species), increasingly, as we will discuss later, genetic data are allowing the identification of morphologically cryptic species. A nice example comes from studies of Nesotes beetles in the Canary Islands, where genetic analyses showed that what had been thought to be a single species Nesotes fusculus, occurring on the islands of Tenerife, La Gomera and Gran Canaria, really represented a paraphyletic group (see Glossary), in which the fusculus phenotype previously recognized

as a monophyletic species (ditto) had evolved independently in the xeric coastal zones of the three islands (Rees et al., 2001). Such analyses have contributed to the pattern of species discovery in the Canaries, whereby, despite dense human populations and some three hundred years of scientific attention, new species have been described at the rate of about one species every six days in recent decades. The discoveries include two large species of lizards (Gallotia intermedia and G. gomerana) and at least two species of trees (Myrica rivas-martinezii and Dracaena tamaranae), and have resulted both from new field and laboratory work (Izquierdo et al., 2004; Whittaker & Fernández-Palacios, 2007). Many new finds on the Canaries, such as the two species of Gallotia, had escaped scientific detection because they persist only in small and endangered populations. Their discovery, as with many recent discoveries around the world, adds both to the global diversity total and to the total of threatened species. These examples are indicative of how difficult a business it is to estimate global diversity. Global estimates of the total species diversity of terrestrial animal and plant species are largely dependent on the estimated number of arthropods. A classic and influential analysis was undertaken by Terry Erwin (e.g. Erwin, 1983), based on a field study of neotropical beetles. Erwin took his samples near the city of Manaus in the heart of the Brazilian Amazon. He used insecticide fumigators to sample beetles from the canopies of three forest types, collecting the specimens from a network of collecting trays spaced out along ten transects, each of 50 m length. He recorded a remarkable 1,080 species (many unknown) from these samples and, significantly for estimates of species richness, 83 per cent of the species he sampled were restricted to one type of forest and 14 per cent to two types. To get from here to an estimate of global arthropod diversity requires some pretty big suppositions. Erwin’s method was to use an estimate of beetle host specificity of 20 per cent (derived from a study of insects on one tree species, c. 163 beetle species per tree species) and multiply this by tropical tree species richness (≈50,000 species). Assuming that beetles comprise about 40 per cent of canopy arthropod species, and that there are twice as many canopy species as ground-dwelling species, it is possible to estimate that there could be as many as 30 million species. Using similar approaches, other researchers have arrived at estimates for global biodiversity of eukaryotes as high as 100 million species (Groombridge, 1992).

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Figure 4.1 Cumulative curves of species description and fitted models of four of the five size categories of mammals analysed on four land masses. The curves start in the year 1890 (year 0), when the number of species was treated as zero. From Medellín & Soberón (1999).

More recently, Ødegaard et al. (2000) have questioned the estimates of host specificity used in such projections. Using data from 2,561 host observations of 697 beetle species on 50 canopy species in a tropical dry forest in Panama, they observed rates of host specificity of 7–10 per cent. Taking into consideration a number of studies, estimates of host specificity of tropical rain forest beetles now range from c. 2 per cent to 20 per cent, which indicates that beetles may only contribute 20 per cent (still a large number!) rather than 40 per cent of canopy arthropods in tropical forests. Based on these new figures, they conclude that 5–15 million species is a far more reasonable range for estimates of global biodiversity than 30 million. This estimate tallies well with Groombridge’s (1992) suggestion that there are probably about 12.5 million species currently in existence. However, as regards prokaryotic species, biologists hesitate even to approach an estimate and, indeed, are unable yet to agree on how to delineate prokaryotic species (Curtis et al., 2006; Doolittle & Zhaxybayeva, 2009).

Global eukaryote species richness can also be estimated based on extrapolations of rates of discovery of new species. Medellín & Soberón (1999) used this method to predict the actual number of mammals in different taxa and size classes in each of Asia, Africa, Eurasia and Oceania. Mammals are typically regarded as a well-known group and it is therefore unsurprising that the rate of new discoveries has slowed over time (Figure 4.1). Even so, extrapolating up to the year 2032 gives an overall forecast of 4,875 species. This represents 247 more than the data set (1992) used in the analysis, with the majority of the expected new discoveries being small-bodied ( 1000 m. (a) The distribution of information of species occurrences. (b) The distribution of the expected diversity as predicted by a bootstrap model that compares the contents of the checklists within a circle with a radius of five degree squares of the focal square. (c) The distribution of the diversity that can be explained by modelling the distributions of the 1,584 species as predicted by assuming that each has a likelihood of occurrence of 50% in degree squares adjacent to those where they are already known to occur, and this additive effect extends within a radius of five degree squares. (d) The modelled distribution of incompleteness of knowledge, derived as the difference between layers b and c. Major rivers of the Amazon Basin are shown in grey. From Hopkins (2007). (See Plate 4.3 for a colour version of these images.)

hypothetical full distribution might be modelled from the known geo-referenced specimens. In Figure 4.3, we see the emergent outcome of repeating this analysis across his full data set of 1,584 plant species. Comparison of the panels shows that the hypothesized real diversity map of Amazonian plant richness might be very different from the ‘known’ pattern of diversity. We stress ‘might’ because, of course, the key point here is that the distributions are incompletely known; the models used to interpolate and extrapolate the real distributions of each species could be unreliable guides (below; Chapter 7).

Information on the geographical distribution of species is not only limited by accessibility of sampling sites, but by the particular history of plant and animal collecting, analysis and compilation for particular countries. Hence, much of our knowledge pertains to political geographical units rather than ‘natural’ ones. Even within comparatively wealthy countries with a shared vision for conservation, such as those in the European Union, there is a damaging paucity of information for many taxa. For example, the Atlas Florae Europaeae (AFE), a project launched in 1965 by a group of botanists and

The distribution of diversity: challenges and applications

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Figure 4.4 Distribution map of Filipendula vulgaris (Rosaceae). The symbols in the key (lower right of figure) indicate distributional status and, from top to bottom, are as follows: native (including archaeophytes); status uncertain; introduction; probably extinct native; and extinct introduction. Reproduced from Kurtto, A., Lampinen, R. & Junikka, L. (eds), Atlas Florae Europaeae (AFE) 13, 2004, by permission of the Committee for Mapping the Flora of Europe and Societas Biologica Fennica Vanamo.

biogeographers at the Museum of Natural History in Helsinki, with the objective of mapping the geographical distribution of all the vascular plants in Europe, has to date published some 13 volumes, including 3,912 maps (example shown in Figure 4.4), yet is reported to include only a little over a fifth of the vascular plants of Europe (www.fmnh.helsinki.fi/english/botany/ afe/). The maps for the first 12 volumes were manually

produced and only recently has the project shifted to direct database entry for mapping purposes. The slow progress of the project reflects the large geographical extent, covering many different countries, and the need to collate the taxonomic information on the species and subspecies during the process. The significance of the Linnean and Wallacean shortfalls for conservation planning is potentially

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immense, not least for convincing politicians and the public of the need to take action. The crux of the problem is that if we don’t know what is out there or how widely species are distributed, how can we convince people about the reality and form of the biodiversity crisis? An equally problematic issue is how to go about filling the shortfalls when funding for conservation in general – and taxonomy in particular – is extremely limited.

4.2.3 The extinction estimate shortfall Whereas the births and deaths of most individuals are discrete, easily recognizable events, it is often difficult to determine when a new species has come into existence, and when the last individual of a species has died. (Lomolino et al., 2010, p. 207) Extinction is a natural process and it is often remarked that of all the species that have lived, only a small fraction are alive today. Processes generating species extinctions over geological time periods include volcanic eruptions, meteorite impacts, climatic changes, marine transgressions, ocean closures and the disappearance of lakes, in combination with biotic forcing as new forms or newly arrived forms have displaced others (Raup, 1991; Lomolino et al., 2010). Rates of extinction have varied throughout the history of life (a period of some 3,500 million years), so it is difficult to distinguish between so-called background rates of extinction and short episodes of unusually rapid or extensive losses. However, analysis of the fossil record suggests that there have been five mass extinctions, each of which may be defined as a major episode of extinction involving many different taxa and occurring fairly suddenly in the fossil record. These five events are each recognizable in the marine record, with the most recent three, the end-Permian, endTriassic and end-Cretaceous events also notable in the terrestrial tetrapod (four-limbed vertebrates) record. Other, lesser, pulses of extinction have also been recognized in the fossil record for terrestrial animals, although mass extinctions are not clearly distinguishable for plants (Willis & Bennett, 1995; Willis & McElwain, 2002). The most recent generally recognized mass extinction event, the so-called K–T event (Cretaceous–Tertiary), occurred at around 65 million years ago and saw the disappearance of the land

dinosaurs, flying reptiles, large sea reptiles and ichthyosaurs. The imperfect nature of the fossil record, the challenge of applying modern species concepts to fossilized remains and the difficulty of distinguishing mass extinction events and pulses means that it is in turn extremely difficult to estimate what might be thought of as the ‘business as usual’ or background rate of extinction. Those who have attempted such calculations estimate that most species persist for perhaps around 4 million years, with a broad range of 1–10 million years average duration, allowing the background rate to be estimated based on average duration and total richness in a group (Raup, 1991). By such back-of-the-envelope calculations of the background rate, anthropogenic extinction rates for birds and mammals have been estimated to be 100 to 1,000 times faster than background (e.g. Primack, 1993). Leaving apart the problems of calculating the background rate, where do the estimates of extinction rates for the period of human-dominance of extinction (sometimes termed the Anthropocene) come from and how good are they? The first step in dealing with the question of estimating anthropogenic extinction rates, past and present, is to recognize that the term ‘extinction’ is used in very different ways, linked to differing sets of assumptions. The typology of extinctions provided by Ladle & Jepson (2008) provides a novel framework for exploring the meaning of extinction (Table 4.2), in which the first two categories refer explicitly to the previous two sections of this chapter, viz. the Linnean and Wallacean shortfalls. Hypothetical population trajectories that might accompany these forms of extinction are provided in Figure 4.5. 1 The term Linnean extinctions refers to attempts to estimate extinctions for areas or regions that are poorly known scientifically, but where the available data leads scientists to believe that large numbers of species exist, many of which are likely to be endemic to the region. When this is applied, for example, to large swathes of tropical forest in the equatorial regions, the use of area-based extrapolations of how much diversity may be present in the pristine state of these systems allows us, in turn, to use species–area relationships to predict how many species will be lost as the habitat is destroyed. In essence, the approach taken in such estimates is closely akin to the Erwin method for extrapolating from local sampling, based on rates of host specificity and

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Table 4.2 Typology and definitions of extinction (modified from Ladle & Jepson, 2008). Type

Definition

Linnean extinction

Extinctions of undiscovered species inferred from the species–area relationship and estimates of species diversity for a given ecosystem or region. The assumed losses of these inferred species are termed Centinelan extinctions by Wilson (1992).

Wallacean extinction

Species that have not been documented for many years but in which extinction is uncertain because populations might survive in areas that have not been surveyed within the potential distributional range.

Phoenix extinction

Extinct in wild but genetic material available in the form of stored material or closely related conspecific or congeneric variety/breed/hybrid, allowing the possibility of a future reintroduction of the same or a functionally equivalent form.

Ecological extinction

Extinct in the wild but with extant captive bred population, or present in the wild but at such low densities that it no longer interacts to a meaningful degree with other species in the community (i.e. it is functionally extinct).

Local extinction

Extinct in the wild within a clearly defined geographic area but with extant free-living populations outside that area.

True extinction 1: Contemporary extinction

Extinction since the birth of the international conservation movement (mid-19th century). Last known population has been monitored and surveyed and is now considered globally extinct in the wild. No captive-bred population or genetic material available.

True extinction 2: Historical extinction

Extinction prior to the birth of the international conservation movement. No authenticated record of an extant population. No captive-bred population or viable genetic material available.

knowledge of tropical tree diversity to arrive at a global species number figure in the region of 30 million. Indeed, a figure of this order of magnitude is implicit in many estimates used by scientists in popular discourse about extinction rates. By starting with a global estimate of this magnitude, and then eating away at it by using estimates of habitat loss (e.g. rates of tropical deforestation) according to a given species–area relationship, it is possible to derive startling figures. Such estimates are characteristically provided in terms of numbers of species that are being lost per annum, often expressed in relation to the size of a country or, if using a shorter time frame, the size of a football pitch. When projected into the future, round number dates such as 2020 or 2050 are used, which allow figures in the order of a million or several million to be deployed within a time frame of relevance to the human life span; such estimates grab the attention of the media (Ladle et al., 2004), signifying the risk of serious losses and the need for action, but are also easy to criticize scientifically.

The typical reliance of such estimates on the dominant island biogeographical paradigm of the last half century, MacArthur & Wilson’s (1967) equilibrium theory, is discussed further in Chapter 8, but for present purposes it is worth noting that estimates following this rationale involve a variety of ecological methods and assumptions (e.g. Wilson, 1992; Thomas et al., 2004), yet tend to make use of the same generalization regarding the relationship between the loss of habitat (area) and inferred loss of species, that the slope (z) of the species–area relationship expressed by the power model (S = cAz, where S = species number, A = Area, and c and z are constants) can be approximated as 0.25. As Gershwin put it in the song from Porgy and Bess, ‘It ain’t necessarily so …’ (Chapter 8; Thuiller et al., 2004; Whittaker & Fernández-Palacios, 2007). 2 Wallacean extinctions is a term given by Ladle & Jepson (2008) to the apparent extinction of a species, whereby the species can no longer be detected in areas in which it previously occurred, but where there is a chance of it persisting in areas within the range that have either

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Figure 4.5 Typology and biocultural context of extinction: (a) Linnean extinction; (b) Wallacean extinction; (c) Phoenix extinction; (d) ecological extinction; (e) local extinction; (f) true extinction (contemporary and historic). From Ladle and Jepson (2008).

The distribution of diversity: challenges and applications never been surveyed or which have not recently been surveyed. In an island biogeographical context, where biologists are interested in the turnover of species through time on a single island, the problem is termed pseudo-turnover, meaning that a species has appeared to go extinct from that island and then re-immigrate from elsewhere when it was actually present throughout, but this was not detected due to inadequacies of survey efforts. This problem has, for example, influenced estimates of extinction and turnover rates of plants and of birds on the Krakatau Islands, Indonesia (Bush & Whittaker, 1991; Thornton et al., 1993). 3 Phoenix extinctions are of species that have the potential to be resurrected through human ingenuity, and they take at least two forms. The first is where species have been transformed by human action for the purposes of domestication. Thus, although European wild cattle (Bos primigenius) became extinct in the 17th century, through back-breeding and artificial selection of domesticated cattle, a wild-acting replicate of the auroch is currently part of the ecology of the Oostvaardersplassen nature reserve in Holland (Ladle & Jepson, 2008). The term ‘phoenix extinction’ might also be applied to the plains bison (Bison bison), which hybridized extensively with domesticated cattle during the late 1800s and early 1900s (Freese et al., 2007), or more recent attempts to create a quagga-like animal by selectively breeding zebra. The second form of phoenix extinction identified by Ladle and Jepson (2008) is where technology is able to recreate extinct species using genetic material stored in gene banks or extracted from preserved remains. This idea has been around for decades and is already used in agriculture for resurrecting different varieties of domesticated species. However, recent advances in genetic technologies have dramatically increased the potential for scientists to recreate extinct species and have stimulated public interest in this possibility. There are currently no credible examples of such resurrections, although the whole mammoth genome has recently been sequenced (Miller et al., 2008), leading to renewed speculation that such a technological feat might one day be possible. 4 Ecological extinction refers to a species that persists in captivity, but which is no longer found in the wild, or occurs in densities so low that it ‘no longer interacts significantly with other species’ (Estes et al., 1989, p. 253). Alternatively, ecological extinction could be described as the avoidance of complete extinction through the intervention of captive breeding. One of

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the classic examples of this is Spix’s macaw (Cyanopsitta spixii), a magnificent neotropical parrot that was collected into extinction, but which clings on to existence in the aviaries of private individuals and institutions (Juniper, 2003). 5 Local extinction or ‘extirpation’ occurs when a species has disappeared from a clearly delimited geographical area (often relating to a geographical or physical boundary), but where extant, free-living populations exist outside that area. A classic example is the red kite (Milvus milvus), which disappeared from England in the late 19th century, but was successfully reintroduced in the late 1980s and early 1990s using populations sourced from Spain, Sweden and Wales (Evans et al., 1999). Another dramatic example that combines properties of both ecological and local extinction is the still-experimental reintroduction of the Californian condor (Gymnogyps californianus) into the vicinity of the Grand Canyon in western North America, a region that had been without condors since the late Pleistocene. While the drivers of local extinction are the same as the drivers of global extinction, the response of the conservation movement may be radically different and, for collectable species, global rarity may even enhance the pressure on the remaining populations. 6 True extinction can be defined as occurring when there is no reasonable doubt that the last population is extinct and where no captive population or genetic material exists. One of the most high profile recent examples of this is the Yangtze River dolphin, or baiji (Lipotes vexillifer), whose disappearance was reported after extensive surveys in November and December 2006 (Turvey et al., 2007). Ladle and Jepson (2008) distinguish between extinctions that occurred before (e.g. the dodo, Steller’s sea cow) and after the advent of the global conservation movement, because of the differing degrees of associated knowledge, certainty and conservation action surrounding such events (Table 4.2). So how many species have humans already driven to extinction in the recent past, and how many may go extinct in the future? First, to deal with the recent past, figures given by different authorities vary depending on the criteria adopted, so the values given in Table 4.3 should be taken as of largely indicative value. They are certain to significantly underestimate the true magnitude of extinction in the last four centuries. For example, based on extrapolations from historical records and sub-fossil remains, it has been suggested

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Table 4.3 Summaries of known extinctions on islands, continents and oceans between c. AD 1600 and c. AD 2000, thus corresponding with categories 6 and 7 (combined) of Table 4.2. These numbers undoubtedly underestimate the real magnitude of species extinctions in the historic period (reproduced from Whittaker & Fernández-Palacios, 2007, their Table 11.2). Group

Islands

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Oceans

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% insular

Mammals Birds Reptiles Molluscs Insects Plants Total

51 92 20 151 51 139 504

30 21 1 40 10 245 347

4 0 0 0 0 0 4

85 113 21 191 61 384 855

60 81 95 79 84 36 59

that as many as 2,000 bird species may have become extinct following human colonization of the Pacific island region during the course of the late Holocene (Steadman 1997), including moas, herons, swans, geese, doves, parrots, owls, many passerines and perhaps hundreds of species of flightless rails. This estimate illustrates the gulf that can exist between recorded extinctions and what we have termed Linnean extinction estimates. Going back a little further in time, the role of humans in the undisputed loss of what is commonly termed the Pleistocene mega-fauna, especially in North America and Australia, remains a hot topic of debate (Barnosky et al., 2004; Lomolino et al., 2010). It seems most likely that a combination of environmental (principally climatic) change, plus human activity, hunting and altered fire regimes, etc., were frequently involved rather than humans being the sole cause. As regards the present-day and future extinctions, there is again a strong scientific consensus that we are entering another phase of accelerated extinction, and some use the phrase a ‘sixth mass extinction spasm’ to describe it. The main message here, however, is that estimating the extent of the building biodiversity crisis presents a significant challenge. An integral and crucial part of this challenge is to identify where the threats are concentrated, which species are most at risk, and where in the world preventative action can have greatest benefits for biodiversity in the various ways in which we define and value it. It is this challenge that motivates much of the remainder of this chapter and of this book as a whole.

4. 3 T H E FU N DAMEN T AL T AXON OMI C U N I T S OF CON S ER V AT I ON B I OGEOGR APH Y As systematists and biogeographers continue to turn to molecular approaches (Riddle et al., 2008), there are signs that reductions in Linnean and Wallacean shortfalls will not progress in parallel, for reasons that include a lack of consensus on criteria used to diagnose species and the utilization of units other than named species in biogeographical and conservation analyses. Here, we explore the conceptual and empirical consequences of using a variety of different metrics of biodiversity for the practice of doing conservation biogeography as outlined in detail in subsequent chapters.

4.3.1 Species versus other geneticallybased conservation units No one definition has as yet satisfied all naturalists; yet every naturalist knows vaguely what he means when he speaks of a species. (Darwin, 1859, p. 101) What are species? An answer to this seemingly straightforward question has eluded evolutionary biologists ever since Darwin and Wallace proposed that species are products of divergence with modification from common ancestors. The species is arguably the fundamental unit for conservation, and preventing species

The distribution of diversity: challenges and applications

Table 4.4 Six major species concepts (modified from Van Dyke, 2003). Name

Basis

Morphological

Morphology

Biological

Reproductive (and geographical) isolation

Genetic

Genetic data, e.g. nucleotide sequence mutations or restriction length polymorphisms

Paleontological

Character state gaps among fossils comparable with those between present-day species

Evolutionary

Ancestral descendant sequence of populations

Cladistic or Phylogenetic

Branch within a cladogram; formal analysis of character states

from becoming extinct the primary goal of the conservation movement. These generalizations mask a high degree of uncertainty and debate regarding both the definition of a species and what taxonomic unit is the most appropriate focus for conservation interventions. More than twenty different species concepts have been recognized (Mayden, 1997), but they can be gathered under six broad headings (Table 4.4), with the starting point being the use of morphological criteria within a traditional taxonomic framework. The most widely accepted modern definition was formulated by Ernst Mayr (1942) and is referred to as the Biological Species Concept (BSC). Mayr defined species as ‘groups of actually or potentially interbreeding natural populations which are reproductively isolated from other such populations’ (Mayr 1942, p. 120). In other words, members of the same species can breed and produce viable offspring, while unrelated species cannot. In this context, speciation is the evolution of reproductive isolation, typically through behavioural or physiological mechanisms, or both. Once a population is reproductively isolated from similar populations, it sets out on a unique evolutionary trajectory. There are several practical and conceptual problems with the BSC. First, it is exceedingly difficult to demonstrate under natural conditions because spatially

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disjunct (non-overlapping, i.e. allopatric) populations rarely have the opportunity to interbreed. Second, as would be predicted under gradualist models of evolution, reproductive isolation is frequently incomplete. For example, the oak species Quercus robur and Quercus petraea sometimes interbreed and produce viable progeny, yet seem to have maintained their biological integrity over millions of years. In contrast, red deer (Cervus elephas) and introduced sika deer (Cervus nippon) in the UK are now so interbred that it is difficult to distinguish them in many localities. Other taxa, such as the bdelloid rotifers, do not even reproduce sexually (Welch & Meselson, 2000), rendering the BSC meaningless. The difficulties encountered with trying to apply the BSC have led to a number of new species concepts being proposed (Table 4.4), based on the use of genetic data. In contrast to the BSC, with its practical focus on reproductive isolation, the phylogenetic (PSC; Baum & Donoghue, 1995; Wheeler & Meier, 2000) and evolutionary species concepts place emphasis on the historical pattern of relationships resulting in distinct entities through a history of descent from a common ancestor. The application of the PSC, in common with the BSC, is beset by an array of practical problems. It has also been argued that any general attempt to replace traditional species units (morphological or biological) with the PSC would be prohibitively costly and would serve to delay progress in pure and applied biology. Nonetheless, the use of phylogenetic approaches to delimiting ‘species’ is rapidly gaining ground, so we need to know whether choice of adherence to traditional or phylogenetic species concepts really matters. As judged by an analysis of the relative numbers and boundaries of entities recognized empirically under each concept, it does (Agapow et al., 2004). In a literature survey of a broad variety of vertebrate and invertebrate groups, the re-analysis of non-PSC-based species using PSC-based criteria increased the number of species by 49 per cent and the average number of species within a group by 121 per cent. Discrepancies of this nature could significantly influence decisions about conservation prioritization if the geographical boundaries of biological species differ from those of phylogenetic species (Peterson & NavarroSiguenza, 1999; Agapow et al., 2004), although some provisional evidence suggests that these discrepancies need not lead to drastic alterations in considerations of protected areas and patterns of endemism (Fjeldså, 2000; Dillon & Fjeldså, 2005). While these issues need

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further analysis across a broader range of taxa and geographical regions, the practical offshoots of a PSC perspective have taken root in several tangible forms, which we will discuss next.

4.3.2 Evolutionarily Significant Units (ESUs) The most frequently cited molecular-based units that sidestep the task of formally naming new species are called Evolutionarily Significant Units, or ESUs. The concept of an ESU originated with Ryder (1986), but has since been developed variously by Waples (1991), Moritz (1994), and Crandall et al. (2000). The standard way to identify ESUs is through analysis of mitochondrial DNA (mtDNA) in animals or chloroplast DNA in plants (cpDNA). This is because these two organelle molecules, being maternally inherited, evolve very rapidly and therefore are more likely than nuclear genes to reflect the history of long-term population divergence through isolation. Moritz (1994) offered the definition that an ESU should be reciprocally monophyletic for mtDNA alleles and show significant divergence of allele frequencies at nuclear loci. This is an attempt to delineate the ‘major historical units’ that arise from long-term geographical isolation and evolutionary divergence (Avise, 2005). Such ‘natural groups’ may be morphologically cryptic but hold within them not only the imprint of their past history, but the possibility of differing responses to future environmental change. Whether or not they satisfy the criteria for the application of the biological species concept, or merely represent sub-specific variation, these ESUs may warrant individual conservation attention. Alternative definitions have attempted to align ESUs with the ‘Distinct Population Segments’ (DPS) amendment to the 1973 Endangered Species Act in the USA (Waples, 1991), with incorporation of a broader recognition of evolutionary processes (Crandall et al., 2000) or with the desirable goal of achieving genealogical concordance across multiple genes (Avise, 2005). Regardless of definition, the underlying power of the ESU concept – with its reliance on molecular data and the ability to ‘do’ biogeography and conservation without the necessity of formally naming new species – appears to remain intact. 420 hits were returned from a query using ‘Evolutionar* Significant Unit*’ in a recent (23 October 2010) topic search of the Web of

Science from 1990 to 2010, with 47 per cent of those records falling between 2006 and 2010, suggesting an ongoing acceleration in the use of ESUs. ESUs represent something of a conundrum for those projects that seek to fully enumerate more traditional and intuitive conservation units (i.e. species) as a baseline unit of biodiversity (e.g. the Catalogue of Life, Species 2000, Encyclopedia of Life; Table 4.1), in the sense that they do not coincide perfectly with either a species or subspecies as a formal taxonomic category (e.g. Zink, 2004) and therefore are not fully accounted for in taxonomic lists. Furthermore, discrepancies between species and ESUs could lead to very different assessments of local and regional biodiversity, which might lead to different inferences in biogeographical and ecological analyses (see Riddle & Hafner, 1999; Kelt & Brown, 2000). Under the Moritz (1994) definition, the discrepancy between ESU diversity and species diversity can be estimated and appears to be considerable, but not overwhelming. In a survey of vertebrate taxonomic species (Avise & Walker, 1999), 56 per cent contained at least two, but less than seven, lineages that could be considered as separate ESUs; Riddle and Hafner (1999) estimated an average of 2.7 ESUs per species in desert rodents from western North America; and Zink (2004) estimated 1.9 ESUs per avian species. Some have argued that ESUs delineated using the Moritz definition should qualify for recognition as distinct species under the PSC (Vogler & DeSalle, 1994). Most often, however, practitioners are properly hesitant to name new species formally based on a singlegene or -genome (most commonly mtDNA in animals and cpDNA in plants) assay of geographic, phylogenetic, and population genetic variation because a single gene tree will often not be entirely congruent with the species tree (Avise, 2005; Edwards, 2009). In other words, a single mtDNA gene tree, even if resulting in several reciprocally monophyletic lineages (i.e. ESUs), is not always going to reflect the same phylogeny for different genes drawn from the nuclear genome. At the same time, biogeographers and conservation biologists do recognize the great operational utility of mtDNA or cpDNA assays as a powerful ‘first approximation’ of evolutionary and biogeographical pattern and processes (Zink & Barrowclough, 2008). They have indeed been used extensively to assay patterns of biodiversity, to postulate associated historical processes and to develop conservation prescriptions.

The distribution of diversity: challenges and applications

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Moritz (1994) also recognized that his definition of ESU may well be too stringent to allow the recognition of some important variation and, for this reason, also proposed a lower level for conservation application, known as the Management Unit (MU). Management units are populations which may not show reciprocal monophyly for mtDNA alleles (and by extension, presumably cpDNA in plants), yet which have diverged in allele frequencies at nuclear or organelle DNA loci and are therefore worthy of some level of monitoring or protection. MUs are significant for conservation in that they represent populations connected by such low levels of gene flow that they are functionally independent.

The most promising aspect of DNA barcoding is that it represents a true union between the goals of systematists, biogeographers, ecologists and an array of other constituencies, including public health organizations. For purposes of conservation biogeography, DNA barcoding will likely have its greatest impact by filling in some of the details of species diversity in lesserknown and morphologically cryptic taxa, and of the distributions of taxa, thereby reducing the magnitude of both Linnean and Wallacean shortfalls.

4.3.3 Other conservation units

4.4.1 Mapping species individually and collectively

One of the latest additions to the units of conservation debate is the suggestion that for large, complex animals such as mammals and birds, conservationists should try to identify culturally distinct population segments, especially if the populations are small or endangered. Such Culturally Significant Units (CSUs) could be vital to future survival if the behaviour confers a distinct adaptive advantage or confers greater adaptability on the population (Ryan, 2006). A good example is the different cultures of tool use seen in wild populations of chimpanzees (Pan troglodytes) in west Africa (Whiten & Boesch, 2001). CSUs are the cultural equivalent of ESUs, in that they also require some form of population isolation although, in contrast to ESUs, significant levels of genetic divergence are not implied. Finally, yet another tool conceptually affiliated to the PSC and phylogeography, and which may speed up identification of new ‘species’, is DNA barcoding: the use of short, unique sections of the genetic profile of a species sample for the purposes of identification, rather like the barcode on your groceries (Hebert et al., 2003). DNA barcoding relies on sequencing a standardized segment of DNA – originally and still most often a portion of the mitochondrial cytochrome c oxidase subunit 1 (COI) in animals and plants, or the nuclear ribosomal internal transcribed spacer region (ITS) in fungi – and comparing that sequence against a database of many thousands of homologous (see Glossary) sequences to determine if it represents a previously unknown species or one that is already registered in the database (http://barcoding.si.edu/; http://www. boldsystems.org).

4. 4 S PAT I AL DI S T R I B U T I ON S : FR OM GEN ES T O B I OGEOGR APH I CAL R EGI ON S

The distributions of species are often represented by what are sometimes termed range-filling maps, i.e. coarsely drawn envelopes encompassing the known outer limits of the species’ distribution, within which ranges are represented as solid entities. In poorly surveyed areas of the world, such maps may involve extrapolations based on general knowledge of the habitat in which the species is known to be found. These maps do not provide a sound basis for conservation planning purposes. The generally recognized protocol for mapping species, however, is to use some form of grid-cell system, preferably using equal areas and typically using grid cells of 10 × 10 or 50 × 50 km or 0.5 degrees (latitude/longitude). The species is recognized as present in the cell only when established to be present by direct observation, ideally backed by voucher specimens collected and stored in herbaria/museums. When species ranges are mapped in this way, it becomes evident that species’ distributions are discontinuous (as exemplified in Figures 4.2 and 4.4). This may be more apparent for some taxa than for others, but it actually applies to all species if distributions are recorded and mapped on a fine enough scale of analysis, although some highly endangered species have such small ranges that they may be considered to consist of a single population and provide a single dot on the map. Given that environments are patchy, it is, of course, unremarkable that species’ distributions should also be patchy. So, for example, in the UK a plant adapted to

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Figure 4.6 The distributions of two plant species within England and Wales. (a) Vulpia unilateralis and (b) Pulmonaria longifolia, demonstrating that assessment of the range size of the two species for the same geographical extent is dependent on the size of the grid used in analysis (i.e. the focal scale). Contrast the number of dots (representing the underlying grain of the data) with the number of 100 km2 cells occupied. (c) The contrasting aggregation patterns of the two species are reflected in the slopes of the relationship between focal scale of analysis (sensu Box 1.2) and estimated range. Triangles and dashed lines for V. unilateralis, circles and solid lines for P. longifolia. From Kunin, W.E. (1998) Extrapolating species abundance across spatial scales. Science, 281, 1513–1515. Reprinted with permission from AAAS.

alkaline environments (a calcicole) would typically be found on scattered limestone outcrops rather than on the acidic soils surrounding them, while birds dependent on reed beds are again likely to have a patchy distribution, restricted to these relatively rare habitats. With further reflection, it is also unsurprising that different species are likely to exhibit different scales of patchiness, reflecting their own individual niche requirements and how they are distributed across landscapes and also reflecting the scale at which the species concerned interacts with the environment. So for example, individual snails explore less geographical space in their lifetime than large raptors or ungulates, but even within a single taxon (e.g. butterflies, or birds) there can be striking differences in the mobility or home range of different species. The upshot of these two properties – the discontinuous nature of ranges and the varying scales of pattern

of individual species – is that assessments of the range occupancy of the species found in any particular region of the world are scale dependent. This is neatly illustrated in Fig. 4.6, which shows that while two plant species, Vulpia unilateralis (panel a) and Pulmonaria longifolia (panel b), have the same range size in England and Wales when judged using 100 km2 grids, but when judged at a finer scale of analysis, Pulmonaria is shown to have a significantly larger range, while at the coarsest scale plotted the reverse occurs (panel c). The figure thus illustrates two things: first, the smaller the grid size used, the smaller is the estimate of the range size; and second, that the relative order of this measure of rarity can change as a function of the scale of analysis. The reason for this is that V. unilateralis has a generally scattered distribution while P. longifolia has a tightly constrained distribution within the south of

The distribution of diversity: challenges and applications England (Kunin, 1998). Extrapolating the scale/area curves suggests that P. longifolia may be very much more common than V. unilateralis at still finer scales, although it is dangerous to attempt to infer the size and viability of the populations of either species from such range map data. As range size is a key criterion in determining whether a species should be considered endangered (Box 4.1), especially where other data are lacking, the

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scale-dependency of range size estimates is a potentially important issue in making these assessments and in comparing data from different regions of the world (for further discussion of scale dependency see Box 1.2; Rahbek & Graves, 2000; Lennon et al., 2001; Whittaker et al., 2001; 2005). As we have repeatedly emphasized in this chapter, knowledge of the geographical distribution is fundamental in making sense of the ecological requirements,

Box 4.1 Rarity, range restriction and the Red List Rarity is often a precursor to extinction. However, not all rare species are rare for the same reasons. Rarity can mean that a species occurs at low densities, is adapted to a narrow range of environmental conditions, or occupies a small geographical range. These three categories (abundance, habitat breadth and geographical range) form the basis of Deborah Rabinowitz’s widely used categorization of rarity (Rabinowitz, 1981). Under this simple classification scheme, seven types of rarity can be recognized (Figure B4.1a), with the most vulnerable category being species that have low density, low population size and which utilize a narrow range of habitats. It should be noted that this is not the only scheme for defining rarity (e.g. Manne & Pimm, 2001), but it is one of the most frequently applied. Species are by no means evenly spread across the eight categories shown in Figure B4.1a. For example, it has long been known that while a high proportion of species have relatively small geographical ranges, there are few that are widespread and abundant. However, Brown (1995) has shown graphically that within taxonomically or ecologically similar species there tends to be a positive correlation between range size and density. One consequence of this is that species occupying large ranges tend to be more abundant throughout those ranges than are range-restricted species. When examining whole faunas, such as the Breeding Bird Survey data for North American land

Figure B4.1a Rabinowitz’s seven forms of rarity. Species in the (eighth) upper left cube at the front exhibit no component of rarity, being common across a large geographic range and possessing a wide habitat breadth. Those at the lower back right have all three components of rarity: small geographic range, narrow habitat breadth and low local density. Taken from Ricklefs (2000).

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Basic biogeography: estimating biodiversity and mapping nature

birds, one finds that a lot more scatter is evident than within similar groups of species, but still the condition of ‘abundant and localized’ is extremely rare, at least within continental avifaunas (Gaston, 1994; Brown, 1995). Perhaps the more important question from the perspective of conservation biogeography is what constitutes a small range and how might the range size/abundance pattern vary in relation to biogeographical context (e.g. continental versus island archipelago contexts) or for taxa of differing body sizes? As previously noted by Whittaker et al. (2005), within the North American breeding bird data set analysed by Brown (1995), range size varied from c. 10,000 km2 to over 10 million km2 and fewer than 10 species had ranges 99% barriers of its habitable volume; with few physical barriers, the majority of the ocean is physically interconnected; ocean currents, gyres, upwellings, can operate on a huge scale; physical structure is limited to benthic topography e.g. sea mounts, plateaus, trenches and coastal morphology

Implications for biogeography

Implications for conservation

Potential for species to have vast ranges and for individuals to move over huge distances; depth and size of sea precludes use of many of the remote sensing devices that are used on land

Need to plan on large scale; wide species distributions do not necessarily reduce their vulnerability to extinction; only ‘one ocean’ means that action taken in one area affects others (this can be positive, e.g. allowing replenishment of depleted areas; or negative, e.g. movement of pollutants, invasive species)

b) Buoyancy and density

Increased density of water compared to air enables organisms to be buoyant with little energetic expenditure and thus enables life-history strategies that would be untenable on land (e.g. pelagic jellyfish, giant kelp)

Three-dimensional nature hard to map

Challenge for planning and implementing conservation actions in three dimensions, e.g. restricting fishing to certain depth above seamounts

c) Currents, waves and tides

Sea is very dynamic over timescales from hours to years (waves, tides, currents, El Niño); these physical movements affect many marine organisms and can act as corridors or as barriers

Biogeographical boundaries can be fluid on a variety of timescales

Humans activities, e.g. dredging, sea-filling (land reclamation), coastal development, global warming, etc., can alter hydrology and therefore affect recruitment and distributions of marine organisms; planning of protected areas on the high seas may have to be fluid

Many organisms make use of multiple habitats and widely separated regions during their lives (e.g. many reef fish have pelagic larvae which grow up as juveniles in seagrass habitats before settling as adults on reefs); some species make vertical migrations

Marine protected areas cannot necessarily be treated as islands; need protection of different habitats for different parts of the life cycle to ensure population viability

LEVEL 2 – BIOLOGICAL a) Ecological High levels of connection connectivity between systems, e.g. between benthic/pelagic zones, between different habitats, between widely separated areas, between land/sea

The distribution of diversity: challenges and applications

Table 4.6 Continued

Marine situation

Implications for biogeography

Implications for conservation

b) Genetic connectivity

Reproduction and recruitment can be widely separated spatially and temporally; gene flow can occur over large spatial scales; metapopulation structure important

Some areas are net sources, others are net sinks of propagules; recruitment highly dependent upon physical processes

Potential for areas to receive recruits from far away; systems or networks of marine protected areas may be far more effective than single large ones

c) Primary productivity

Mainly in the form of highly mobile phytoplankton with high rates of turnover; results in a very responsive system

Pelagic bioregions very hard to map (compared to on land, where rooted plants are used to define bioregions and transitions are generally swift and stable over time)

Marine systems can adjust quickly to physical or biological changes (e.g. changes in community composition and phase shifts due to over-fishing) because they are not buffered by long-lived primary producers

d) ‘Rooted’ ecosystems

Some systems are ‘rooted’ (e.g. coral reefs, seagrass, mangroves, kelp beds, hydrothermal vent communities) and are stable over years

‘Rooted’ habitats can be relatively easily mapped

MPAs encompassing ‘rooted’ habitats need not be mobile

e) Land–sea connection

Sea is always ‘downstream’ from land; coastal systems in particular are highly affected by land-based activities (e.g. silt from rivers smothering coral reefs; outflow of rivers heavily polluted by agricultural, or mining chemicals causing dead zones; rivers providing input of iron to the sea); sea affects land, too, but it is not an equal exchange

Very strong ecological/ diversity gradients moving from land to coast to sea; challenge for GIS methods and area-selection algorithms which use species richness as criteria for selection (these will inevitably choose coastal regions, because their transitional nature results in high species richness)

Land–sea links need to be included in conservation plans for organisms that directly (e.g. sea snakes, marine mammals, salmon) and indirectly (e.g. bears eating salmon, riparian systems that derive nitrogen from salmon carcasses) utilize both marine and freshwater ecosystems; need whole catchment conservation

Humans live on, not in the sea, and hence we are less familiar with patterns of diversity beneath the waves

Coastal and continental shelf regions are highly threatened and also highly valuable, so should be conservation priorities

LEVEL 3 – SOCIO-POLITICAL a) Occupation Approximately 60% of the and use global population live within 50 miles of the sea; coastlines and shelf areas are heavily impacted from exploitation, pollution and development; much of the open ocean is less affected

85

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Table 4.6 Continued

Marine situation

Implications for biogeography

Implications for conservation

b) Political boundaries

Political boundaries have been designated in the sea only recently, e.g. United Nations Convention on the Law of the Sea (UNCLOS), 200 nm Exclusive Economic Zones (EEZ) (1982) or designation of state or municipal waters

Political designations in the sea generally have no ecological meaning (on land they often coincide with major ecological divisions, e.g. seas, mountain ranges)

Political boundaries frequently fail to coincide with biogeographical ones, especially in the sea; successful conservation, particularly on the high seas, requires international co-operation; innovative funding that crosses political boundaries is needed

c) Institutional infrastructure

Responsibility for the high seas lies with international organizations, e.g. International Maritime Organization (IMO), UNCLOS; national responsibility often under the jurisdiction of agriculture, forestry or fisheries; new marine and coastal ministries are being created, e.g. Indonesia

Institutional classification schemes, e.g. Food and Agriculture Organizations (FAO) statistical areas, generally have no basis in ecology – for example, Area 57 stretches from the tropical Indian Ocean to the Antarctic

Co-operation among nations at international level is a challenge; ministries of agriculture or forestry may have little understanding of the differences between terrestrial and marine systems as they relate to biogeography and conservation

d) Exploitation

We use the sea extensively and damage it through destructive exploitational activities, e.g. trawling, dredging, mining and through dumping of wastes; the sea is where the last wild vertebrates are hunted on a large scale; fishing as occupation of last resort

Biogeographical classifications will be most useful if they are at a scale at which exploitation is managed

High exploitation requires that any conservation is management-linked and aware of stakeholders, particularly subsistence fishers; conservation criteria commonly economic and utilitarian in the sea (e.g. MPAs for fisheries management)

e) Social attitudes

Historical perceptions of marine organisms as resources, not wildlife; sea as being limitless; open access

Lack of interest/ engagement, hence hard to get funding for taxonomic/ biogeographical research

Need education and public participation even to put forward the concept that the sea is in trouble

f) Knowledge level

Scientific and public knowledge of marine organisms, distributions and processes that maintain them is extremely limited; diversity on seamounts recently discovered; little is known about the scale of dispersal/ connectedness in most systems

Limited knowledge is hampering accurate determination of marine biogeographical regions, even where they could be easily mapped; potential to use modelling to help determine spatial distributions in unstudied areas

Clearly fewer data, but this should not result in inactivity; would be wise to follow the Precautionary Principle

Regional classification of coastal zones: based on coastal geomorphology and biotic associations

Habitat classification: specific biotic characteristics

Divisions: 14 subdivisions of the domains plus shelf regions which are interpreted as shallow variations of the domains involved

4. Domains: 3 defined by temperature

5. Zoogeographical regions: 3 based on temperature, plus 1 oceanic region

Provinces: oceanic regions based on flow characteristics

Ecoregions: biomes subdivided by ocean basin

3. Biomes: 3 main oceanic biomes based on prevailing winds, plus 1 coastal biome

(Ray, 1975)

(Bailey, 1998)

(Longhurst, 1998)

(Ekman, 1953)

Provinces: based on faunal changes, but biased towards certain taxa

2. Regions: 7 based on temperature and depth, e.g. tropical warm shelf, Mediterranean, temperate

Micro (10% faunal endemism

Provinces and biotones: defined by endemic fish (2 pelagic plus 9 benthic); Biotones: defined by overlapping fish distributions (2 pelagic plus 8 benthic) plus 9 watermass characteristics classes and 4 seafloor topology classes for off-shelf areas

Meso-scale regions: 60 defined using biological and physical information and geographical distance along the coast

Meso (≥1 : 1,000,000)

Ecosystems/ communities: at scale of 1 : 100 000 or finer; 8 habitat categories for shallow water, from satellite images and ground truthing

(Environment Australia, 1998) [meso-scale regionalization] – N.B. different states used different methods of analysis

(Hayden et al., 1984)

(Briggs, 1974)

Micro (1 million ha) to tiny (10a >12.5b

34.1 19.9 18.4

54.6 54.4 54.4

49.1 29.9 27.6

80.3 80.1 80.0

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The shaping of the global protected area estate

Coastal and continental shelf areas are the most heavily exploited areas; they provide the majority of the global fisheries catch, contain the highest levels of biodiversity, generally operate on smaller spatial scales, are more predictable than pelagic systems and are the most tractable politically (de Fontaubert, 2001). However, it is becoming clear that deep water and pelagic systems are also in trouble. The most serious impacts are from fishing, including the removal of 25–35 per cent of the primary production from upwelling and temperate continental shelves, the loss of top predators from coastal and pelagic food webs, gross depletion of target stocks and massive wastage from by-catch (Hyrenbach et al., 2000). The scale of these activities is shocking, as is the general lack of public awareness of the extent of over-harvesting of the seas, notwithstanding the publication of research findings in prominent international journals (e.g. Jackson, 2008). Pelagic species, processes and ecosystems occur in the water column and are not attached to the substrate. They present major challenges to place-based conservation planning. Such systems are highly dynamic in space and time. Upwellings and fronts can shift tens to hundreds of kilometres between seasons and years, while gyres and eddies may be ephemeral, lasting weeks or months. On the other hand, many pelagic species use highly predictable habitats in which to breed, forage, or travel. Static bathymetric features (e.g. reefs, shelf-breaks, seamounts, hydrothermal vents) create discontinuities in the ocean that can lead to aggregations of biodiversity; persistent hydrographical features (e.g. fronts, currents) act as oceanic highways and signposts; and ephemeral hydrographical features (e.g. eddies, winddriven upwellings) provide resources for high levels of productivity. These habitat types reflect increasing unpredictability and therefore increasing challenges to MPA planning, given the desirability of MPAs having explicit geographical boundaries (Hyrenbach et al., 2000). One of the limiting factors to setting conservation priorities beyond EEZs is a lack of available information on geomorphology, oceanography and marine species’ distributions. To date, only 5–10 per cent of the sea floor has been mapped with a resolution comparable to that on land (Wright & Heyman, 2008), yet even midresolution bathymetric data suggests that there are over 14,000 seamounts, the majority of which are beyond national jurisdiction (Harris, 2007). In the last

few years, temporally and spatially resolved ocean provinces have been generated from satellite data (Oliver & Irwin, 2008) and a global open oceans and deep seabed classification (GOODS) has been published (UNESCO, 2009). Recent research has also mapped global distribution patterns for seabirds, turtles, and marine mammals, with a view to using these data to set conservation priorities (Cheung et al., 2005; Worm et al., 2005). Deep benthic and pelagic systems occur outside, as well as inside, political boundaries. Where they fall in international waters, additional challenges of governance, ownership, and enforcement arise (De Fontaubert, 2001; Hislop, 2007). Despite all the aforementioned challenges, major NGOs are developing initiatives for pelagic and high seas areas. Greenpeace has proposed a set of three high seas marine reserves (the Pacific Commons), delineated primarily by political boundaries (Roberts et al., 2006). The World Wildlife Fund is also preparing a high seas priority list. These initiatives suggest that we may be preparing to take the next steps in marine conservation planning, building on technical, institutional and societal advances to make the case for large MPAs in the open ocean, with dynamic boundaries and extensive buffers. On the balance of evidence, they appear sorely needed, along with other monitoring, conservation, and management measures.

5. 5 CU R R EN T T R EN DS AN D FU T U R E DI R ECT I ON S The last 50 years have seen a remarkable period of expansion of the protected area estate globally, albeit in response to a dramatic increase of human demands for land conversion and natural resource extraction. Over this period, we may trace several phases in which different approaches and organizations have taken the lead. The initial lead given by the IUCN has been followed more recently by a phase in which a few major conservation NGOs, such as Conservation International, WWF and BirdLife International, have been dominant forces in shaping strategic conservation on the global stage, generating a raft of (to varying degrees) complementary and competing planning frameworks and initiatives. These schemes stem from a quite small number of core concepts, rationales and objectives but, as we have seen, they also sum to provide a rather

The distribution of diversity: challenges and applications bewildering array of frameworks and designations. The most rapid developments in this field have occurred in the last two decades, as very substantial resources have been mobilized towards protected area planning and conservation prioritization on land and, increasingly, in the oceans. One notable recent trend is the extent to which many of the major players are pooling resources and schemes together once again (cf. Mace et al., 2000), after a ‘breakaway’ period in which a few newly founded or re-founded NGOs changed the organizational frame of global conservation science and planning. The Key Biodiversity Areas scheme is one manifestation of this process, as is the re-drawing of hotspots boundaries on to the WWF Ecoregions maps in the latest CI-2004 hotspots iteration (above). A second notable trend is the extension of protected area planning schemes into the productive landscape and the engagement of major corporate actors in the identification, protection and management of areas with high conservation value. The number and extent of such ‘commercial’ protected areas seems set to expand, and this will require their future integration in global protected area planning and accounting frameworks. There is no particular reason to think that the phase of planning protected area systems is at an end. Even when the broad outlines of a protected area system have been decided and legally recognized, there is a need for reiterative phases of data refinement, monitoring, examination of shifting priorities, changes of designation and resourcing. This remains the case globally, as well as for particular regions (e.g. Loucks et al., 2008; Soutullo et al., 2008). It is evident, especially at finer, landscape scales of analysis, that, even in countries with strong legal codes and frameworks, designation decisions can sometimes be revised. Further discussion of the tools available for conservation planning at these finer scales of analysis, and in the light of ongoing global environmental change, is provided in the following two chapters.

F O R DI S C USS I ON 1 How well do the modern NGO-driven protected area planning approaches discussed above capture and reflect the foundational social values of the conservation movement discussed in Chapter 2?

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2 What are the respective strengths and weaknesses of hotspot approaches and biogeographical representation approaches to global conservation planning? 3 How important is it to continue to refine the science underpinning protected area planning? 4 How appropriate is it to transfer the hotspots rationale to the marine realm? 5 Ecoregion-based conservation planning enables nested scales of analysis. How important and unusual is this feature? 6 Are contemporary protected area planning approaches truly science-based, or might it be better to consider them the products of an interplay between science, politics and organizational interests?

S U GGES T ED R EADI N G Belt, H. van den (2004) Networking nature, or Serengeti behind the dikes. History and Technology, 20, 311–333. Dasmann, R.F. (1972) Towards a system for classifying natural regions of the world and their representation by national parks and reserves. Biological Conservation, 4, 247–255. Eken, G., Bennun, L., Brooks, T.M., Darwall, W., Fishpool, L.D.C., Foster, M., Knox, D., Langhammer, P. Matiku, P. Radford, E., Salaman, P., Sechrest, W., Smith, M.L., Spector, S. & Tordoff, A. (2004) Key Biodiversity Areas as site conservation targets. BioScience, 54, 1110–1118. Hughes, T.P., Bellwood, D.R., Folke, C., Steneck, R.S. & Wilson, J. (2005) New paradigms for supporting the resilience of marine ecosystems. Trends in Ecology & Evolution, 20, 380–386. Jackson, J.B.C. (2008) Ecological extinction and evolution in the brave new ocean. Proceedings of the National Academy of Sciences USA, 105 (Supp. 1), 11458–11465. Jepson, P. & Whittaker, R.J. (2002) Ecoregions in context: a critique with special reference to Indonesia. Conservation Biology, 16, 42–57. Lourie, S.A. & Vincent, A.C.J. (2004) Using biogeography to help set priorities in marine conservation. Conservation Biology, 18, 1004–1020. Olson, D.M., Dinerstein, E., Wikramanayake, E.D., Burgess, N.D., Powell, G.V.N., Underwood, E.C., D’Amico, J.A., Itoua, I., Strand, H.E., Morrison, J.C., Loucks, C.J., Allnutt, T.F., Ricketts, T.H., Kura, Y., Lamoreux, J.F., Wettengel, W.W., Hedao, P. & Kassem, K.R. (2001) Terrestrial ecoregions of the world: a new map of life on Earth. BioScience, 51, 933–938. Soutullo, A., De Castro, M. & Urios, V. (2008) Linking political and scientifically derived targets for global biodiversity conservation: implications for the expansion of the global network of protected areas. Diversity & Distributions, 14, 604–613.

CHAPTER 6 Systematic Conservation Planning: Past, Present and Future James E.M. Watson1, Hedley S. Grantham1, Kerrie A. Wilson1 and Hugh P. Possingham1 1

School of Biological Sciences, The University of Queensland, Brisbane, Australia

6 . 1 I N T R OD UC T I ON In general, the best farming land is the first to be cleared. In long-settled regions of the world, this has meant that by the time biodiversity conservation became a social priority, a very much non-random subset of the ‘original’ habitat types has been available for conservation management. Historical decisions on where protected areas were located were rarely based solely (if at all) on scientific assessment of biodiversity value or biogeographical representativeness. Rather, these decisions were based on other factors, such as the suitability of alternative land uses, availability of an area for conservation management, scenic beauty, and recreational values (Chapters 2 and 5; Pressey et al., 1993; Margules et al., 2002; Gaston et al., 2008). This has resulted in a legacy of protected areas that are biased towards habitats that are generally not threatened, such as dry, infertile or steep habitats (Pressey et al., 1993, 2002; Soulé & Terborgh, 1999). For example, five per cent of the Earth’s entire terrestrial protected area (972,000 km2) is the Greenland National Park, which contributes little to biodiversity conservation as it contains mostly ice (Chape et al., 2003; WDPA Consortium, 2006). This form of bias can be demonstrated quantitatively, as shown in a regional-scale analysis by Pressey

et al. (2002). In their paper, a part of their analysis was an assessment of protected area coverage as a function of slope and fertility in the northern eastern region of New South Wales, Australia (Figure 6.1). This analysis highlights the bias often found in reserve systems, with the steepest slopes and the soils of lowest fertility being far more represented in the reserve system than the converse. There are examples like this found on all inhabited continents on Earth (Brooks et al., 2004; Rodrigues et al., 2004a; Joppa & Pfaff, 2009). The threat to biodiversity as a result of habitat loss and change over the second half of the 20th century led to an increased interest in enhancing the coverage and representativeness of the protected areas network (McNeely, 1994). The efforts taken towards these goals at a global and regional scale gained impetus from the development of the IUCN biogeographical regions (Dasmann–Udvardy) framework discussed in Chapter 5. This coarse-scale analysis did not, however, offer guidance on designing networks within regions at the scale of landscapes. The first efforts to take a more scientific approach to designing protected area networks were based on the theory of island biogeography (e.g. Chapter 8; MacArthur & Wilson, 1967; Diamond, 1975a). The rationale followed was that nature reserves and other protected areas can be considered forms of

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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Figure 6.1 Assessment of reserve coverage as a function of slope and fertility in the northern eastern region of New South Wales, Australia. The vertical axis represents the percentage of the total area of each broad environmental unit captured in reserves in the region. The other two axes are measures of slope and soil fertility, with the lower numbers (i.e. 1) indicating flatter slope and lower fertility and the higher numbers (i.e. 3) indicating steep slopes and high fertility, respectively. From Pressey et al. (2002).

habitat islands, isolated from other reserves by anthropogenically transformed habitats (sometimes named the ‘matrix’) that are generally unsuitable for the species of conservation concern. These early efforts were guided by basic ecological principles, such as that bigger protected areas are better than smaller ones because they are likely to contain more species (Diamond, 1975a). Initial approaches to systematic conservation planning were developed based on simple scoring systems, using criteria such as species richness or number of endemic species, to provide an indication of how new areas might contribute to protected area networks if they were chosen (Margules & Usher, 1981; Smith & Theberge, 1986). The integration of these basic principles into conservation planning was a useful first step, but both conservation scientists and practitioners have since criticized their simplicity (e.g. Simberloff & Abele,

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1982). For example, it has been argued that the number of species contained in any single area alone should not determine its priority. More important is how any one area complements the existing protected area network, along with a suite of wider landscape conservation issues (Moilanen, 2008). Similarly, simply relying on the idea that ‘a big reserve is better’ is not useful for making planning decisions in landscapes also required for other human uses (e.g. agriculture, mining). Systematic conservation planning has evolved as a discipline to enhance the efficiency of protected area network design and, through creating alternative proposed networks, to allow scientists and stakeholders to better engage with the complexity of multi-sectoral spatial planning across landscapes and within regions. Therefore, in general, the tools discussed in this chapter are typically employed at finer scales of analysis than the global/regional approaches discussed in Chapter 5 (but see Venter et al., 2009; K.A. Wilson et al., 2009). The 1980s saw the first attempts to use detailed biogeographical information and selection algorithms in the design of protected area networks (Kirkpatrick, 1983). The field of systematic conservation planning has grown significantly since. It has influenced conservation planning by some of the major environmental organizations such as The Nature Conservancy (Groves et al., 2002) and Conservation International (Myers et al., 2000), and it has shaped policy legislation and conservation in both terrestrial (Knight et al., 2006; Kremen et al., 2008) and marine (Davis, 2005; Fernandes et al., 2005) environments. It has featured in hundreds of peer-reviewed papers (Pressey et al., 2007) and in recent books (e.g. Margules & Sarkar, 2007; Moilanen et al., 2009). In this chapter we review the key principles of systematic conservation planning and some of the current decision support tools available to assist conservation planners in making decisions. Decision support tools are information systems intended to help decisionmakers compile and analyse data to help solve conservation problems. The increasing power and ease of use of such computer-based systems in the last two decades has opened up exciting possibilities for applications to conservation planning. We illustrate some of these applications from contemporary case studies, providing examples of the use of different techniques and tools. The field of conservation planning is rapidly changing, and we discuss advances (and future challenges) in systematic conservation planning at the end of the chapter.

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6. 2 W HAT I S SY ST E M AT I C CO N S E R V AT I ON P L ANNI NG A N D WHY USE I T ? The science of systematic conservation planning is concerned with the optimal application of spatiallyexplicit conservation management actions to promote the persistence of biodiversity and other natural features in situ (Margules & Pressey, 2000; Margules & Sarkar, 2007). It involves a transparent process of setting clear goals and objectives, and of planning conservation actions that meet them (Bottrill & Pressey, 2009). A fundamental characteristic of systematic conservation planning is the principle of complementarity (Kirkpatrick, 1983). Since the first publications in this research field, systematic methods have identified systems of conservation areas that are complementary to one another in terms of collectively achieving objectives. Areas identified in this way will each contain, for example, different species or complementary portions of the required areas of different habitat types. As will be discussed further in this chapter, this represents a major improvement on the additive scoring procedures that were used extensively before the application of complementarity methods. Additive scoring approaches are incapable of dealing with the fundamental notion of building a system of protected area where the value of the whole system is not the same as summing the values of the separate protected areas. Systematic conservation planning has traditionally been applied to design strict protected area networks (those areas that are managed for conservation values only, e.g. IUCN management categories I–IV; see Table 2.2). More recently, however, it has been expanded to include planning other types of conservation actions, such as stewardship payments or other land management, in space (Carwardine et al., 2008) and time (Wilson et al., 2007). Here we use the term ‘protected area’ loosely, in reference to any place where an action is applied for conservation purposes. We acknowledge that much of what is written in this chapter is focused on the literature behind planning protected area networks, but at the end of the chapter we provide examples of other forms of systematic conservation planning. It should be noted that systematic conservation planning involves designing protected area networks based on clear objectives, as well as an understanding of constraints on where and how implementation can

occur (Smith et al., 2006). Constraints include factors such as the cost of acting in a particular area or the willingness of landholders to participate in a conservation initiative (Knight et al., 2009). Good systematic planning processes, as we will see, include input, information and values from a wide variety of stakeholders, incorporated within a transparent and inclusive process (Knight et al., 2006; Bottrill & Pressey, 2009) in order to reduce conflicts between opposed interests. However, as in any such field, the approaches discussed herein have also spawned many analyses that are of largely heuristic rather than immediate practical value. This allows analysts to explore ‘what-if ’ scenarios concerning future landscapes and climate surfaces or to undertake ‘tests’ of the effectiveness of existing protected area networks or schemes (e.g. Araújo et al., 2004a,b, 2008; see Chapter 7).

6. 3 CON CEPT S AN D PR I N CI PLES 6.3.1 Representativeness An overarching goal of conservation is to ensure that there is no loss of biodiversity. As discussed in earlier chapters, representativeness is a fundamental principle in systematic conservation planning and refers to how well protected area networks contain representative samples of every feature of biodiversity that we aim to protect. Biodiversity features normally reflect some combination of genetic, species and community diversity. However, it is also important to consider the structure of habitats, e.g. the availability of coarse woody debris in temperate woodland, and ecological processes, such as fire dynamics in Mediterranean ecosystems. It is often difficult for protected areas to achieve complete representation for two reasons: 1 in regions with high species compositional turnover over small distances, such as Mediterranean ecosystems (Judd et al., 2008), a large proportion of the region will be required to represent all of the unique biodiversity features; and 2 even for the best studied regions, systematic data are lacking for some aspects of biodiversity. This second problem has two elements, termed ‘Linnean’ and ‘Wallacean’ shortfalls (Chapter 4; Whittaker et al., 2005). The Linnean shortfall refers to our lack of knowledge of how many, and what kind,

The distribution of diversity: challenges and applications of species there are. Almost two million species have had formal scientific names given to them, but this is still only a fraction of the total of all species (Groombridge & Jenkins, 2002). Estimates have been made that if the collection and description of new species were to continue at the current rate, it would take several thousand years to catalogue the world’s biodiversity (Soulé, 1990). The Wallacean shortfall refers to our inadequate knowledge of the global, regional, and local distributions of the species that we know. Even for the best known taxa such as birds and mammals, and in the best studied regions, there are still huge gaps in our knowledge of distributions (Chapter 4).

6.3.2 Persistence (adequacy) Having a representative protected area network does not ensure that biodiversity within the network will persist into the future. This is because although protected areas might contain a particular species or habitat type, the area might not alone be sufficient to ensure their persistence. Therefore, protected areas should ideally also be designed to maximize persistence. This can involve an analysis of viability requirements (Lande et al., 2003); the configuration of protected areas, including dealing with issues such as connectivity and the permeability of the matrix (McIntyre & Hobbs, 1999; Lindenmayer & Franklin, 2002); and predicting what ecological processes are needed to sustain biodiversity (Soulé et al., 2004). While persistence is considered one of the most fundamental concepts of systematic conservation planning, exactly what constitutes adequate conservation is not well defined (Woinarski et al., 2007; Watson et al., 2008; Carwardine et al., 2009). For example, is a conservation plan that gives every species a 75 per cent chance of persisting for 1,000 years adequate?

The principle of efficiency is based on the idea that conservation planners should try to achieve biodiversity objectives for the least possible cost. ‘Cost’ here may reflect the financial cost of implementing and managing protected areas or the costs of lost opportunities for economic development (Naidoo et al., 2006). It can also include other socio-economic considerations, such as the willingness of people to assist with conservation management, with the expectation that it is more cost-effective to do conservation where people are willing to act. For example, take the matrix on the distribution of four species at five sites shown in Table 6.1. If you were to select the minimum number of sites to represent each species, the optimal combination would be sites 1 and 2 (at a cost of $25). However, when we take cost into account, the combination of sites that represents all species with the least cost is the set comprising sites 1, 4 and 5 ($11). By such consideration of cost, conservation planners are able to maximize the conservation ‘return on investment’ and hence make an efficient plan. There is an increasing number of studies that provide evidence that incorporating financial constraints into conservation planning increases the likely biodiversity benefits for a given amount of money (Ando et al., 1998; Naidoo et al., 2006; Carwardine et al., 2008). Other benefits from biodiversity conservation can be factored into such analyses, including ecosystem services – the benefits that humans derive from natural systems, such as clean air and water. By dealing with multiple measures of benefit, conservation planners may provide a more comprehensive evaluation of the returns from conservation investments.

Table 6.1 Matrix showing the distribution of four species at five sites.

6.3.3 Efficiency A simple way to ensure representativeness and persistence is to conserve everything. This is obviously impossible, and so some degree of compromise is necessary. If the impact of conservation actions on the rest of society is minimized, there is a better chance that the plan will succeed politically and socially and thus provide a platform from which to expand further actions.

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Species Species Species Species Cost

1 2 3 4

Site 1

Site 2

Site 3

Site 4

Site 5

1 0 0 1

1 1 1 0

1 0 1 0

0 0 1 0

0 1 0 0

$5

$20

$5

$4

$2

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It is also possible to modify the cost of conservation by incorporating into the analysis the benefits obtained from the delivery of ecosystem services, such as the amount of carbon sequestrated or the reduced cost to filter water (Venter et al., 2009). Such payment for ecosystem services has the potential to increase the support and resources available for conservation (Costanza et al., 1997; Daily et al., 2009).

6.3.4 Flexibility Objectives can often be achieved in a number of alternative places, particularly when the distribution of biodiversity features is widespread. Moreover, proposed new conservation areas or networks must be accepted and implemented by planning bodies, which brings economic, political and social considerations to bear upon decisions. Therefore, a key principle of systematic conservation planning is flexibility. A flexible conservation plan provides alternative solutions and assists planners to take account of opportunities (Knight & Cowling, 2007). This is because socio-economic constraints may not be fully understood and, in any event, may constantly be changing. For example, a piece of land with high conservation value might not initially be available for conservation management, but may later become available for sale, lease or other management intervention (McDonald-Madden et al., 2008). Adopting a flexible plan also gives scope for sensible resolutions of resource/use conflicts.

6. 4 DEV E L OP I NG A S YS T E M AT I C CO N S E R V AT I ON P L AN In this section we provide examples of how objectives can be set against the key principles outlined in section 6.3. The process of defining measurable objectives is one of the principal components of systematic conservation planning (Nicholson & Possingham, 2006). Defining objectives gives the planning approach transparency and a benchmark by which to evaluate progress towards goals. We discuss how all stakeholders (and not just planners sitting in academic or government institutions) need to be involved in the process of developing these objectives to ensure the plan is successfully implemented. We also provide two real-world case studies to help describe how each of these

principles has been successfully integrated into an applied systematic conservation plan through the careful choice of conservation objectives.

6.4.1 Achieving representation As discussed earlier, ‘Linnean’ and ‘Wallacean’ shortfalls in biogeographical data are highly problematic for any plan trying to achieve representation. Given such a deficit of knowledge and data on biodiversity, a partial measure of biodiversity is almost always used as a surrogate for the rest of biodiversity. To develop biodiversity surrogates, conservation planners must gather all existing data sets and determine which are fit for purpose. Decisions on which data sets to use will often be based on the likely effectiveness of the particular data set and biodiversity metric as a surrogate for other components of biodiversity for which we have no data or poor data. However, the mere existence of a data set does not necessarily guarantee fitness for purpose (see examples in Chapter 4). For instance, where the underlying survey regime is too geographically biased, it could skew the selection of protected areas towards places that have been well-surveyed but which are not particularly biodiverse. In data-poor areas, one alternative is to use environmental surrogates (e.g. vegetation types) as a ‘coarse filter’, with the aim of capturing biodiversity attributes that are likely to correlate with the chosen data layers (e.g. Faith & Walker, 1996). A limitation of such approaches is that unless very finescale environmental data are available, ‘fine filter’ features indicative of resource hotspots, such as saltlicks, are likely to be missed, as may be the factors controlling the distributions of the subset of threatened and rare species (see: Araújo et al., 2001, 2003, 2004a; Faith et al., 2004a; Noss, 2004). Below are some examples of different theoretical approaches in developing surrogates for conservation planning purposes. Species-based surrogates A variety of criteria have traditionally been used to select species-based surrogates in systematic conservation plans. These surrogates have often been called ‘indicator’ species and there are a number of different types that have been used in past planning techniques (see Box 6.1).

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Box 6.1 Some examples of species-based surrogates that have been used in systematic conservation planning approaches Keystone species have a disproportionate effect on the ecosystem relative to their abundance (Mills et al., 1993; Paine, 1995). As such, they affect the types and abundance of many other species in a community. The identification and management of these species can be important in conservation planning (Fleishman et al., 2000). The keystone concept, although intuitive, has received criticism because it is not always clear whether ecological communities have keystone species – and even if they do, this may be hard to demonstrate quantitatively because of the complexity of community structure and environmental dynamics of many ecosystems (e.g. Power et al., 1996; Andelman & Fagan, 2000).

Figure B6.1a The African elephant (Loxodonta africana) plays a significant role in altering the vegetation structure and type throughout its range, and as such is considered a keystone species. Photograph courtesy of Peter Baxter.

Focal species are, in the present context, species that are most endangered by the threatening processes within a system (Lambeck, 1997; Watson et al., 2001). The logic of using a focal species is that if a conservation plan meets their minimum needs, they should capture the needs of all the other species in that system in relation to that particular threat. This approach has, however, been criticized by Lindenmayer et al. (2002), who have argued: (i) that it may be too difficult to identify species most affected by each threatening process because of a lack of data on all taxa, and (ii) that the approach is over-reliant on the untested assumption that protecting the most threatened species will inevitably protect those that that are less threatened. Umbrella species are those species that are used as surrogates to represent the ‘health’ of an ecosystem or the distribution patterns of other species; or they are species that require such extensive resources for their conservation that many other species will be protected by default. Top predators are often used as umbrella species. There has been mixed support for the umbrella species concept in conservation planning. Andelman and Fagan (2000), in a study of umbrella species of the southern Californian sage-shrub community, found that selecting areas using umbrella surrogates performed barely better than a randomly selected set of species. However, Fleishman et al. (2001) have reported more positive results.

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Figure B6.1b The hooded robin (Melanodryas cucullata) has been identified as a focal species in the woodland ecosystems of south-eastern Australia as it is highly threatened by habitat fragmentation and requires large woodlands remnants that are close together to persist (Watson et al., 2001). Photograph courtesy of Mat Gilfedder.

Figure B6.1c The tiger (Panthera tigris) is often used as an umbrella species for conservation planning in countries such as India. Photograph courtesy of Liana Joseph.

Threatened taxa. It has been argued that conservation planning should concentrate on the needs of species currently endangered or threatened with extinction (Sarakinos et al., 2001; Conservation International, 2004). It is often less controversial to use these species as they should be of special concern for biodiversity conservation and, in some cases (and in particular regions), they may be relatively well known (and their locations mapped) (Gaston et al., 2002; Bottrill et al., 2009).

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Figure B6.1d The marine iguana (Amblyrhynchus cristatus) is found only on the Galapagos Islands. Uniquely among modern lizards, this animal lives and forages in the sea. It is threatened by predation by exotic species. Photograph: James Watson.

Phylogenetic difference. Some ecologists argue that species that are more phylogenetically distinct contribute more to the total genetic and morphological diversity and so should be given priority for protection (Weitzman, 1993; Faith et al., 2004b; Faith, 2009; and see Box 7.1). It has been suggested that a good way to generate a plan using this criterion is to use higher taxa (i.e. genus, family) instead of species in the planning process (Mooers, 2007).

Figure B6.1e The little known Guianan cock-of-the-rock (Rupicola rupicola) is a spectacular, phylogenetically distinct member of a two-species family inhabiting northern South America. Photograph: James Watson.

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Associated with the use of species surrogates (in particular using umbrella species, keystone species and focal species) to achieve representation is the concept of achieving functional or ecological redundancy, which refers to the situation where there are multiple species within an ecosystem that play similar ecological roles. Achieving functional redundancy is seen as important objective, because the consequences of losing all of the species that perform a particular ecological function within an ecosystem (e.g. losing all of the algae feeders from a coral reef) could result in a dramatic shift to a lower biodiversity system. Thus, the amount of functional redundancy in a system is of considerable importance in terms of retaining ecosystem integrity (Walker, 1992). This approach requires that species with similar ecological roles (termed functional groups) are identified, along with the key processes that maintain ecological integrity. Conservation efforts can then be aimed at maintaining a full suite of functional groups, and functional groups with little or no redundancy should be prioritized for conservation action (Walker, 1992). The functional redundancy concept has been enthusiastically applied to the problem of conserving coral reef ecosystems (Steneck & Dethier, 1994; Bellwood et al., 2004). Coral reefs are particularly prone to collapsing from a high diversity system to a low diversity system dominated by algae and a few species of fish (Scheffer et al., 2001) although, until recently, the causes of such phase shifts were poorly understood. Recent comparative analysis of functional groups in coral reefs from around the world strongly suggests that high species diversity provides the potential for functional redundancy (Bellwood et al., 2004). Hence, Caribbean reefs that are lacking several critical functional groups, or have groups represented by a small number of species, have been particularly prone to phase shifts to low diversity systems (Scheffer et al., 2001). However, it should be noted that even in high diversity coral reef systems, such as the Great Barrier Reef in Australia, there are still some functional groups with low redundancy (e.g. that are represented by a small number of species) (Bellwood et al., 2003). Despite the support of concepts such as functional redundancy by some systematic conservation planners, the overall level of support for species-based surrogates has been variable (Beger et al., 2003, 2007; Faith et al., 2004a). Since it is unlikely that it will ever be possible to measure the true variation of

biodiversity within or between regions, or the overall functional role played by all species in a region, the true effectiveness of a species-based surrogate is indeterminable. Moreover, the underlying assumption that the needs of a particular surrogate group of indicator species will ensure the long-term persistence of all of biodiversity may never be true as all individual species, have, by definition, evolved to have their own specialized needs (see discussion on individualism in Chapter 3) and these needs will never be captured by a surrogate. Because of this, many recent conservation planning exercises have used sets of species covering entire taxa (i.e. all birds, all mammals, etc.), or assemblages of species in a given area (e.g. combining plant, vertebrate and invertebrate data), as a surrogate for biodiversity in developing a conservation plan (Chapter 5, and see, for example, Williams et al., 1996; Sarakinos et al., 2001). In the case study outlined in Box 6.2, 53 species were identified that, when taken together, were considered representative of the system in Maputaland. These data were then combined with other data layers in a systematic conservation planning exercise. Environmental surrogates In the last decade, systematic conservation planning studies have predominantly used environmental surrogates as general surrogates for biodiversity representation (e.g. Carwardine et al., 2008; Klein et al., 2008). ‘Environmental surrogate’ is a generic term covering land or ecological classifications based primarily on physical and climatic variables, which can incorporate some biotic variables, such as vegetation type (Margules & Sarkar, 2007). These variables are assumed to correlate with the patterns of species distribution, and have been argued by some to be more useful than species-based surrogates (compare: Araújo et al., 2001, 2003, 2004a; Ferrier, 2002; Lombard et al., 2003; Faith et al., 2004a). Environmental surrogates are often used because these data are usually more readily available compared to more detailed biological data. In the Californian marine case study outlined in Box 6.3, a number of key habitats and a range of different depth classes were considered good environmental surrogates. In the Maputuland case study outlined in Box 6.2, it was argued that capturing the 44 land-cover types, as well as the 53 species, was the most effective way to get a representative system.

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Box 6.2 Conducting systematic conservation planning in the terrestrial environment: a Maputaland case study Written by Robert J. Smith, Durrell Institute of Conservation and Ecology, University of Kent, drawing on Smith et al. (2008). The Maputaland Centre of Endemism is a region of high conservation value that falls within the countries of Mozambique, South Africa and Swaziland (Figure B6.2a). Its climate and soils have played a large role in maintaining high levels of species richness and endemism (Steenkamp et al., 2004), but they have also influenced the conservation of this biodiversity, because much of the land has little agricultural value and so has not been cleared by commercial farmers. Instead, most people rely on small-scale farming and harvesting natural resources for their livelihoods. This, together with an increasing human population and a history of political marginalization, has led to widespread poverty. The governments of the region are keen to reduce these poverty levels and have recognized that ecotourism and game ranching are the most profitable forms of land use. Consequently, they have developed the Lubombo Transfrontier Conservation Area (TFCA) initiative, which seeks to conserve the region’s biodiversity and reconnect important large mammal populations while creating jobs by developing new conservation areas, both privately and communally managed. The TFCA initiative is guided by the Maputaland conservation planning system, which is based on systematic conservation planning principles (Margules & Pressey, 2000). This approach involves producing a list of important conservation features, setting targets for each feature and then identifying priority areas for meeting these targets.

Figure B6.2 (a) Protected areas (PA) and TFCA (Transfrontier Conservation Area) zones in the Maputaland Centre of Endemism. (b) Priority areas for conservation outside the existing protected areas. (c) Proposed conservation landscape. From Smith et al., 2008, with permission from Elsevier.

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The Maputaland system involved identifying 44 land-cover types, 53 species and 14 ecological processes as important conservation features, and mapping their distributions using satellite imagery and expert opinion (Smith et al., 2008). It also involved using data on the predicted spread of subsistence agriculture as a measure of both threat and opportunity cost, together with data on potential revenue from game ranching, which has key relevance for implementing the results (Knight et al., 2006). The first conservation assessment used the Marxan conservation planning software, which uses a simulated annealing approach (Ball & Possingham, 2000). This involved: 1 dividing the region into a number of planning units; 2 assigning a cost to each planning unit based on its modelled risk of being cleared for agriculture; 3 using Marxan to identify near-optimal portfolios of these units for meeting the targets, maintaining connectivity and minimizing impacts on subsistence agriculture (Figure B6.2b). These initial outputs were then used to develop a conservation landscape plan that could boost economic development through nature-based tourism and game ranching. The analysis identified 4,291 km2 of new core protected areas and 480 km2 of land that would function as ecological linkages (Figure B6.2c). The game ranching data were then used to estimate potential revenue from this proposed expansion of the protected area system. The results showed that these new areas could provide US$18.8 million per annum, thereby helping to create jobs and reduce poverty. These results have already been used to guide South Africa’s National Protected Area Expansion Strategy and the Critical Ecosystem Partnership Fund initiative in the Maputaland–Pondoland–Albany hotspot, although more work is needed to ensure that the system becomes part of day-to-day land use planning in all three countries.

6.4.2 Achieving persistence Identifying how to secure the long-term persistence of species, ecosystems and the ecological and evolutionary processes that maintain them is difficult. For most systematic conservation plans, persistence objectives are formed as targets. These targets should be informed by ecological theory and empirical knowledge of species autoecology and biogeography (Carwardine et al., 2009). The research that went into designing a conservation plan for the critically endangered Leadbeater’s possum (Gymnobelidius leadbeateri) is a good example of how an objective for persistence can be calculated using a species minimum viable area. This possum, considered an umbrella species (see Box 6.1), inhabits the tall forests of southern Victoria, Australia, but its habitat has receded due to industrial logging and changed fire regimes. Lindenmayer and Possingham (1995) showed that the species needed several patches, each of at least 100 ha in size, in each forest catchment in which they were present, to ensure their persistence in the long term. It was argued that all remaining patches of

habitat containing this species must be protected and, if possible, enlarged by restoration activities to hit this minimum viable patch size, which has been the basis of conservation plans in the region. In a similar example, Carroll et al. (2003) developed a conservation plan based on the needs of mammalian carnivores in the Rocky Mountains region of North America, using a spatially explicit population model that informed the design of the protected area network. Persistence targets can also be set for environmental surrogates, especially when planning at coarser spatial scales. These are often based on achieving representational targets for biodiversity features while implicitly accounting for consequences for other stakeholders (e.g. agriculturalists or the forestry sector). For example, in a series of Regional Forestry Agreements developed in Australia, it was agreed by all stakeholders, including conservation biologists, that each distinct forest type was adequately protected if at least 15 per cent of its area was within a protected area (Pressey, 1998). In the Californian marine case study outlined in Box 6.3, different persistence targets based on

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Box 6.3 Conducting systematic conservation planning in the marine environment: a case study from the central coast of California Written by Carissa J. Klein (The University of Queensland, Australia). California’s Marine Life Protection Act mandates the design and management of a network of marine protected areas to protect marine life, habitats, ecosystems and natural heritage, and to improve recreational, educational and study opportunities provided by marine ecosystems (State of California 1999). As part of the initiative to implement the Marine Life Protection Act, California’s central coast (from Pigeon Point to Point Conception, covering an area of 2,978 km2) was the first of five regions to undergo a stakeholder-driven process to design a network of marine protected areas. To help inform the design of marine protected areas consistent with the Act’s goals, a representative group of stakeholders from California’s central coast developed a very broad set of Regional Goals and Objectives with the help of administrators, managers, and scientists in the period 2005–2006. A scientific advisory team was then tasked to provide guidelines that quantified the science-related Regional Goals and Objectives. These guidelines were as follows: 1 The diversity of species, habitats, and human uses prevents a single optimum network design. 2 Every ‘key’ marine habitat should be represented in the network. 3 Protected areas should extend from the intertidal zone to deep waters. 4 Protected areas should have an alongshore span of 5–20 km. 5 Protected areas should be placed within 50–100 km of each other. 6 Each ‘key’ habitat should be replicated at least 3–5 times. 7 Placement should take into account local resource use and stakeholder activities. 8 Placement should take into account the adjacent terrestrial environment and associated human activities. 9 Network design should account for the need to evaluate and monitor biological changes within the protected areas. Systematic conservation planners were asked to produce a network of marine protected areas consistent with the scientific guidelines. These planners decided that it could be accomplished using the systematic conservation planning decision support tool Marxan (see Klein et al., 2008a,b and www.uq.edu.au/marxan/). Marxan identifies possible locations for protection that achieve a set of conservation targets (e.g. protect 20 per cent of each habitat type, 50 per cent of threatened species’ distributions) for a minimal ‘cost’ (e.g. cost of closing conservation areas to fishermen; see Box 6.4 for more information on what a minimal-set problem is). The nine guidelines outlined above, with the exception of 8 and 9, were able to be factored into the Marxan analysis as follows: • Guideline 1, which is related to the systematic conservation planning principle of flexibility, was accounted for by using Marxan to produce a number of different reserve networks which achieved a similar objective. Marxan produces multiple different solutions for the location of protected areas, all of which achieve the same set of conservation goals. • Guideline 2, which is related to the principle of representation, was addressed by representing each key habitat identified in the different reserve networks. • Guideline 3, which is also related to the principle of representation, was addressed by targeting each feature in five different depth zones. • Guidelines 4, 5, and 6, which are related to the principle of persistence, were addressed by employing user-defined parameters to ensure reserves were of an adequate size, spacing, and replication. • Guideline 7, which is related to the principle of efficiency, was addressed by minimizing the impact on commercial and recreational fisheries.

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Figure B6.3a Comparison of the impact on commercial and recreational fisheries of marine reserve networks designed by stakeholders, based on expert judgment, with that designed using Marxan. The fishing impact (defined as loss of overall fishing yield) of both solutions is displayed, and the Marxan analysis is defined as an average ± standard deviation of 100 different solutions that achieved the planning objectives. Network 1 was developed by commercial and recreational fishers, network 2 by conservationists, network 3 by a mixed interest group, and network 4 was the solution considered for implementation by The California Fish and Game Commission. Adapted from Klein et al. (2008b).

Biodiversity data used in the analysis were environmental surrogates which included rocky reef, soft bottom, kelp forests, submarine canyons, eelgrass, surfgrass, and estuaries. Each of these habitats was targeted for inclusion in a network of marine protected areas under a number of different scenarios (e.g. 10 per cent representation, 20 per cent representation and 50 per cent representation in the reserve system). The socio-economic data included information on the number of recreational fishing trips and an expert-derived assessment of the relative importance of an area for commercial fishing. The expertdriven assessment involved 109 commercial fishermen being interviewed to determine accurate spatial data on fishing effort and to map their fishing grounds. From this, an index of relative fishing effort was used to calculate the impact of fisheries in the reserve design (i.e. those marine waters that would be closed to fishing). The two types of fishing data were combined to deliver a relative index of fishing effort, which was used as a ‘cost’ to minimize in the Marxan software. Marine reserves were chosen that would meet the different biodiversity targets and minimize the impact on fishermen in terms of lost fishing effort due to reservation. Explicitly considering commercial and recreational fisheries in the analysis allowed the impact to the fisheries to be reduced by up to 21 per cent, depending on the scenario selected (Klein et al., 2008a). In a separate analysis, Klein et al. (2008b) were able to compare the marine reserve network designed without using a systematic planning tool by the three stakeholder groups (commercial and recreational fishermen, conservationists and a mixed interest group) against those designed using Marxan. They found that the Marxan analysis represented an equal or greater amount of habitat, yet for a lower cost in terms of the impact on commercial and recreational fisheries (Figure B6.3a). Interestingly, of all stakeholder groups, the proposal developed by stakeholders from the fishing industry was the most proficient at representing biodiversity and minimizing the impact to the fishing industry. These results indicate the important role stakeholders have in systematic conservation planning and that conservation planning decision support tools should be used to support stakeholder-driven planning processes, not replace them.

The distribution of diversity: challenges and applications environmental surrogates were used. The research team formed a scientific advisory team that gave them advice on what would be a good target for reservation for each habitat and water depth class. They were advised that the key habitats and different depth classes had to be captured and replicated at least three to five times to achieve an adequate outcome. See Box 6.3 for the results of this exercise. Despite their continued use, there has been a large amount of criticism over the use of simple percentage targets in systematic conservation plans (Soulé & Terborgh, 1999; Recher, 2004; Watson et al., 2008). The main criticism is that fixed percentages do not account for landscape context. The habitat fragmentation literature (Chapter 8; and see Lindenmayer & Fischer, 2006) reveals that the size and isolation of the protected area, its ‘shape’ in terms of edge to core ratio, and also the similarity (or ‘hostility’) of the matrix habitat surrounding the protected area, can each affect the chances of persistence for many species. Fixed percentage targets do not take these patch- and landscape-scale effects into account. There have been a number of recent analyses in the systematic conservation literature to address this problem. Specific design criteria based on the characteristics of environmental surrogates (e.g. a specific habitat type) have been incorporated into the persistence objective in some systematic conservation plans. For example, Leroux et al. (2007) introduced a framework for determining a minimum reserve size required to incorporate natural disturbance and maintain ecological processes by identifying criteria for estimating the size, location, and efficacy of a minimum dynamic reserve. The size and location of such a reserve is determined by the estimated maximum extent of the largest disturbance event, and by the extent and distribution of communities of species that are differentially affected by disturbance. They illustrated their approach using a study of the Mackenzie Valley region of Canada, where forest fire is the major natural disturbance that influences vegetation community dynamics and dependent fauna. In this research, Leroux et al. (2007) designed and evaluated a candidate minimum dynamic reserve using a spatially explicit dynamic simulation model that incorporates locally calibrated fire and the vegetation dynamics (i.e. the minimum area they need for persistence and recolonization following a fire event) of five broad vegetation types (closed spruce, open spruce, mixed-wood, tall shrub, small shrub).

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Using simulations, they showed that minimum extent of vegetation types ranged from 10 km2 of tall shrub to 1,948 km2 of open spruce, while the mean extent of the five communities available to burn in the study area varied from 118 km2 of mixed-wood to 3,407 km2 of open spruce. Using these thresholds, they showed that their minimum dynamic reserve maintained its recolonization sources through time, suggesting that minimum dynamic reserves may provide an operational framework for determining reserve size in dynamic landscapes under the influence of large natural disturbances such as fire. Of course, it is very difficult to validate such an estimation – hence the use of simulations. Another way in which systematic conservation planners have attempted to achieve persistence is to develop a form of redundancy within the plan, i.e. to set multiple representation targets. Here, the idea is that reserve planning algorithms are set with the goal of selecting a network of areas that ensures, for example, that each species occurs in a minimum of five separate sites. Building in this degree of redundancy may be desirable to provide the protected area networks with a degree of resilience to ensure that a species (or other desired biodiversity attribute) survives in the face of natural catastrophes, disease epidemics, the chronic ecological and genetic effects of small population size, or the loss of a reserve to legal or illegal human intervention. It should be noted that this use of the term ‘redundancy’ has somewhat negative connotations in conservation planning, as it was used as a key theme in criticisms of formerly widely used scoring procedures that disregarded complementarity and yielded systems of protected areas that had high redundancy and were inefficient (i.e. they were expensive and achieved few targets – see Pressey & Nichols, 1989; Pressey, 1994). Thus, the term redundancy is rarely used and ‘multiple representation’ is the favoured expression. This is considered more appropriate because multiple representations are not a by-product of the selection process but, rather, they are actively pursued. Rodrigues et al. (2000) provide a useful demonstration of the potential advantages of multiple representations. They used presence/absence data from the Common Birds Census (CBC) in the UK to test the effectiveness of three families of selection models: i Single and multiple representations. Single representations calculated the minimum area such that each species was represented in at least one site. The

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multiple representations method selected the minimum area needed to ensure that each species was represented in at least n sites (or the maximum number of sites, if this was less than n). ii Percentage of range. This method was used to select the minimum area of sites so that each species was represented in at least p per cent of its range within the study area. iii Permanence rate. A permanence rate was calculated for each species in each site, being the frequency with which a species was recorded in relation to the number of visits to a site within a specified time period. The minimum area was selected so that each species was represented in the site, or one of the sites, where it has the highest permanence rate. The results of the Rodrigues et al. (2000) study clearly demonstrated that a single representation strategy (a minimum of one site containing each species) leads to very high efficiency but low longterm effectiveness. A multiple representation strategy appeared to be safer than a strategy based on percentage of area. This is explained by the prioritization of rare species that is an inevitable by-product of the multiple representation approach. For example, if a rare species only occurs in three sites and the multiple representation criterion (n) is set to three sites or more, then all the sites containing the species necessarily will be included in the selection. The drawback of a simple multiple representation approach is that it assumes that all sites where the species occur have a similar potential for sustaining a population over a period of time. Strategies that target sites where species are most likely to persist give the greatest probability of long-term effectiveness (Williams, 1998). Unsurprisingly then, the Rodrigues et al. (2000) study found that choosing the best site based on permanence rate was a better strategy than investing in multiple, but blind, redundancy. Unfortunately, estimating persistence rate requires a lengthy and accurate time series, and other methods of choosing the ‘best’ site such as using abundance data are also expensive and time-consuming. Ultimately, the decision about whether built-in redundancy is a good way to select a reserve network depends on data and resource availability (e.g. what area/pattern of reserves can be maintained). An additional approach beyond planning for multiple representation is to plan ways to maximize the biophysical connections among protected areas. This is considered important for a number of reasons:

• First, as natural landscapes become more fragmented, an increasing number of species will need to disperse through an increasingly ‘hostile’ landscape matrix if they are to maintain their genetic variability in viable metapopulations. It is probable that connected landscapes improve the chances of this happening (Mackey et al., 2008). • Second, it is increasingly recognized that a large number of species need a very large area to survive – far larger than a protected area network will provide. For example, the European goshawk (Accipter principalis) has a home range of 30–50 km2, and male mountain lions (Felis concolor) in the western United States have home ranges in excess of 400 km2 (Wilcove et al., 1986). Moreover, many species have evolved to be highly dispersive and regularly migrate vast distances to find suitable conditions. These species clearly require more space than could reasonably occur in a small number of isolated protected areas, as the resources they require for existence vary both spatially and temporally (Gilmore et al. 2007). The survival of these species will depend on their ability to move between protected areas, and also the hostility of the matrix habitat between protected areas. • Third, habitat connectivity is likely to play an even larger role with the onset of anthropogenic climate change. Studies have estimated that by the middle of the 21st century, range shifts due to climate change will commonly span tens of kilometres (Kappelle et al., 1999). There will be a clear need to have some form of connectivity to find suitable locations to which species can migrate or take refuge (Peters & Darling, 1985; Mackey et al., 2008). Planning for ‘connectivity’ has recently moved beyond simply creating corridors or stepping-stones between protected habitat patches. The concept of connectivity conservation is now encompassed within the concept of maintaining the ecological and evolutionary processes that generate and sustain biodiversity at various spatial and temporal scales (Soulé et al., 2004, Pressey et al., 2007; Watson et al., 2009). Incorporating information on connectivity within a systematic conservation planning framework enables networks of priority areas to be designed with the goal of maintaining genetic and demographical flows, which may thus ensure the resilience of populations to the effects of landscape conversion and climate change. To date, few studies have incorporated ecological and evolutionary processes into conservation planning

The distribution of diversity: challenges and applications (Rouget et al., 2003; Possingham et al., 2005; Pressey et al., 2007). However, in a national scale analysis in Australia, Klein et al. (2009) accommodated ecological and evolutionary processes in four ways: 1 using sub-catchments as planning units rather than arbitrarily delineated grids; 2 targeting refugia from drought; 3 targeting evolutionary refugia; 4 preferentially selecting planning units along connected waterways. The researchers identified drought refugia as areas with relatively high and regular herbage production, while evolutionary refugia were identified as areas thought to be important for maintaining and generating biota during long-term climatic changes. They identified priority areas for conservation in Australia that met biodiversity and ecological process targets while minimizing acquisition cost. Other examples of incorporating ecological processes in conservation planning include the comprehensive analyses undertaken in South Africa, where spatial surrogates for processes, such as edaphic interfaces, animal movement corridors, and macroclimatic and environmental gradients were targeted (Cowling et al., 1999, 2003; Rouget et al., 2003, 2006). Clearly, the dynamic nature of ecological processes makes them difficult to quantify (Possingham et al., 2005), but they are now recognized as an important consideration when persistence objectives are being defined. See further discussion in Chapter 7.

6.4.3 Achieving efficiency As discussed in section 6.2, a key concept in identifying areas to achieve representation efficiently is complementarity. The basic idea behind complementarity is that conservation areas should complement one another in terms of the ‘features’ they contain, the species, communities, habitats, ecological processes, etc. Each conservation area should be as different from the others as possible until all the ‘differences’ (e.g. different species, communities, etc.) are adequately represented. Complementarity can be defined in a number of ways. The most commonly used implementation is that a proposed new conservation area is assigned a higher complementarity value than another if it has more surrogates that have not already met their assigned target of representation in a conservation area

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network. For example, two proposed new areas, both with high species richness, may have different numbers of surrogates that can be captured in the reserve network. The efficient choice would be selecting the area that adds the most complementarity. Complementarity is, therefore, related to the concept of beta diversity (Whittaker, 1972), but whereas beta diversity is the difference between two areas, complementarity is a measure of the dissimilarity between the species complements of sets of selected areas. It is important to note that the principle of efficiency is not simply about achieving complementarity. As we discussed earlier, achieving an efficient network is also a matter of achieving objectives for the least possible cost, where cost may reflect the financial cost of implementing and managing protected areas or the costs of lost opportunities for economic development (Naidoo et al., 2006). There is an increasing number of examples of where cost data have been implemented into systematic conservation analyses. For example, in the Californian marine case study outline in Box 6.3, the authors conducted and interviewed 109 commercial fishermen to find spatial data on fishing effort and to map their fishing grounds. From this, an index of relative fishing effort was used to calculate the impact of fisheries in the reserve design (i.e. those marine waters that would be closed to fishing). Using these stakeholder data, Klein et al. (2008a,b) were able to produce a systematic conservation plan that was efficient in that it maximized biodiversity conservation and minimized cost to livelihoods for fisherman. As outlined in Section 6.3.3, it is also possible to factor in the returns from ecosystem service protection into conservation planning analyses. There may, however, be trade-offs between the achievement of objectives (Mertz et al., 2007; K.A. Wilson et al., 2009), depending on the spatial congruence between ecosystem services and between ecosystem services and biodiversity features. Some analyses have found high levels of congruence (Turner et al., 2007; Venter et al., 2009), but in other areas overlap has been more limited (Chan et al., 2006; Naidoo et al., 2008). There are several ways to integrate ecosystem services into conservation planning analyses (Egoh et al., 2007). Ecosystem services can be included as a feature for which a target can be set (Chan et al., 2006) and the set of planning units that meet these and other targets for the lowest cost can be identified. Alternatively, it is possible to modify the relative weighting for

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conserving ecosystem services versus the conservation of other biodiversity features (K.A. Wilson et al., 2009) and then seek to maximize the overall benefit that is derived.

6.4.4 Achieving flexibility As we have discussed throughout this chapter, the selection and creation of new protected areas in a network is not a simplistic, one-off process. Protected area networks have to be accepted socially and politically, and it is therefore of critical importance that there should be several alternatives available when a systematic conservation plan is developed. These alternatives mean that the plan is flexible (Pressey et al., 1993). It must be clear, however, why areas are selected and why some areas are not, and hence transparency is a clear part of flexibility (Nicholls & Margules, 1993). Measuring the ‘irreplaceability’ of sites is arguably the commonest way to show flexibility in a systematic conservation plan. The irreplaceability of a site reflects the importance of including that site in the protected area network if all conservation objectives are to be achieved (Pressey et al., 1994; Ferrier et al., 2000). Irreplaceability can be viewed in two contexts: the likelihood that an area is necessary to achieve conservation objectives for the features it contains; or the extent to which the options for achieving conservation objectives are reduced if the area is unavailable for conservation. In systematic conservation planning, a completely irreplaceable area is essential for a plan to meet its conservation objectives, whereas an area with a very low irreplaceability can be substituted by other sites. For example, when planning a reserve system in a landscape, you may find that some areas are completely unique or have been altered to such as extent that the last remaining sites are highly irreplaceable. If there is a risk of these areas being lost to threatening processes, then it might be a large loss for biodiversity conservation in that region. Consequently, irreplaceability can be used as a measurement of conservation value. It is important to note that although irreplaceability can help determine which areas are priorities for conservation, other constraints and considerations may mean that areas with lower irreplaceability are more suitable for conservation. For example, some

combination of vulnerability, ecological condition, and financial cost of an area might influence its priority for protection. When this occurs, it is important to acknowledge the conservation cost of not including these sites within the overall plan. Moreover, the irreplaceability rank of an area will change as individual areas are designated as part of the conservation area network. Therefore, the process of identifying irreplaceable sites must be reiterated after each stage, when new areas are included in a network and others are removed. Such a process was involved in the Maputaland study highlighted in Box 6.2.

6. 5 DECI S I ON S U PPOR T T OOLS T O I DEN T I FY AN D PR I OR I T I Z E N EW PR OT ECT ED AR EAS As discussed in the introduction to this chapter, the development and use of systematic planning tools for designing protected areas is only a recent phenomenon. The early approaches to designing systematic conservation plans using simple scoring systems (e.g. Margules & Usher, 1981; Smith & Theberge, 1986) were perceived to be a great improvement on previous approaches due to their transparency and repeatability. However, due in large part to technical limitations of data processing up until the end of the 1980s, these early systematic conservation planning approaches did not take into consideration complementarity, nor did they have the ability to set spatial objectives like connectivity and spatial compactness (Margules et al., 1991; Pressey, 1997). Over the past decade, decision support tools have been increasingly used to help inform conservation planning decisions. Decision support tools are often computer-based information systems intended to help decision-makers compile and analyse data to help solve conservation problems. A range of mathematical techniques have been developed that are incorporated into these tools (see Box 6.4 and Moilanen et al., 2009). It is important to note that the use of any decision support tool, simple or complex, requires properly defined conservation problems. A common framework for defining conservation priorities is through the use of decision theory. This framework centres on achieving explicitly stated objectives while acknowledging constraints on conservation actions and the levels of uncertainty involved within the decision process. In Table 6.2, we outline a

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Box 6.4 Three broad classes of mathematical problems used in systematic conservation planning: the minimum set problem, the maximal coverage conservation prioritization problem and the conservation resource allocation problem Conservation planning began without a well-posed mathematical problem, which is not uncommon in conservation science (Possingham et al., 2001). Cocks and Baird’s (1989) seminal paper provided the first formal statement of a conservation planning problem – the minimum set problem. In the minimum set problem the goal is to conserve a variety of conservation features to an adequate level for minimum total cost where cost can be the cost of acquisition and management or the estimated foregone opportunity cost (Naidoo & Ricketts, 2006). The simplest variant of this problem is: NS

min ∑ ci x i i =1

given that Ns

∑x r

i ij

≥ T j,

for all features j,

i =1

where rij is the occurrence level of feature j in site i, ci is cost of site i, Ns is the total number of sites and Tj is the target level for feature j. The control variable xi has value 1 for selected sites and value 0 for sites not selected (Moilanen et al., 2009). This became the foundational problem of systematic conservation planning. Since then, various authors have produced alternatives, but arguably the maximal coverage conservation prioritization problem is the most dominant. This problem is used when resources are insufficient for satisfying all targets and the objective is to find the solution that satisfies the largest number of conservation targets, given a budget constraint. The maximal coverage problem is related to the minimum set coverage problem, in that minimum set coverage can be achieved by solving the maximal coverage problem at different budget levels and finding the minimum budget level that satisfies all targets. A simple version of the maximal coverage problem can be written as: max ∑ I j ( ∑ x i rij ≥ T j ), j

i

given that

∑xc i

i

≤ B,

i

where B is the conservation budget (money, trained personnel, time, etc.), and I(z) is an indicator function, with Ij(z) = 1 when condition z is true, i.e. the target for feature j is met when ⎛ ⎞ x i rij ≥ Tj ⎟ , ⎝⎜ ∑ ⎠ i

and Ij(z) = 0 otherwise (Moilanen et al., 2009). Both the minimum set and maximum coverage problems are limited to specific problems. However, it is possible to define a fairly general conservation resource allocation problem that includes most, if not all, previous problem definitions. In general, all of conservation involves taking actions in a place and at a time in an attempt to achieve a variety of outcomes. Our general task is therefore to decide how much to spend on each kind of action (e.g. invasive species control, changed logging practices, or reduced grazing) in each place, at a particular time

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(which we will refer to henceforth as the year). These actions will alter the dynamics of a variety of state variables, yijt., such as the size of the population of a species in a site, or the amount of an ecosystem service in a site. Mathematically this means that our control, or decision, variable is ajkt, the amount of money we spend on action k in place j in year t. A fairly general formulation of the conservation resource allocation problem is to: T

max ∑ f ( y ijt ,1 ) t =1

subject to a budgetary constraint each year N

P

∑∑ a

jkt

≤ bt for all t,

j =1 k =1

and contingent on the dynamics of the state variables, that is, how key system states move from year to year in response to actions or forces we do not control: y ijt . 1 ⋅ g( a..t y..t, x..t,) for all i, j and t, where f is a function that turns our state variables into a reward function that we are trying to maximize (this could be highly non-linear), g is a function that determines how the state variables, yijt, evolve in space and time as a consequence of actions and forces we do not control, xijt. In this formulation, N is the number of different places and P is the number of different sorts of actions. This mathematical formulation of a problem that considers expenditure of money on different conservation actions in space and time is a fairly general formulation of all resource allocation problems. It is called a resource allocation problem because there is a fixed annual budget. Evaluating actions based on their cost-effectiveness (Joseph et al., 2009) provides one algorithm that can often provide rough solutions to this very complex optimization problem.

seven-step decision theory framework which has been articulated by a number of authors for systematic conservation planning (Table 6.2; Possingham et al., 2001; K.A. Wilson et al., 2009). There is now a large amount of literature on optimal protected area design based on this decision theory framework (summarized in Moilanen et al., 2009). The problems generated using this framework can be expressed mathematically and then solved by one of a number of methods. There are two classic problem definitions commonly used in conservation planning, the minimum set and maximal coverage conservation prioritization problem (Box 6.4). The minimum set problem minimizes the resources expended while meeting the conservation objectives. For this problem, the objective is to minimize cost and the constraint is the conservation objectives. The maximal coverage conservation prioritization problem maximizes the objectives (e.g. target level achievement) given a fixed amount of resources. Here,

the problem is reversed: the constraint is the budget and the objective is to maximize conservation objectives. Methods for solving systematic conservation planning problems fall into several classes: local heuristic algorithms, which select sites in a stepwise manner (Pressey et al., 1993; 1994); global heuristic algorithms, which select sites in sets (e.g. simulated annealing, Ball & Possingham 2000); and optimization algorithms (Cocks & Baird, 1989). These methods are dealing with increasingly large and more complex problems (see section 6.7), which includes having multiple and conflicting objectives and multiple types of management actions. It must be noted that decision problems can be quite complex, and there are now several software packages that can support systematic conservation planning (e.g. Marxan, C-Plan, Zonation, ConsNet; see Moilanen et al. (2009) for a thorough review of each platform). However, as Bottrill & Pressey (2009) point out, these

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Table 6.2 The application of a seven-step systematic conservation planning decision theory framework (Possingham et al., 2001; K.A. Wilson et al., 2009) to a hypothetical example based on the problem of acquiring new land to add to a protected area network to protect threatened species.

Step

Details

Example: Acquiring new land to add to a protected area network with the aim of protecting threatened species.

1 Statement of objective(s)

This is a statement of what is hoped to be achieved and is measurable.

To maximize the representation of threatened species in protected areas.

2 List of management actions

This can range from one action to a number of actions.

Purchasing new areas to add to the protected area network. The available option is either to acquire each parcel of land or not.

3 State variables

This is the knowledge about the system, including both biodiversity and human variables.

Where the threatened species are located and how much each parcel of land costs.

4 State dynamics

This step requires knowledge about how the state variables may change (which may be dependent or independent of the management action).

Fluctuation of property prices for parcels of land. These may vary independently or may increase with the implementation of the extended reserve network (Armsworth et al., 2006).

5 Constraints

The constraints are what limit the application of any management action.

Size of budget, willingness of landholders to sell their properties, etc.

6 Uncertainty

Most data will contain a degree of uncertainty.

Inaccuracies in species data regarding presence and absence due to surveying methods and species detectability variation.

7 Solution methods

A range of mathematical approaches are used to solve problems (Box 6.4).

Algorithm to maximize representation and minimize cost.

software systems are not designed to replace people by making decisions for them; they operate interactively to facilitate decisions by people. 6. 6 CO N SUL T AT I ON AND I MPL EM E NT AT I ON OF S YS T E M A T I C CO N S ER V AT I ON P L ANS Much of the systematic conservation planning literature to date has focused on advancing the ‘tools’ of the systematic conservation planning trade. Far less attention has been dedicated to implementing conservation plans in the ‘real world’ (Salafsky et al., 2002; Knight & Cowling, 2003). Indeed, some experts have argued that the discipline of systematic conservation planning is mired in an ‘implementation crisis’ (Knight &

Cowling, 2003), because ‘… few academic conservation planners regularly climb down from their ivory towers to get their shoes muddy in the messy, political trenches, where conservation actually takes place’ (Knight et al., 2006, p. 410). There has been some critical discussion around this quite stark assertion (see, for example, Pressey & Bottrill, 2009), and a number of operational case studies show that development of a systematic conservation plan for a particular area by academics can integrate the diverse disciplines and activities needed for successful conservation action into a single, comprehensive process (Boxes 6.2 and 6.3 are good examples). Nonetheless, this debate highlights the point that while the tools of systematic conservation planning are important, they do not in themselves deliver conservation action.

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To be successfully implemented, all systematic conservation plans must be complemented with social, political, and institutional tools and processes (Knight et al., 2009). There are several operational models established that outline key considerations that can help to guide a transparent planning and implementation process (Pressey & Cowling, 2001; Knight et al., 2006). Those who have participated in such processes stress the importance of providing participants with clear and transparent explanations of the stages of a conservation planning process and what things need to be done to achieve them. This includes, for example, an assessment of conservation issues; identifying opportunities for and constraints on conservation actions; developing an implementation strategy; and products such as maps that help guide implementation. The flexibility of an overriding conservation model is also important, as it has been increasingly shown that implementing successful conservation actions is an ongoing process with feedbacks between planning and implementation. Bottrill & Pressey (2009), for example, have produced a detailed 11-step process for the successful implementation of a conservation plan. The steps are outlined in Table 6.3. They argue that all 11 stages must be completed for a conservation plan to be conducted successfully. An alternative operational framework is provided by Knight et al. (2006), who identified the key components for ‘doing’ pragmatic conservation planning (Figure 6.2). In their schematic, the thematic components are grouped into three interlinked foundations: 1 empower individuals and institutions; 2 undertake the systematic conservation assessment; 3 secure effective action. Each foundation is essential for an effective conservation planning process. The reality is that the implementation of any conservation plan is difficult. Wherever systematic plans are actually implemented, it quickly becomes apparent that human society is not an entity with a single value system (see Chapter 2). Whereas a conservationist or amateur naturalist may value a particular site because it contains habitat for an endangered species, a timber company may value that site because of the potential revenue that might be generated from harvesting trees, or a group of mountain bike enthusiasts may value the site for its recreational values.

Table 6.3 Steps in the process of developing and implementing a conservation plan, as outlined by Bottrill & Pressey (2009). Steps

Processes

Stage 1

Scoping and costing the planning process

Stage 2

Identifying and involving stakeholders

Stage 3

Identifying the context for conservation areas

Stage 4

Identifying conservation goals

Stage 5

Collecting socio-economic and threat data

Stage 6

Collecting data on biodiversity and other natural features

Stage 7

Setting conservation objectives

Stage 8

Reviewing objective achievement in existing conservation areas

Stage 9

Selecting additional conservation areas

Stage 10

Applying conservation actions to selected areas

Stage 11

Maintaining and monitoring established conservation areas

However, when a plan is integrated with expert knowledge (Pressey & Cowling, 2001) and coupled with an implementation strategy that takes into context the needs for stakeholder collaboration (Driver et al., 2003), the planning process itself can provide a foundation for effective conservation action. This is a major, often forgotten, value of systematic conservation planning – it not only identifies the priority conservation areas, but also provides a mechanism for stakeholder collaboration. 6. 7 W H AT DOES T H E FU T U R E OF S Y S T EMAT I C CON S ER V AT I ON PLAN N I N G H OLD? In this chapter, we have delved into the fundamental principles of systematic conservation planning, while also providing some contemporary case studies demonstrating the use of different techniques and tools. In

The distribution of diversity: challenges and applications

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Figure 6.2 An operational model for pragmatic conservation planning. From Knight et al., 2006.

this final section, we discuss future challenges in the field of systematic conservation planning and some recent advances. The first two decades of systematic conservation planning primarily focused on a restricted suite of problems. These problems have generally assumed that systematic conservation planning: 1 is a static problem that determines a once-off reserve system; 2 can ignore the dynamic nature (including evolution) of biodiversity assets (e.g. species, habitats); 3 assumes a binary world where sites are either protected or not;

4 can use the area or number of sites as a surrogate for cost; 5 can ignore uncertainty; 6 can ignore risk and threat, and 7 can rely on simple targets for biodiversity assets so that once achieved, we are content that persistence is achieved. All of these issues are challenges that need to be overcome if the discipline is to be taken seriously by those responsible for implementing conservation action. Here we briefly discuss some recent work that has contributed to this furthering this research agenda.

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6.7.1 Conservation planning is a dynamic problem

6.7.3 A mix of conservation actions could occur at any site

Possingham et al. (1993) provided one of the first analyses that formulated the dynamic site selection problem. In each year they assumed one site could be bought (due to a constrained budget), sites had a fixed probability of becoming available and sites that were unreserved had a fixed probability of being destroyed. At the time, these authors found that taking a static approach was suboptimal compared to solving the dynamic problem using stochastic dynamic programming. Various authors have subsequently considered and solved larger and more realistic versions of this original problem (e.g. Costello & Polasky, 2004; Meir et al., 2004; Drechsler, 2005; Strange et al., 2006). Such advances enable systematic conservation planners to include complexities like dynamic budgets and feedback between acquisition actions and the cost of reservation. In principle, any sort of dynamic complexity can be included in the site selection problem; however, the optimal solution of stochastic dynamic problems can only be found exactly using stochastic dynamic programming, which is computationally intractable for any but the smallest problem. There is therefore a need to derive simple heuristics that sequentially choose actions through time, such as choosing the actions that maximize the short term gain in biodiversity or minimize the short term loss of biodiversity.

As discussed briefly in the introduction, formal protection of habitat is just one of many conservation actions. In many cases, especially where there are multiple players in land ownership issues plus complex social and cultural constraints, reservation is an unlikely option for conservation. What we need is tools to help us determine which package of actions to activate at any site. This sort of idea is effectively zoning – a common practice in fisheries, forestry and conservation where there are multiple interests (Watts et al., 2009). These zoning tools are useful to guide broad policy decisions, and other methods have been developed to systematically select among specific conservation actions. For example, the Project Prioritization Protocol is a costeffectiveness analysis that has been demonstrated to be useful for selecting among specific management projects for threatened species in New Zealand (e.g. Joseph et al., 2009).

6.7.2 Conservation assets change through time The biodiversity assets that we would like to conserve are continually changing: local populations become extinct; species’ distributions change; species evolve; and vegetation types change through succession (as discussed in Chapter 3). This adds further complexity and uncertainty to the dynamic conservation planning problem described above, and in principle it can be dealt with within the same approach. However, there are some short cuts possible. Sites with evolutionary potential can be preferred in planning (Cowling et al., 2003), present and future predicted distributions can be accommodated in the plans (Hannah & Hansen, 2005) and successional changes can be predicted and allowed for in target setting (Drechsler et al., 2009).

6.7.4 Better economics and socio-economics Ando et al. (1998) were arguably the first to highlight in the peer-reviewed literature the naivety of building conservation plans that ignored realism in respect of financial costs. While the inclusion of the estimated cost of conservation is now more common in conservation planning (see section 6.3.3) it is still a challenge for most conservation researchers who are more familiar with the nuances of biological data (Bode et al., 2008). To this end, there is a need for more real collaboration between economists and conservation biologists. However, it is also being recognized that using simple cost-layer data (i.e. the price of land), without considering socio-economic factors such as a landholder’s willingness to conduct a conservation action, regardless of cost, may lead to some erroneous results.

6.7.5 Dealing with uncertainty There is some level of uncertainty in every aspect of conservation planning (Regan et al., 2009). For

The distribution of diversity: challenges and applications example, semantic uncertainty underpins the actual definition of the conservation problem, while parametric uncertainty is rife in all the data that are used to develop conservation plans (Whittaker et al., 2005). While Regan et al. (2005, 2009) argue that we can generally deal with parameter uncertainty quite well using sensitivity analysis, uncertainty about problem formulation or issues like species viability represent serious challenges at the interface of social science, philosophy, economics, mathematics and ecology. So far there are too few papers that deal credibly with uncertainty in conservation planning (but see Moilanen & Wintle, 2006).

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Probably the best way forward for conservation planners is to explicitly acknowledge and derive tradeoffs, recognizing that no single answer is best but offering a range of good options that reflect different societal aspirations (Whittaker et al., 2005; Polasky et al., 2008). An alternative might be to represent different levels of risk (e.g. 75 per cent, 80 per cent or 95 per cent probability of persistence for 100 years) or varying levels of persistence (80 per cent probability of persistence for 10, 100 or 1000 years) based upon available knowledge. Further discussion of the challenges of planning for persistence in a changing world is provided in the following chapter.

6.7.6 Properly accounting for threats There are several ways of dealing with threats in conservation planning. One is to rate sites in terms of the likelihood that they will be destroyed relative to their irreplaceability, with preference given to sites that are under most threat (Araújo et al., 2002a; Pressey et al., 2007). In practice, some planners use the likelihood of a site being converted to other uses as a surrogate of conservation cost and hence, by reference to the principle of efficiency, they avoid sites with a high probability of conversion. Ironically, this will give us the reverse outcome to the first approach. Indeed, some of the confusion about how we deal with threats arises because some threats are mitigated by conservation action, while others are not. Ideally, threats are dealt with properly in a full dynamic framework (Wilson et al., 2006; Game et al., 2008) within which the consequences of taking action at a site, or not, are explicitly modelled.

6.7.7 Persistence – attainable goal or impractical utopia? Persistence (also known as adequacy) is the bugbear of systematic conservation planning science because the question it asks – how much is enough? – is probably unanswerable. Governments and non-government organizations would often like to know that a suite of conservation actions in time and space is sufficient. However, in reality, more is always better, although that ‘more’ comes at an additional cost.

6.7.8 How much should we invest in improving a conservation plan? As we have discussed throughout this chapter, there are usually many assumptions about what the most appropriate conservation actions in any given area may be and whether the data are truly fit for purpose. Recent research has shown that if learning processes and data collection strategies are intentionally included into the conservation planning process, it is likely that future conservation decisions will become more effective (Grantham et al., 2009). There is a complex and not very well understood trade-off between acting and learning when developing and implementing a systematic conservation plan. It is important to recognize that any given planned conservation action has been traded off with all other actions and also against the cost of delaying a conservation action.

FOR DI S CU S S I ON 1 Describe and give examples for each of the key principles of systematic conservation planning. Describe some ways of achieving each of these principles when developing a hypothetical systematic conservation plan in both the marine and terrestrial environments. 2 How should scientists assess the fitness for purpose of data for use in systematic conservation planning?

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3 What are the limitations and strengths of using targets for achieving persistence in a conservation plan? 4 What are the respective strengths and weaknesses of using species-based surrogates versus environmentalbased surrogates when developing a systematic conservation plan? 5 What is the difference between conducting a minimum set problem as opposed to a maximum coverage problem when undertaking a systematic planning process? Give some examples of when each type of problem should be applied. 6 Are the key principles of systematic conservation relevant to both marine and terrestrial environments? What differences are there between how they are interpreted and the data used to achieve them in each of these environments? 7 Why is it important to ensure that all stakeholders participate in the planning process? How can planners ensure that stakeholders participate in the conduct of the systematic conservation plan?

S U GGES T ED R EADI N G Margules, C.R. & Pressey, R.L. (2000) Systematic conservation planning. Nature, 405, 243–253. Margules, C. & Sarkar, S. (2007) Systematic conservation planning. Cambridge University Press, Cambridge, UK. Moilanen, A., Wilson, K.A. & Possingham, H. (eds) (2009) Spatial conservation prioritization: quantitative methods and computational tools. Oxford University Press, Oxford. Possingham, H.P., Wilson, K., Andelman, S. & Vynne, C. (2006) Protected areas: goals, limitations, and design. Principles of conservation biology (ed. by M.J. Groom & G.K. Meffe & C.R. Carroll), pp 509–533. Sinauer Associates Inc, Sunderland. Williams, P.H. & Araújo, M.B. (2002) Apples, oranges and probabilities: integrating multiple factors into biodiversity conservation with consistency. Environmental Modelling and Assessment, 7, 139–151. Wilson, K.A., Carwardine, J. & Possingham, H.P. (2009) Setting conservation priorities. Annals of the New York Academy of Sciences, 1162, 237–264.

PART 3 CONSERVATION PLANNING IN A CHANGING WORLD

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

CHAPTER 7 Planning for Persistence in a Changing World Richard A. Fuller1,2, Richard J. Ladle3,4, Robert J. Whittaker3 and Hugh P. Possingham1 1

School of Biological Sciences, The University of Queensland, Brisbane, Australia CSIRO Climate Adaptation Flagship and CSIRO Sustainable Ecosystems, Brisbane, Australia 3 School of Geography and the Environment, University of Oxford, Oxford, UK 4 Department of Agricultural Engineering, Federal University of Viçosa, Brazil 2

7. 1 I N T R OD UC T I ON There are millions of different species on Earth, each responding uniquely to the environment and to other species, and each with a distinct geographical distribution. This results in enormously complex spatial patterns in nature, from latitudinal gradients in global species richness to the patchy distribution of plants across a meadow. Documenting and understanding these patterns has occupied biogeographers and macroecologists for decades (Chapter 4). There has been considerable progress toward a general understanding (Gaston & Blackburn, 2000; Lomolino et al., 2004), but how should we go about trying to conserve biodiversity in the face of such complex spatial patterns? So far, the conservation community has focused overwhelmingly on elements of the pattern of biodiversity, features that can be mapped spatially and thus ‘captured’ by conservation management activity (Chapters 4–6). For example, a conservation plan might attempt to represent a certain proportion of each vegetation type in a region, or to ensure that places where threatened species occur are designated as protected areas.

This approach would be sufficient if simply capturing elements of the natural world ensured their long term persistence. Unfortunately, there are at least three reasons why this is not the case: 1 Processes generate and maintain biodiversity, so these processes must themselves be conserved. 2 Conservation efforts must track continuously changing patterns of biodiversity. 3 Threats are dynamic, so mitigation efforts must be similarly dynamic for conservation to be efficient. We now expand on each of these points in turn. First, biodiversity is generated and maintained by processes. It is these processes that need conservation, along with the patterns that emerge from them, to ensure that biodiversity persists in the long term. Ecosystems are shaped by myriad physical and biological processes, including climate regimes, oceanic currents, hydrological flows, plate tectonics, demography, migration, dispersal, extinctions, colonizations, predation, competition and distributional shifts, to name but a few. Given the apparent primacy of climate in determining the distributions of species, there has been an intense effort to predict the impacts of climate change on biodiversity and how we might go about responding to the challenges this poses (Section 4.4; Peters &

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Darling, 1985; Opdam et al., 1999; Parmesan & Yohe, 2003; Felton et al., 2009). Despite the importance of dynamic processes such as climate in continually shaping biological systems, a recent review of the conservation planning literature showed that about 80 per cent of studies assume that neither biodiversity nor the processes threatening the persistence of biodiversity change over time (Pressey et al., 2007). This is despite the fact that a decade has now elapsed since the first demonstrations of dynamic approaches to planning protected area networks to promote the long-term persistence of biodiversity (Cowling et al., 1999). Second, even if we successfully conserve processes that generate and maintain biodiversity, the resulting pattern that we see today is only one brief snapshot of a continuously changing system. As if huge spatial complexity isn’t challenge enough for conservation, ecosystems are continuously changing over time (Chapter 3), species’ distributions and abundances expand and contract (Figure 7.1), land bridges between continents come and go with changes in sea level and communities are thrown together and then separate. Deep history leaves a signature in the structure of present-day communities (Box 7.1). For example, modern marine bivalve assemblages show a clear break in species’ ages corresponding with recovery from the end-Cretaceous mass extinction event (Krug et al., 2009). Pleistocene climatic change has promoted massive movements of species populations, often forming novel assemblages, and alternatively isolating and rejoining populations, generating speciation in some lineages and strong phylogenetic structure in many others (e.g. Riddle & Hafner, 1999, 2006; Hewitt, 2000; Bush & de Oliviera, 2006; Avise, 2009). To provide a specific example, Neotoma woodrats appear to have tracked oscillating climates through time by changes in their body size (Smith et al., 1995). The biology of many species thus plays out across time as much as across space, and conservation efforts must therefore track an ever-moving target. Third, threats to biodiversity change in type, distribution and severity over time (for an analysis focused on changing threats on islands, see Whittaker & Fernández-Palacios, 2007). If conservation is about buffering samples of biodiversity from threats, then both the nature of the buffers employed, and where they are instigated, must depend on the type, location and intensity of threats and how these change through time. The cost (e.g. social, economic, political) of

particular conservation action relative to its likely benefit is also highly dynamic, so conservation plans must be continually updated to reflect changing social and economic circumstances.

7. 2 U S I N G T H E PAS T T O U N DER S T AN D T H E PR ES EN T AN D PR EDI CT T H E FU T U R E To ensure long-term persistence of biodiversity, conservation is a dynamic problem in which we must plan for future changes in biodiversity pattern and process, as well as our changing ability to instigate conservation action. Much of conservation is about trying to build scenarios about the future and acting accordingly. The past is the most obvious place to look for guidance about what will happen in the future (Chapter 3). Consider a few simple questions and the critical importance to conservation of long-term ecology quickly becomes clear: • How quickly have species’ distributions responded to past changes in environmental conditions such as fire regimes, habitat availability and climate? • How quickly have species evolved in response to changing environmental conditions? • Where are the places with high rates of extinction and speciation? • Which places across the planet have acted as refugia during past periods of environmental change, and might these prove good long-term investments for conservation? • Do changes in species’ distributions occur predictably in relation to environmental conditions? • How does variability in a species’ current abundance or distribution compare with past fluctuations? • As species head towards extinction, do they show stereotypical patterns of geographical range collapse? Clearly, a long-term perspective is essential if we are to make progress in answering these kinds of questions and using them to guide present and future conservation activity. Human transformation of the planet began long before the science of ecology. While dramatic environmental changes such as Amazonian deforestation are accelerating, and can be observed more or less in real time (Shimabukuro et al., 2006; Hansen et al., 2008), much of the damage and change that humans have wrought on ecosystems occurred hundreds to thousands of years ago, and recent human impacts are

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Figure 7.1 Plots of range size as a function of the age of species within six different groups of species: (a) Old World Acrocephalus and Hippolais reed warblers; (b) New World Dendroica wood warblers; (c) New World Icterus orioles; (d) storks; (e) gannets and boobies; (f) albatrosses. Crosses represent species excluded from the regressions used to estimate the lines of best fit. The inference the authors drew from their study is that species ranges typically show a phase of relatively rapid range expansion post-speciation, followed by a longer, slower phase of range contraction, a scenario broadly consistent with taxon cycle models. Reproduced from Webb & Gaston (2000).

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Box 7.1 Integrating evolutionary considerations into conservation planning Species have traditionally been used as the primary taxonomic level for conservation and, as we have seen, many conservation prioritization analyses essentially rely on weighing up tallies of rare, threatened, or range-restricted species. This strategy has the advantage of needing little or no explanation for policymakers or the public. However, even within a single taxonomic group such as plants or mammals, such a focus may result in some highly valued species or assemblages being omitted from conservation prioritization because they happen not to fall in the most diverse areas. One aspect of the way in which we may value species is in terms of their evolutionary distinctiveness, and here the question arises as to whether a strategy focused on maximizing species richness of a taxon will also succeed in capturing a broad representation of the evolutionary tree, thus preserving the phylogenetic diversity and future evolutionary potential of that group. Phylogenetic diversity (PD) can be defined as a biodiversity index that measures the length of evolutionary pathways that connect a given set of taxa, and it is regarded as a surrogate for the ‘feature diversity’ that arises along the branches of the evolutionary tree (Faith, 2006; Forest et al., 2007). Quantifying the diversity of such ‘features’ is important because it is difficult to know which features of an organism will be advantageous as the environment changes. Thus, maximizing PD may be the best bet-hedging strategy to ensure that evolution has the necessary raw materials on which to work in a rapidly changing world. The potential application of phylogenetic diversity within conservation planning is illustrated by a recent study by Forest et al. (2007) on the remarkable flora of the Cape region of South Africa – an undisputed biodiversity hotspot containing more than 9,000 plant species, a staggering 70 per cent of which are endemic. They collected and compiled distribution data for the entire Cape and created an inventory of species and genera per quarter-degree square. They also reconstructed the phylogeny (phylogenetic tree) of the Cape flora based on analysis of plastids from 735 genera, each indigenous to the Cape. The results were fascinating, if somewhat complex. PD and species (or genus) richness were broadly correlated (areas of high PD also had high richness), suggesting that traditional conservation planning strategies of maximizing species richness may be equally effective at maintaining PD. However, as is often the case, the devil is in the detail. PD was found to scale with richness in a complex manner, so that some regions had more or less PD than would be expected from their species richness. Forest et al. (2007) uncovered a distinctive east/west division in the distribution of PD that broadly corresponded to climatic zones in the Cape region. Specifically, PD for a given number of taxa was higher in the eastern region than in the west (Figure B7.1a). The consequences of such a geographical decoupling of PD and taxon richness is that, in the event of extinctions of non-prioritized species, a traditional taxon (species or genus) richness approach to conservation planning might lose disproportionate numbers of species possessing unique evolutionary characteristics. This, in turn, would reduce the evolutionary potential to respond, adapt and diversify in the light of changes in climate, land use, species composition, etc. Whether such PD approaches will become widespread in conservation planning is debatable. As the authors themselves note, taxon diversity will remain an important conservation target, and it is by no means straightforward to balance prioritizations based on these two indices. Moreover, there is a serious scale issue that may be largely intractable: any conservation plan that operates at less than a global scale runs the risk of finding solutions that are optimal only within the study region. Using distinct phylogeographical regions such as the Cape region reduces, but does not totally resolve, this issue.

Conservation planning in a changing world

a)

b)

c)

d)

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Figure B7.1a Genus richness and phylogenetic diversity in the Cape flora: (a) genus richness (ten quantile intervals from yellow to deep red); (b) phylogenetic diversity (PD) per cell calculated using absolute age estimates in million years (colour code as for a); (c) residuals from a loess regression of PD on genus richness. Cells with negative residuals are indicated in blue, and those with positive residuals are shown in red (shading increments of half a standard deviation); (d) the distribution of unusual PD values, as assessed by comparing the observed PD in each cell with 10,000 PD values calculated by random selection of the same number of genera from the Cape flora. Cells with significantly lower PD (P 50 per cent of species exhibited z-values for the SARs of native species higher than those deemed typical of archipelagic SARs. Moreover, for the three cases in which the percentage of SIE equals or exceeds 90 per cent, the mean z-value is unity (1.00 ± 0.21). The findings thus approximate the schematic shown in Figure B8.1b. Hence, the results of the present work provide significant support for Rosenzweig’s proposition that z-values from inter-provincial ISARs should be very high, approaching unity. This should hold not only for the scale of recognized global biogeographical regions, but also for any system in which speciation is the major process. Of course, the timescales in which species evolve and go extinct may differ between island systems and global biogeographical provinces. Nevertheless, it seems that the two system types exhibit analogous patterns of species accumulation with area. Therefore, studies of evolutionary dynamics in relation to area, employing data for single island endemics, would be well worth pursuing in other taxa and regions, as they could be used as model systems to test: 1 variation in critical island sizes below which within-island diversification does not occur; 2 how these thresholds vary with taxa and island groups; 3 how consistent the form of inter-provincial ISARs are; 4 their capability for predicting future diversity; 5 and they may also be used to recognize and examine the influence of additional factors such as climate, productivity, etc. Such studies can offer great insights into basic questions of conservation biogeography, such as the potential impact of habitat fragmentation and loss and homogenization on biological diversity.

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Applied island biogeography

These macroecological approaches to island data, generated by the stimulus of the MacArthur–Wilson theory, have promoted the wide use of species–area curves for conservation purposes. These include predicting species endangerment globally, regionally and locally (McDonald & Brown, 1992; Tilman et al., 1994; Pimm & Askins, 1995; Brooks & Balmford, 1996) as functions of habitat loss and fragmentation; devising general reserve-design principles (Diamond, 1975a; Wilson & Willis, 1975); and identifying conservation targets for specific habitat types (Desmet & Cowling, 2004). Among the most controversial uses of the species–area relationship based implicitly on ETIB is its application in the forecasting of future species extinctions as a function of habitat loss due to factors such as deforestation (e.g. Brooks et al., 1997, 2002) or future climate change (e.g. Thomas et al., 2004 – see Box 7.3). Projected extinctions based on species–area models involve several uncertainties (Heywood et al., 1994; Whittaker & Fernández-Palacios, 2007) and can never completely replace species-level assessments for the identification of extinction threat (e.g. Kotiaho et al., 2005). However, for many species of conservation concern, the collection of appropriately detailed information is an unrealistic target. It is vital, therefore, that conservation biogeographers develop more realistic indirect measures and theoretical projections of extinctions, based on as pragmatic a set of assumptions as possible (May et al., 1995; Laurance, 2007). The wide variations in outcomes can be seen from efforts to estimate likely extinctions arising from tropical deforestation. Results of current and future rates of deforestation have varied dramatically, ranging from the alarming (e.g. Ehrlich & Wilson, 1991) to more modest (but still significant) losses (Wright & MullerLandau, 2006), thus strongly affecting projections of future species losses. Recently, Wright & Muller-Landau (2006) noted that the estimates of net tropical deforestation rates during the 1990s differ by 250 per cent (see their Table 2). Using a number of criteria, they considered 45 humid tropical countries that support 89.6 per cent of all extant closed tropical forest and 89.9 per cent of all potential tropical forest cover. They concluded that deforestation rates will decrease as population growth slows, and that a much larger area will continue to be forested than previous studies suggest. Such uncertainties, along with differences arising from choice of assumptions about species persistence

in degraded habitats, from the high sensitivity of predictions to uncertainty or errors in species–area slopes and from large uncertainties about both the global species totals and the geographical distribution of biodiversity, mean that all currently available predictions of future losses inherently posses great uncertainty (see Table 8.2, and Chapter 7 and see Laurance, 2007, 2008; Willis & Bhagwat, 2009, for general discussion). Although the most recent of the estimations presented in Table 8.2 was made in 1992, we consider the information to be useful in pointing out the problems in predicting global extinctions that can arise through different assumptions on a number of critical issues. In short, extinction rate estimates based on species–area projections involve many uncertainties (Heywood et al., 1994). The precise form of the relationship describing the loss of species from an original habitat as a function of the remaining habitat area is still an open question. There are two main associated issues. First, many species are not restricted to their ‘native’ habitat and can persist in certain anthropogenic habitats. Second, the slope of the species–area relationship used for the loss of total area of a habitat is still uncertain; there is no strong theoretical or empirical justification for the use of a ‘global’ slope value of z = 0.25 (or any other single value). Whittaker & Fernández-Palacios (2007) have criticized the use of SAR as a means of forecasting species threatened by, or committed to, extinction, noting ‘the way in which the species–area models are used … is conceptually decoupled from the island theory from which it seemingly derives’. They argue first that, a z of 0.25 is a subjective ‘middle’ value to take (see discussion above about the z-values of the different SAR categories). Second, and more crucially, this z-value has been derived from analyses of true isolates. It describes approximately how many species are held in each of a series of isolates/islands of different size. Yet, in several recent studies the z-value is applied not to separate fragments but to an entire region (e.g. Brooks & Balmford, 1996; and see also Box 10.2, pp. 272–273 in Whittaker & Fernández-Palacios, 2007). As will be discussed in the section on nestedness below, depending on the degree of shared species between different habitat islands, it is possible for relatively low or very high proportions of the original species found in a region to be represented in a

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Table 8.2 Some estimates of global species loss due to tropical deforestation and the key assumptions made (adapted from Krishnamurthy, 2003; see therein for references).

Extinction estimate

Total species (millions)/ per cent tropical

Tropical forest loss 2

Extinction/ area lost

Source

1 species/hour by 2000

5–10/40–70%

245,000 km /year

50% species extinct when 10% area left

Myers, 1979

33–50% of all species between 1970 and 2000

3–10/25%

50% deforestation by 2000

Species–area, concave curve

Lovejoy, 1980

1 million species by 2000

4/40%

33% of remaining forest destroyed

50% species in area will go extinct

Myers, 1985

10% of all species by 2000; 25% by 2015

4–5/ 50%

2% deforestation/ year

50% species in area will go extinct

Raven, 1988

17,500 species/year

10/50%

0.7% deforestation/ year

50% species in area will go extinct

Wilson, 1988b

8.8% of all species by 2000

3–10/25%

12.3% deforestation between 1980 and 2000

Species–area, concave curve

Lugo, 1988a, b

5–38% of all species between 1990 and 2000

10/>50%

0.8–1.6 % deforestation/year

Species–area; z = 0.15, 0.35

Reid & Miller, 1989

27,000 species/year

10 in tropical rain forests

1.8% deforestation/ year

Species–area; z = 0.15

Wilson, 1992

series of habitat islands. Therefore, treating what are actually archipelagos of habitat islands as though they were a single island in analyses of extinction threat is a potentially crucial oversimplification – and it is one reason why we cannot rely upon the ‘90 per cent area loss = 50 per cent species loss’ generalization with which we began this section. As a further note of caution, it is important to emphasize that while the species–area relationship is indeed a very general pattern, area rarely explains all interpretable variation in species richness, with some residual variation being attributable not only to system isolation but to other variables such as habitat diversity, elevational range, disturbance regime, etc. (Whittaker & Fernández-Palacios, 2007; Triantis et al., 2008). It follows that SARs can only provide a crude approximation for use in conservation planning. Hence, as noted by Whittaker et al. (2005), the application of the species–area relationship for informing conservation sciences is one area within conservation biogeography where the theory appears to require further work.

8.2.2 Relaxation and the extinction debt Newly emerged islands present new habitat and accumulate species through time via immigration. In contrast, habitat islands created through isolation by rising water levels or by habitat destruction (e.g. deforestation) are typically assumed to support something approximating a full complement of local species at their formation. That is, they are expected to contain both source populations (having positive population growth within the area itself) and sink or casual populations that happened to be present at the time of isolation, but which do not exhibit positive population growth within the area itself. Upon isolation these islands are thus ‘supersaturated’ for a patch of their newly reduced area and increased isolation. With time, these islands lose species, a phenomenon called species relaxation (Diamond, 1972; Wilcox, 1980). Immigration (at a lower rate than before) and extinction (at a higher rate than before isolation) should both continue during the relaxation period and subsequently they come back into balance; at this

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point, the island has found its new, lower equilibrium richness level (Figure 8.1). The time taken for relaxation to occur is referred to as the ‘lag time’ and the anticipated eventual species loss is termed the ‘extinction debt’ (see Ewers & Didham, 2006). Two classic examples of relaxation are presented in Figure 8.5. The first example takes the form of a data set for the mammals living in isolated high mountains forests of the south-western USA (Brown, 1971). The radical shift in climate following the end of the latest glacial period resulted in these mammalian populations becoming isolated from each other by the increasing aridity of the valleys surrounding them. The second example is for the mammals of the Sunda Islands of Indonesia (Wilcox, 1980). These islands, interconnected during the last glacial period, were isolated by the ocean as the glaciers melted and raised the sea level. Thus, both these systems became isolated around the start of the Holocene (c. 10 ka) and since then are thought to have gradually been losing species. Biogeographers and conservationists have been interested in three general questions related to species relaxation. First, how does relaxation proceed? In other words, what is the shape of the curve of species loss over time? Second, how much time is needed between fragmentation and extinction (the lag time)? Third, and critically, how many species will be left after relaxation is complete? Conversely, how many and which species will eventually go extinct? Relaxation after habitat loss and fragmentation is typically expected to proceed in a sequence of stages (after Wilcove, 1987): • Stage 1. Initial exclusion. Some species will be lost from the landscape simply because their original ranges did not include any of the remnant patches. • Stage 2. Extirpation due to lack of essential resources. Species vary greatly in their resource requirements and many require very large areas and/ or very rare resources. Thus, the likelihood that all of a species’ resource requirements can be met decreases as the remaining area decreases. • Stage 3. Perils associated with small populations. Small populations are much more susceptible to a host of genetic, demographical, and stochastic problems. As the total area of the remnant patches decreases, and the ability to sustain large populations decreases, these problems become increasingly severe (e.g. Frankham et al., 2002).

0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1

1

2

3

4

5

Figure 8.5 Mammal diversity on Sunda Islands (circles) and south-western US mountaintops (diamonds). These isolates started forming about 10,000 years ago at the end of the Pleistocene. In both archipelagos, larger islands have experienced proportionately fewer extinctions. The above estimations are based on two general assumptions. First, it is assumed that the extinction rates are comparable. Rosenzweig (1995) considered that, as the same taxon is studied and given that the two systems have been formed due to the same event (the switch into the current interglacial) and thus began losing species at approximately the same time, we can hypothesize a similar rate of extinctions. Second, it is assumed that the original number of species for each island can be estimated from a mainland area (Malaysian mainland for Sunda Islands and Sierra Nevada for US mountaintops) of the same size as the island. Re-drawn after Rosenzweig (1995, his Fig. 6.5).

• Stage 4. Deleterious effects of isolation. Some populations may be rescued from extinction by migration and recruitment of individuals from other populations. The likelihood of such rescue effects decreases as isolation increases. • Stage 5. Ecological imbalance. Most species are strongly influenced by interactions with other species. Loss of one species during any of the aforementioned stages of relaxation may result in the subsequent loss of its predators, parasites, mutualists, or commensals (e.g. Koh et al., 2004). In addition, habitat disturbance and reductions in community diversity during the earlier stages of relaxation may facilitate the establishment of introduced species, triggering a cascade of subsequent extirpations.

Conservation planning in a changing world It has been argued that it may take several generations for the processes causing relaxation to play out following habitat destruction and fragmentation, meaning that there is a substantial lag time between the initial stimulus and the end of the process of species losses (Tilman et al., 1994; Ewers & Didham, 2006; Vellend et al., 2006). This creates an ‘extinction debt’ – a future ecological cost of habitat destruction that may not be initially apparent in studies made shortly after habitat fragmentation has occurred. Indeed, Brown’s (1971) mammal assemblages were hypothesized to still be in the process of relaxation from their relatively large mountain top habitat islands thousands of years after isolation (Figure 8.5; and see further discussion in Lomolino et al., 2006). Whether such protracted response times are typical is unknown, but it does seem highly likely that the true ecological costs of the historically recent spate of anthropogenic habitat disturbance, destruction and fragmentation across the globe are yet to be realized (see, for example, Figure 8.6). It is also noteworthy that, although the majority of recorded species extinctions since AD 1600 have occurred on oceanic islands, predictions of increasing numbers of future extinctions suggest a significant shift to continental areas (Millennium Ecosystem Assessment, 2005). Developing methods to quantify the magnitude and taxonomic distribution of the extinction debt is clearly vitally important for effective conservation planning and prioritization. However, this objective is by no means simple to attain. Accurate assessment of extinction rates and their extrapolation into the future requires good quality long-term data on species occurrences – data which are generally lacking, especially for less conspicuous and/or numerically much more species rich taxa. This lack of appropriate knowledge (Chapter 4) has led to an inevitable reliance on indirect measures and theoretical projections of extinction debt. These include: species–area models; rates at which wellknown species are shifting to increasingly more threatened categories of conservation concern; extinction probabilities associated with the IUCN categories of threat; impacts of projected habitat loss on species currently threatened with habitat loss; and the extrapolation of correlations of species loss with climate change (e.g. McDonald & Brown, 1992; Mace & Kunin, 1994; Pimm & Askins, 1995; Thomas et al., 2004 – for further discussion see Ladle, 2009).

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One of the best-known empirical examples of relaxation on an ecological timescale is the loss of bird species from Barro Colorado Island in Panama. The island was formerly a hilltop in an area of continuous neotropical lowland rain forest, but abruptly became a 15.7 km2 island when the central section of the Panama Canal Zone was flooded to make Lake Gatun in 1914. Of about 208 bird species estimated to have been breeding on Barro Colorado island immediately following isolation in the 1920s and 1930s, 45 were no longer present by 1970 (Wilson & Willis, 1975). However, only a minority of these losses were directly attributable to stochastic processes of relaxation. The others could be attributed to ecological changes such as forest regeneration following abandonment of farming activity, which reduced the availability of open habitats, or predation by terrestrial mammals (see review in Whittaker & Fernández-Palacios, 2007). For example, many of the birds lost were typical of second growth or forest edge, suggesting that the regeneration of the forest following abandonment of farming activity must have reduced the availability of these more open habitats. Additionally, some groundnesting species were probably eliminated by their terrestrial mammalian predators, which became abundant after the disappearance of top carnivores with large area requirements. This effect, of increasing numbers of smaller omnivores and predators due to the absence of large ones, has been termed mesopredator release (Soulé et al., 1988) and has been documented to occur in several other similar contexts (e.g. Laurance, 2002). A later avifaunal survey of Barro Colorado Island reported sightings of 218 species from the island or the waters immediately around it between 1994 and 1996, including five new records, none of which were thought to be of breeding species (Robinson, 1999). As anticipated from the island theory (Figure 8.1), the rate of species loss appears to have declined over time, especially for forest-interior birds. However, overall, species extinctions do appear to have continued to exceed colonizations. So, in summary, the isolation of the hilltops to form this lake-bound island has been followed by around a century in which the process of relaxation has been the dominant trend. Future changes in avifaunal species richness and composition on the island are likely to be dependent on the extent to which the

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Extinctions per thousand species per millenium 100,000

Distant past (fossil record)

Recent past (known extinctions)

Future (modelled)

10,000

Projected future extinction rate is more than ten times higher than current rate

1,000

Current extinction rate is up to one thousand times higher than the fossil record

100

10

For every thousand mammal species, less than one went extinct every millennium

1 Long-term average extinction rate

0.1

Marine Mammals species

Mammals Birds Amphibians

All species

Figure 8.6 Past and future extinctions. ‘Distant past’ refers to average extinction rates as calculated from the fossil record. ‘Recent past’ refers to extinction rates calculated from known extinctions of species (lower estimate) during the last 100 years or known extinctions plus ‘possibly extinct’ species (upper bound). ‘Future’ extinctions are model-derived estimates using a variety of techniques and, in general, refer either to future loss of species based on the level of threat that exists today or to current and future loss of species as a result of habitat changes. The techniques involved for modelling future extinctions are: species–area models; rates at which species are shifting to increasingly more threatened categories; extinction probabilities associated with the IUCN categories of threat; impacts of projected habitat loss on species currently threatened with habitat loss; and correlation of species loss with energy consumption. According to the authors of the assessment, the lower bound estimates for future modelled extinctions are low certainty estimates, and the upper bound estimates are speculative (i.e. even lower certainty). Adapted from Millennium Ecosystem Assessment (2005).

nearby mainland forest cover is retained, as these forests provide the source of the transient birds and occasional new colonists (those that stay to breed) observed on Barro Colorado Island. The most significant problem with predicting future extinctions in such systems is that we have an inadequate theoretical and empirical basis by which to estimate the rate at which species will be lost over time or the total time period required for a new (dynamic) equilibrium to be achieved. Precise estimates of the ‘time to extinction’ of each species under threat remains an unrealistic aim for both true and habitat islands, as it will largely be species- and system-specific.

A classic illustration of this problem is provided by the tropical moist forests of the Atlantic seaboard of Brazil, known as the Mata Atlantica, which have been reduced over the past few centuries to only about 7 per cent of their estimated former cover (Ribon et al., 2003). This is a large region and the remnants are numerous and widely distributed, so the ‘90 per cent habitat loss = 50 per cent species loss’ rule of thumb (above) should not really be expected to apply. Nonetheless, such habitat loss and insularization should have driven significant losses. To date, however, no extinctions have been documented with any degree of certainty, although many species appear on IUCN ‘Red Lists’ as ‘vulnerable’,

Conservation planning in a changing world ‘endangered’ or ‘critically endangered’, largely based on reductions in range or population estimates (Ribon et al., 2003). Despite the criticism, especially of methodology and taxonomic bias (e.g. Régnier et al., 2009), the IUCN Red List has become an essential source of information for conservation action and is widely recognized as the most comprehensive compilation of extinct and threatened species (Mace & Lande, 1991; Rodrigues et al., 2006). Brooks and Balmford (1996) compared losses of birds in the region, projected using a species–area model, with those listed by the IUCN as ‘threatened’, and they found congruence. They concluded that the forecasts of looming extinction are basically correct, but that there is a substantial lag between the habitat loss/fragmentation process and global extinction of the species. There is, moreover, good evidence of local extirpation within the existing range of many bird species that live in devastated habitats such as the Mata Atlantica. Thus, within the Viçosa region (a 120 km2 area in south-eastern Brazil) over the last 70 years, it appears that at least 28 bird species have become locally extinct, with 43 being classified as ‘critically endangered’ and 25 ‘vulnerable’. In total, 61 per cent of the original avifauna has been significantly reduced in incidence (Ribon et al., 2003). Nectarivorous species appear to have been affected least, followed by omnivores and carnivores, with frugivores and insectivores hit the hardest. Assuming relaxation to be well under way in these fragmented systems, the questions of estimating the time lag between habitat loss and eventual species losses, and of predicting the identities and numbers of these losses, remain unanswered. In this context, as noted by Raheem et al. (2009), it is surprising that fragment age (i.e. the time of isolation/creation of a fragment) has received little attention from ecologists and conservationists. In most landscapes, it is the productive and/or most accessible areas that are deforested first, thus producing a nonrandom spatial and temporal distribution of habitat fragments (Laurance et al., 2002; Ewers et al., 2006). A topographically diverse landscape, such as on most oceanic islands, will therefore typically contain an assortment of older, smaller and more degraded fragments at lower elevations and younger, larger and less degraded fragments at higher elevations (see for example Box 8.2). The above example illustrates the difficulties of untangling causal processes underlying species

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relaxation within real landscapes. Part of this complexity is generated by the operation of two overlapping temporal scales that are critical in determining net rates of species loss across fragments: the rate at which habitat is being lost from a region (considering also the possible recovery of habitat; see Wright & Muller-Landau, 2006); and the age of the habitat fragments created within that region. Untangling the contribution of these two interlinked age-dependent factors may be critical to a better understanding of the relaxation process and thus for more accurate predictions of species losses and relaxation lag time. Recently, Raheem et al. (2009), studying the land snails assemblages in fragments of natural rain forest in Sri Lanka’s wet zone, concluded that fragment age, along with fragment shape complexity, were the only two significant determinants of fragmentation-related changes in community composition. Attributes of fragments such as area, distance-to-edge and matrix quality, which have been traditionally linked to species losses, exhibited no obvious effect (see also below). In practice, review of the literature on such effects reveals many such idiosyncrasies between studies. At least some part of the differences in findings from one case study to the next reflects differences in the ‘experimental design’ of the fragmented systems analysed and, in particular, the range in values of properties such as area, age, distance from source, habitat complexity, etc. that each study encompasses.

8.2.3 Ecosystem collapse and threshold responses in habitat islands Reduction of habitat area can causes super-saturation as immigration rate declines and extinction rate rises (above). In the most extreme scenario, where the loss of habitat is so extreme that immigration into a patch virtually ceases, species richness may, in theory, collapse catastrophically (see Whittaker & FernándezPalacios, 2007, their figure 10.5; and also Vandermeer & Lin, 2008). Although the process of species richness collapse and associated loss of ecosystem function is not presently well-defined or understood, it is thought to be linked to extreme impoverishment of the available resources that are required for a system to sustain its functionality (e.g. Dobson et al., 2006). One of the most emblematic examples of an ecosystem collapsing comes from the island literature. Easter Island (Rapa Nui) was once one of the most isolated

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Box 8.2 Extinction debt in the Azores Box prepared by K.A. Triantis and R.J. Whittaker – excerpted and modified slightly from Triantis et al. (2010). See the source paper for a full analytical presentation of the materials and methods used. Habitat destruction is considered to be the leading cause of terrestrial species extinctions. However, there is typically a time lag between the reduction in habitat area and the eventual disappearance of the remnant populations. These ‘surviving but ultimately doomed’ species represent an extinction debt. Calculating the magnitude of such future extinction events has been hampered by potentially inaccurate assumptions about the slope of species–area relationships, which are habitat- and taxon-specific [see text]. We have overcome this challenge by introducing a novel method that uses the historical sequence of deforestation in the Azorean Islands to calculate realistic and ecologically-adjusted species–area relationships. The Azores constitute an ideal model system for assessing extinction debt because: 1 they have lost more than 90 per cent of their original native forest during the five centuries of human occupation; 2 being one of the most isolated archipelagos on Earth they support a significant number of single island endemics (SIE); 3 the history of human settlement and deforestation is well known; 4 extensive biogeographical data exist for a range of taxa. The Azorean Islands were discovered in AD 1432 by Portuguese explorers, and more than 500 years of human settlement have taken their toll on the local fauna and flora, 420 species of which are endemic to the archipelago. Today, approximately 70 per cent of the vascular plants and 58 per cent of the arthropods found in the Azores are exotic, many of them invasive. The destruction of the native ‘laurisilva’, a humid evergreen broadleaf laurel forest, in the Azores has followed a clear temporal sequence. At the time of human colonization (c. AD 1440), the archipelago was almost entirely covered by forest. By 300 years ago (c. AD 1700) human activities had restricted the native forest in most islands to areas above 300 m a.s.l. and, by c. AD 1850, areas with native forest were present only above 500 m a.s.l. The development of an economy dependent on milk production during the last decades of the 20th century drove a further reduction of native forest area, to 2.5 per cent of the total area of the archipelago (Figure B8.2a). The Azorean arthropod fauna has been intensively sampled during the last ten years. The Borges et al. (2005) checklist includes virtually all arthropod species native to the Azores, as well as an accurate description of their presence or absence in all the islands of the archipelago. The endemic arthropods belonging to three groups – Araneae, Hemiptera and Coleoptera – were classified as native forest dependent and non-forest dependent species, and only the forest dependent species endemic to the archipelago were considered for further analyses. We used four different ‘habitat areas’ to calculate our species–area relationships: these were chosen to correspond to the extent of native forest at four known points in time before and since human colonization (≈AD 1440, AD 1700, AD 1850, AD 2000; Figure B8.2a). Although the historical estimates of forest cover are crude approximations, we consider that they are accurate enough to provide a baseline for estimating the present extinction debt. Our analyses follow the rationale that if species ‘relaxation’ has not yet taken place or is incomplete (i.e. the extinction debt has not yet been paid), then the best fitting species–area model will correspond not to present forest area but to a past baseline – a hypothetical dynamic equilibrium from which the system has since departed. However, there is a complication in dealing with a system of endemic species on oceanic islands of varying age, namely that the dynamics of colonization, speciation and extinction may be at different points, depending on the age of the island. Accordingly, we fitted and compared both species–area and species–area–time models.

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a) T4 T3 T2

b) 1440 AD 1700 AD 1850 Present

AD

Log Species

T1

Time T1

T2

T3

T4

Log Area

Figure B8.2a The sequential reduction of the native forest and the respective species–area relationships. (a) The elevational occupancy of native forest in historical times for the island of Terceira (Azores). T1: Before human occupation (almost complete coverage of island’s area); T2: approximately 300 years ago (300–500 m); T3: approximately 160 years ago (above 500 m); T4: current distribution. (b) A schematic representation of the effects of the sequential reduction of the native forest on the species–area relationships of endemic forest arthropods. The dashed lines in T4 represents the future species–area relationships, extrapolated from T1 and T2 (see text). The magnitude of the extinction debt is represented by the difference between current species richness (solid line) and the future predictions (dashed lines). From Triantis et al. (2010).

For the total island area and the habitat area above 300 m, the species–area–time model applied was significant (P < 0.05) for the arthropod taxa considered, with most of the explained variance attributable to area. However, for the area above 500 m and the present area covered by native forest, neither the species–area–time relationships nor the respective species–area relationships were statistically significant. We thus used the first two benchmark relationships, for total area (≈ AD 1440) and area above 300 m (≈ AD 1700), to represent the baseline conditions for estimation of current extinction debt. Hence, we used the parameters estimated for the total area of the islands (Prediction 1 in Table B8.2a) and that of the area above 300 m (Prediction 2 in Table B8.2a) to estimate the number of endemic forest arthropods that ‘should’ be present and, by direct comparison with the number of extant species, to derive the number of future extinctions (i.e. the extinction debt) (Table B8.2a). For the arthropod taxa considered, our results clearly indicate that the majority of the endemic forest-inhabiting species (>50 per cent) are expected to go extinct in time, especially on those islands on which the native forest has been restricted to small areas or has been totally removed. Terceira, the island with the largest remnants of native forest, has the smallest number of predicted future extinctions. At face value, these figures constitute a powerful warning to island conservationists that the worst of the extinction crisis is by no means over. Furthermore, in spite of the fact that some archipelagic endemic species may benefit from a degree of population reinforcement between habitat fragments or islands, the parallel reduction of the native forest across all islands in the last 600 years has greatly diminished the probability of such source-sink dynamics rescuing species from global extinction. Hence, we would also anticipate a correspondingly large number of archipelagic-scale species extinctions for Azorean endemic arthropods in the future as the extinction debt is settled. In point of fact, at least five SIE species of beetles recorded early in the 20th century have not been recorded since 1965 and might therefore be considered extinct.

In the paper, we argued that the figures reported above are likely to be more accurate than previous predictions because we have focused our attention on endemic forest species that have evolved in,

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Table B8.2a Number of forest-dependent archipelagic endemic arthropods for the nine Azorean Islands and the respective predicted number of species that should be found based on the species–area–time models from the total area (Prediction 1) and the area above 300 metres (i.e. area occupied by native forest c. 300 years ago; Prediction 2). Currently there is no native forest on Graciosa and Corvo islands. Note that the results remain similar when the different groups, i.e. Coleoptera, Araneae and Hemiptera, which have been lumped together in the table, are analysed separately.

Island

Arthropods

Prediction 1 (total area)

Prediction 2 (area >300 m)

Species loss

Graciosa Corvo Flores Faial Pico São Jorge Terceira São Miguel Santa Maria

8 3 24 17 28 21 29 34 24

1.14 1.00 4.01 1.11 2.55 1.28 5.79 1.84 0.29

2.42 1.34 8.04 2.59 4.66 2.83 12.10 5.28 1.93

86–70% 66–55% 83–67% 93–85% 91–83% 94–87% 80–58% 95–84% 99–92%

and are only found in association with, the native forest. At the same time, we avoided additional ‘noise’ caused by generalist species that may well be able to survive in other (i.e. anthropogenic) habitats. If this logic is correct, then the implication is that large-scale conservation efforts need to be implemented if the high extinction debt we have identified is to be deferred or avoided. Humaninduced fragmentation, land-use changes and invasive species have already been identified as important threats to Azorean biodiversity. This paper argues that the conservation of the Azorean natural heritage, and that of many other oceanic islands, will largely depend on establishing an integrated large-scale strategy to manage both indigenous and non-indigenous species, while simultaneously protecting the remnants of native habitat and, ideally, increasing their extent. However, as appreciated by the authors (and pointed out by the journal’s reviewers), there are a number of key assumptions embedded in this study that may undermine the power of the analyses and which may serve as points for class discussion. These include: 1 the reliance on an assumption of a dynamic equilibrium, or something approximating to it, prior to human interference; 2 that the endemic species identified as forest-dependent can persist only in the native forests; and 3 that the remaining forest area, although in most cases fragmented, can be treated as if occurring in a single block, with the degree of fragmentation being insignificant.

fragments of inhabited land in the world. When the first Europeans arrived there on Easter Day AD 1722, they found a fascinating enigma: how could this impoverished, nearly treeless island, with its sparse and impoverished population, have supported the construction of the remarkable giant statues (moai) that could be found all over the island. How, and why, had it all gone so terribly wrong?

The flora of Easter Island currently consists of over 200 vascular plant species, of which only 46 are native. However, the native flora was once rather richer, containing several native tree and shrub species (Diamond, 2007). The forests are now known to have contained a giant palm tree and a number of other trees, some reaching over 30 m in height. These forests persisted for at least 33,000 years (as far back as the

Conservation planning in a changing world palaeoecological record goes) and survived the major climatic shifts of the late Pleistocene and early Holocene. We can thus be certain that deforestation caused directly or indirectly by humans was responsible for the treeless state of the island observed by the first Europeans (Diamond, 2007). Recent studies have recorded that more than 20 tree and woody plant species were exterminated as an eventual outcome of Polynesian settlement. The palm was almost completely gone by AD 1450 and the other large trees by AD 1650. What is not known with any certainty is exactly how long this almost suicidal environmental destruction took. The date of the first settlers arriving on the island is still debated. Estimates range from AD 300 to 1200, with the most recent date considered as the most reliable earliest date of occupation (Hunt, 2006; Hunt & Lipo, 2006). Their effect on the forest was soon detectable in the pollen record and it had been entirely eliminated for some time by the end of the 17th century (Figure 8.7; Hunt & Lipo, 2006; Diamond, 2007). The human population reached its peak around AD 1600, but subsequently was intensely reduced, along with the megalithic culture that had sustained the quarry-

AD

900–1200

Small colonization groups arrive; population grows and soon begins building ahu and moai

AD

1150–1250

Population continues to grow causing widespread forest clearance; additive effect of rats on regeneration of palm tree forests

AD

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ing, sculpting, transport and erection of the remarkable giant statues. The loss of trees and other plant species is matched by a more complete loss of native birds than on any comparable island in Oceania (Steadman, 1997, 2006). Bird bones associated with Polynesian artefacts 600–900 years old showed that Easter Island once sustained at least 22 species of seabird, of which only seven now occur on one or two offshore islets, and just one of which still nests on Easter Island. Bones also provide evidence for the existence of six endemic land bird species, a heron, two rails, two parrots, and an owl, none of which survive. Embodied in ecosystem collapse is the concept of trophic cascade, i.e. the chain of knock-on extinctions following the loss of one or a few species that play a critical role (e.g. as a pollinator) in ecosystem functioning. A perturbation at one trophic level propagates through lower levels with alternating positive and negative effects, as highlighted in the phenomenon of mesopredator release, outlined earlier. Thus, the removal or absence of large predators would be expected to lead to increased densities of consumers, which, in turn, would be predicted to have negative

1300–1400

Maximum human population; rate of deforestation peaks

AD

1680

Island largely deforested; plants and grass used for fuel instead of wood; population reduction

AD

1772

First Europeans arrive and find remnant population of around 3,000; some trees still remain

Figure 8.7 Major events in the sequential collapse of the Easter Island ecosystem. Adapted from Hunt (2006) and modified according to the account provided by Diamond (2004, 2007).

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consequences for producers (Oksanen & Oksanen, 2000). Terborgh et al. (2001) studied a set of large predatorfree islands created by a hydroelectric impoundment in Venezuela. The small area of the islands restricted the predator community to species predating invertebrates (e.g. birds, lizards, anurans and spiders) and seed predators (rodents), alongside herbivores (howler monkeys, iguanas, and leaf-cutter ants). Predators of vertebrates were absent, and densities of rodents, howler monkeys, iguanas and leaf-cutter ants were found to be 10 to 100 times greater than on the nearby mainland, suggesting that larger predators normally limit their populations. Moreover, the densities of seedlings and saplings of canopy trees are severely reduced on herbivore-affected islands. Terborgh et al. (2001) found support for the idea that hyper-abundant folivores could reduce species-rich forests to an odd collection of herbivoreresistant plants. The endpoint of such a process is likely to be a biologically impoverished system. All of the above examples suggest the existence of taxon- and system-dependent thresholds, beyond which species losses accelerate (Ewers & Didham, 2006; Whittaker & Fernández-Palacios, 2007). Such thresholds are highly pertinent to understanding relaxation as a result of habitat loss and fragmentation. The most dramatic changes seem to be those following the loss of a trophic tier, typically the loss of top predators. However, similarly dramatic changes can follow the addition of a tropic tier, as seen when terrestrial vertebrate predators are introduced to remote islands previously lacking them (Terborgh, 2010). Given the central importance of the topics of habitat fragmentation and species relaxation in predicting current and future extinction rates, it is surprising that more attention has not been given to experimental analyses of threshold effects and to studies of the timescales over which the ‘extinction debt’ persists (see Simberloff & Martin, 1991; Brooks et al., 1999; Laurance, 2002). Although restricted to metapopulation model simulations (of which, more follows below), Keymer et al. (2000) have shown that persistence in dynamic landscapes depends on the interaction between three factors: the amount of habitat in the landscape; the rate of change of the amount of habitat; and the life history of the species living in the landscape. More generally, they suggest that including temporal considerations into models of landscape

structure changes the extinction threshold – the amount of habitat destruction a population can tolerate – by making the threshold sensitive to the rates of destruction.

8. 3 S PECI ES I N CI DEN CE 8.3.1 Minimum viable populations, minimum areas and incidence functions In his seminal paper, Caughley (1994) identified two prevailing paradigms in conservation biology: the ‘declining population paradigm’ and the ‘small population paradigm’. The declining population paradigm is the identification and management of the processes that depress the demographical rate of a species and cause its populations to decline deterministically, whereas the small population paradigm is the study of the dynamics of small populations that have declined owing to some (deterministic) perturbation, and which are more susceptible to extinction via chance (stochastic) events. These concepts underpin the formulation of extinction-risk criteria. Theoretical and empirical work has repeatedly shown that, once reduced in size and geographical range, populations face a considerably elevated risk of extinction (MacArthur & Wilson, 1967). Or, as Darwin (1872, p. 133) put it: ‘Rarity … is the precursor to extinction’. There are actually several different forms of rarity (Box 4.1), the most extreme form of which is when a species is reduced to a small population entirely isolated from supplementary immigration, or indeed to the very last such population of the species. In the late 1970s, researchers identified the need to characterize quantitatively the long-term viability of such small and entirely isolated populations (Soulé & Wilcox, 1980). This led to the concept of the minimum viable population (MVP), the smallest number of individuals required to provide a specified probability of persistence over a given period of time (Shaffer, 1981). For instance, the MVP could be operationalized as ‘the population size required to ensure a 99 per cent probability of the species’ population persisting for 40 generations or for 1,000 years’ (see, e.g. Reed et al., 2003). Theoretical estimates of MVPs typically vary from as few as 50 to as many as 10,000 individuals, based on the postulated effects of demographical, genetic and

Conservation planning in a changing world environmental variation (Reed et al., 2003; Brook et al., 2006), with the available empirical evidence pointing to the upper end of this range (e.g. Reed et al., 2003). It has been estimated that the maximum tolerable rate of inbreeding is 1 per cent per generation, which has in turn been translated to approximately 50 individuals to ensure short-term fitness (see Shafer, 1990). However, typically only a proportion of the adult population participates in breeding and it is these animals that form the effective population size, which is often substantially smaller than the total population size (Shafer, 1990; see Crandall et al., 1999, for discussion on the concept). A study of grizzly bears in the Yellowstone National Park showed that to prevent inbreeding rates exceeding 1 per cent required an overall population size of at least 220 rather than 50 animals (Shafer, 1990). Further rule of thumb estimates have been collated by Frankham et al. (2002; their Table 14.1, p. 339) as follows: the population numbers required to avoid inbreeding depression and to retain fitness in the short term, >50; to retain evolutionary potential, 500– 5,000; and to avoid the accumulation of deleterious mutations, 12 to 1,000 individuals. Attempts to calculate the viability of single populations (i.e. whether the population is likely to persist for a given period of time) are referred to as population viability analyses (PVA) (see Reed et al., 2003). PVA can take into account the combined impacts of stochastic factors (demographical, environmental and genetic stochasticity) and deterministic factors (e.g. habitat loss, over-exploitation). According to Brook et al. (2006), PVA and the threat categories of IUCN (Box 4.1) each offer an assessment of a species’ probability of extinction based on its current population size and structure and the characteristics of the threatening processes it faces. On the other hand, the main feature of MVP analysis is that the risk of extinction is fixed and the critical question asked is how large a population must be to avoid this risk. Demographical stochasticity of initially small populations can lead to losses from a series of isolates without a need to invoke any specific mechanism such as predation or loss of fitness. However, where small populations persist for a reasonable length of time (e.g. several generations), they may also lose genetic variability as they pass through bottlenecks. They may then lose fitness by lacking the genetic flexibility to cope with either the normal fluctuations of environment or an altered environment, and they may also

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accumulate so-called deleterious genes, i.e. genes that reduce survival or fertility (see Caughley, 1994). A further complication in assessing genetic effects of fragmentation is that where a species is split into numerous separate populations in fragmented habitats, there may be multiple bottlenecks involved. This may result in reduced variation within each population, but increased genetic differentiation between populations (see Leberg, 1991). The viability of an isolated population may also be influenced by the occurrence of environmental change or disturbance, and indeed it has been argued that it is critical to take such environmental catastrophes and fluxes into account when estimating the MVP and designing conservation measures based around protecting such small, endangered populations (Mangel & Tier, 1994). An example of synergetic effects of a catastrophic event and inbreeding is provided by song sparrows (Melospiza melodia) living on Mandarte Island in western Canada. The inbred birds died at a much higher rate during a severe storm than did outbred birds (Keller et al., 1994). Although the severe weather was what caused this mortality, it appeared that inbreeding determined, in part, which individuals survived the storm. Recently, Reed et al. (2003) considered the effects of age structure, catastrophes, demographical and environmental stochasticity, and inbreeding depression, to derive MVP estimates for 102 vertebrate species. They defined an MVP as ‘one with a 99 per cent probability of persistence for 40 generations’. Across this data set, mean and median estimates of MVP were 7,316 and 5,816 adults, respectively. The estimated values did not differ systematically between major taxa, or with trophic level or latitude, but were negatively correlated with population growth rate. Reed et al. (2003) stress that although MVPs provide a useful rule of thumb for species conservation (which is that the size of vertebrate populations needed for successful long-term conservation is about 7,000 adults), MVPs should not be used as precise conservation targets. For further discussion see also Brook et al.’s (2006) study on the MVP of 1,198 species. Closely related to the concept of MVP is the idea of the minimum viable area (MVA). For some species, e.g. snail populations, a fairly small area may suffice to maintain the requisite number of individuals. Species of higher trophic levels generally require more area or space to ensure good survival prospects. It has been

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calculated that a single pair of ivory-billed woodpeckers (Campephilus principalis), a species generally considered extinct, may have required 6.5–7.6 km2 of appropriate forest habitat; that the European goshawk (Accipiter gentilis) has a home range of about 30– 50 km2 (Wilcove et al., 1986); and that populations of North Island brown kiwi (Apteryx mantelli) in New Zealand are unlikely to be viable in protected areas of less than 100 km2 (Basse & McLennan, 2003). Even some plant and insect species may need surprisingly large areas if they typically occur at very low population densities (Mawdsley et al., 1998). Thus, for many species, reserves must be really rather large if their purpose is to maintain a MVP entirely within their bounds. For instance, it has been estimated that the minimum viable area for some large mammals during a time span of 1,000 years exceeds 100 times the area of Yellowstone National Park (Shafer, 1995). The MVA approach, if focused on large-bodied flagship species and converted into policy recommendations, may have benefits for the preservation of entire ecosystems, since many other species with lesser area requirements may benefit from protection within the MVA of the flagship species. However, one limitation of the MVA approach is that it is predicated on the idea that each area is discrete and has no biotic (genetic) exchange with other surrounding areas. If there is such exchange taking place, such that the population is actually part of a network (or metapopulation), then the estimated MVA may be larger than is strictly necessary (see discussion in Whittaker & Fernández-Palacios, 2007). Another way of examining area requirements of particular species is by means of incidence functions estimating the probability of a species occurring as a function of a key controlling variable, such as island species richness, area, isolation, or sometimes a combination of two key variables (Diamond, 1975b; Wilcove et al., 1986; Watson et al., 2005; and see Figure 8.8). In 1994, Hanski introduced the incidence function model (see also Hanski, 1999; Moilanen & Hanski, 2006). The simplest form uses a snapshot of species presences and absences and predicts extinctions based on patch size and colonizations based on isolation (see MacKenzie et al., 2006, for discussion on the concept). Moilanen (2002) has drawn attention to three main types of errors likely to occur in data used for incidence functions analyses:

1 inaccurate measurements of the patch areas; 2 the existence of patches that are unknown within or around the study area (‘missing patches’); 3 patch occupancy is incorrectly observed, with patches considered to be empty actually containing a population of the focal species (the ‘false zero’ problem). Perhaps more fundamental a problem is that species incidence functions tell us, of course, the properties of ‘islands’ on which a target species currently occurs, but not those on which it may persist in the long term or in an altered ecological conditions. Thus, they are not equivalent to estimating MVAs. Biedermann (2003) provides an interesting analysis of area–incidence relationships of 50 species of vertebrates and invertebrates from 15 different fragmented landscapes, ranging from Central European grassland to Asian tropical forest, in which he demonstrates that area requirements increase essentially linearly with increasing body size on a log–log scale. Biedermann

Figure 8.8 Examples of species incidence functions based on logistic regression models across different landscapes and ecosystems, ranging from Central European grassland to Asian tropical forest: (a) Kelisia haupti (planthopper); (b) Arytaina genistae (psyllid); (c) Neophilaenus albipennis (spittlebug); (d) Chazara briseis (butterfly); (e) Dendrocopos minor (lesser spotted woodpecker); (f) Accipiter gentilis (goshawk). Re-drawn from Biedermann (2003).

Conservation planning in a changing world cautions that the relationship was based on analyses for species more or less restricted to the habitat patches considered, and that we should not consider as granted a similar relationship for generalist species. If incidence functions really reflect key controlling variables, then they might be of great value in designing reserve networks but, if they are found to be inconsistent across the range or through time, they will need more careful interpretation. Empirical work suggests that, in practice, they do vary in both time and space. In illustration, Hinsley et al.’s (1996) study of 31 woodland bird species in 151 woods in a lowland arable landscape in eastern England over three consecutive years has shown that the incidence functions vary through time in relation to density-independent mortality (extremes of weather conditions). Additionally, Hinsley et al. (1996) showed that specialist species were more likely to disappear from small woods after severe winter weather than were generalists, and that they could take more than a year to recolonize. We may interpret these patterns as reflecting underlying metapopulation dynamic processes, discussed in the following section. An illustration that incidence functions might vary across the range of a species comes from another study of woodland birds, this time undertaken in Australia and based on data from three different landscapes located quite near one another (and thus within the same biogeographical context and climate regime). The study, by Watson et al. (2005), demonstrated that area- and isolation-based incidence functions differed significantly, seemingly as a function of differences in properties of the landscape matrix within which the woodlands were embedded. The three landscapes were an urban area, a peri-urban area and a rural (agricultural) landscape. Interestingly, it was evident that while some species were able to occupy smaller woodlands within the rural landscape, others actually showed a higher incidence in small woods in the urban area (within the city of Canberra itself). This provides some indication of the difficulty of designing a reserve network system optimized for all members of the community of interest (see also Magle et al., 2009). Recently, Prugh et al. (2008) compiled occupancy data for 1,015 bird, mammal, reptile, amphibian and invertebrate populations from 89 case studies, including in total 12,370 habitat patches that were embedded within unsuitable matrix of land cover on six continents. Using incidence functions, they evaluated

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the predictive ability of patch area and isolation for species occupancy. Surprisingly, both area and isolation performed poorly as predictors. Prugh et al. (2008) concluded that is the type of land cover separating patches that most strongly affects the sensitivity of species to patch area and isolation. Thus, although patch size and isolation are indeed important for the occupancy of many species, they find, as do Watson et al. (2005) in their study, that it is crucial to take account of the properties of the intervening matrix. Hence, a key conclusion of this work is that improving the quality of matrix may result in ‘higher conservation returns than manipulating the size and configuration of remnant patches for many of the species that persist in the aftermath of habitat destruction’ (Prugh et al., 2008, p. 20,770).

8.3.2 Metapopulation dynamics Plant and animal species are typically patchily distributed; indeed all species’ ranges involve discontinuities, and especially so at finer scales of analysis. It is frequently possible to discern that within a landscape, a particular species occupies geographically separated patches that are interconnected by occasional movements of individuals and gametes. The name for this network of local populations is a metapopulation. The first metapopulation models were constructed by Richard Levins in papers published in 1969 and 1970 (Gotelli, 1991). The basic idea can be understood as follows: imagine that you have a collection of populations, each existing on patches of suitable habitat. Each patch is separated from other nearby habitat patches by unsuitable terrain. Although these separate populations each have their own essentially independent dynamics, as soon as one crashes to a low level, or indeed disappears, that patch will provide relatively uncontested space for ‘surplus’ individuals from one of the nearby patches, which will soon colonize the nowunpopulated patch. For example, Lei and Hanski (1997) studied metapopulation structure in a threatened species of butterfly, Melitaea cinxia, and its specialist parasitoid, Cotesia melitaearum, in a large network of small habitat patches. They observed that the incidence of the parasitoid in host populations was positively correlated with the size of the host population and the area of the habitat patch. C. melitaearum is thus expected to have a substantial risk of extinction from patches in which

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the number of host populations is small, meaning that the parasitoid may well go entirely extinct from certain patches. However, the network of patches provides the possibility of recolonization. Metapopulation theory, therefore, examines the dynamics of sets of semi-independent populations connected by dispersal (Hanski & Gilpin, 1991). In Levins’s (1970) model, a metapopulation is a network of extinction-prone subpopulations of a species occupying a variety of habitat patches. These subpopulations inhabit identical patches and are subject to equal but independent probabilities of extinction and recolonization. In practice, habitat patches and the landscapes in which they are embedded are very much more complex and heterogeneous than this, so a key challenge for metapopulation modellers is to develop models that are balanced between the attractive simplicity of the general model and fine-tuning to such a degree that models are restricted in application to a single system (see case studies in Whittaker & Fernández-Palacios, 2007). Sometimes, conservation scientists have suggested managing endangered species via policies that encourage the populations to function as metapopulations, thus allowing for the idea that a mixed-use landscape could be worth conserving, as opposed, for example, to insisting that a large area should remain as, or be restored entirely to, a forest cover. However, a spatial model created by Lamberson et al. (1992), in order to predict how the populations of the northern spotted owl (Strix occidentalis caurina) will survive in patches surrounded by logged forest, eventually failed to predict realistic minimum viable populations of the bird (Harrison et al., 1993). The populations of the bird declined in a pattern not predicted by the metapopulation models. On the other hand, some butterfly species have been found to behave according to the predictions of metapopulation models (Thomas & Hanski, 1997). Therefore, the more basic question is: how broadly and to which species does the metapopulation theory apply in habitat fragments? According to Harrison & Taylor (1997) and Hoopes & Harrison (1998) four scenarios of landscape structure are common in fragmented landscapes (Figure 8.9): 1 where patches are roughly of equal size and dispersal distances are comparable to the distances between patches (classic metapopulation models may apply);

2 where patches are so unequal in size and/or habitat quality that most immigrations are in one direction (from large to small patches); extinctions and recolonizations that occur in very small populations are inconsequential (mainland/island metapopulation models may apply); 3 where patches are so close together relative to dispersal distances that they support a single population and not a metapopulation (patchy populations); 4 where patches are so far away relative to dispersal distances that the populations are not interconnected and the assembly ceases to be a metapopulation (nonequilibrium metapopulation models may apply). Thus, whereas patches of tall forest in a savanna landscape may be treated within the framework of metapopulation theory, the concept may not be suitable for forest patches in a highly tree-covered landscape. This is because ecological boundaries between forest and savanna are clear-cut, whereas those between forest and highly tree-covered landscape are fuzzy. Furthermore, an insurmountable barrier for one group of organisms may be easily navigated by another – for sunbirds, forest patches spread over a 100 km2

Figure 8.9 Structures of metapopulations that can arise from fragmentation. Adapted from Hoopes & Harrison, 1998; after Harrison (1991). The four cases are those described in the text.

Conservation planning in a changing world landscape may sustain a metapopulation but, for dispersal-limited snails, they can at best sustain isolated populations. Metapopulation models, therefore, are not generally applicable to all organisms in fragmented systems (Fahrig & Paloheimo, 1988). Hoopes & Harrison (1998) caution against the general use of such models in conservation decision-making because of the prevalence of situations where functional metapopulation dynamics either do not occur, or where they fail to match the assumptions of the models. The following are additional important shortcomings of the metapopulation approach: • Several authors have noted that metapopulation models are extremely data-demanding and usually require data that are very difficult to obtain (Kindvall & Ahlén, 1992; Doak & Mills, 1994). Moreover, model results tend to be very sensitive to poorly estimated parameters, and the predictions of such models have therefore frequently been found to be inaccurate (e.g. Harrison et al., 1993; Wilson et al., 1994). • Most empirical examples of metapopulations pertain to single species or a group of interacting species (Hanski & Gilpin, 1991), but not to multi-species ecological communities. • Most metapopulation models assume no distance effects (Fahrig & Merriam, 1994) although, in practice, dispersal abilities vary from species to species. For instance, metapopulations of frogs may be influenced by the availability of suitable habitat in the surrounding ≈500 m, whereas for birds this distance may be ≈3 km, because of large differences in mobility resulting in the different abilities of frogs and birds to disperse. The issue of ‘scale’ has therefore been considered important in studying fragmented landscapes (e.g. Doak et al., 1992) – an issue which classic metapopulation models do not address. • It is important for conservationists to recognize that many local populations may not be at equilibrium and regional processes may be critical in sustaining metapopulations (Hanski 1999). In conclusion, metapopulation theory offers a useful framework for thinking about isolation and fragmentation (Hanski, 1999) but, if the concept is to be useful as a theoretical framework for conservation decisionmaking, it must be extended from the original simplistic models to allow for the differing degrees of population connectivity in fragmented landscapes and differing forms of inter-patch relationships, as in realworld systems.

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8. 4 N ES T EDN ES S The concept of nestedness was first introduced some 70 years ago (see Ulrich et al., 2009) to describe patterns of species composition within continental biotas and among isolated habitats such as islands and landscape fragments. In a perfect nested pattern, when a set of habitat patches is ordered by increasing species richness, it will be found that the smallest assemblages make up a subset of the species found in the next larger assemblage, and so on, throughout the series (see Figure 8.10). Nestedness is thus a particular form of nonrandomness of assemblage composition across a set of isolates. Any such non-random pattern is potentially of interest to conservation biogeographers, as it may inform judgements about the design of protected area systems within landscapes and regions. Nestedness may, in theory, arise from differential dispersal and colonization abilities (especially for young islands); or differential rates of extinction (e.g. for land bridge islands or newly fragmented habitat islands); or from a strong nestedness of habitat types with increasing ‘island’ size (Whittaker & Fernández-Palacios, 2007); or possibly from other processes (Table 8.3). Nestedness analyses became popular among ecologists and biogeographers only after Patterson and Atmar (1986) developed a statistically rigorous approach for analysing nested subsets. They were interested in nestedness patterns derived by extinction of species from land bridge islands, and their metric reflects this emphasis. They proposed that nestedness patterns for such islands most likely reflect orderly sequences of extinctions on such islands and in fragmented landscapes (see their Fig. 4). They introduced an intuitive ‘matrix temperature’ metric to quantify the pattern of nestedness. Hot matrices are those with more random presences of species and cool matrices are those where species presences are more nested. The matrix temperature could be calculated with a software package, The Nestedness Temperature Calculator (Atmar & Patterson, 1993, 1995). The nestedness concept, as applied by Patterson and Atmar, is based on ordering the data matrix by the size of fauna or flora, i.e. it is richness-ordered nestedness. Some authors, however, have ordered the data matrix not by species richness but by island area, which has been termed area-ordered nestedness, or even by island isolation, i.e. distance-ordered nestedness (e.g. Lomolino & Davis, 1997; Whittaker &

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Figure 8.10 Nested subset relationships. Circles represent islands of different size and letters represent species. Circle size is positively correlated with species richness. The left biota is perfectly nested; all the species present in relatively species-poor assemblages are present in relatively species-rich assemblages. The right biota is perfectly non-nested because none of the species in the species-poor assemblages is present in the richer ones. Note that although the species richness of the three islands is the same in the two cases, the overall species richness of the non-nested set is greater than of the nested system.

Table 8.3 Causes of nested subset patterns (adapted from Ulrich et al., 2009). Assumption/precondition

Hypothesis

Site and species properties

Gradient of:

Explanation/Example

Passive sampling

carrying capacities of sites

regional abundance

Species are drawn randomly from the pool with the constraint that the availability of propagules is itself strongly non-random (e.g. Higgins et al., 2006).

Neutrality

carrying capacities of sites

dispersal ability

The availability of propagules is random and species’ dispersal ability is driving the pattern (e.g. Ulrich & Zalewski, 2007).

Selective colonization

isolation

dispersal ability

There are predictable limits to species’ dispersal abilities. The system consists of islands ‘sampling’ a series of species’ isolation/incidence functions (e.g. Darlington, 1957; Patterson, 1990).

Selective extinction

carrying capacities of sites

extinction susceptibility

Selective occupancy of sites according to the area of sites, which sets their carrying capacity. Relaxation in the case of mainland/habitat islands (e.g. Patterson & Atmar, 1986).

Nested habitats

habitat heterogeneity

degrees of specialization

Absence of certain habitat types in smaller and/or resource-poor patches. Higher proportion of generalist species in smaller and/or resource-poor patches (Wright & Reeves, 1992).

Selective environmental tolerances

environmental harshness, environmental tolerances

Selective occupancy of sites according to species tolerance of environmental stress (e.g. Blake, 1991).

Habitat quality

environmental harshness

Species are distributed according to the harshness exhibited by patches of the same habitat (e.g. Bloch et al., 2007).

Conservation planning in a changing world Fernández-Palacios, 2007). Intuitively, however, ordering by species richness would appear the most appropriate approach. Apart from sequential extinctions, a variety of different mechanisms can also produce nestedness patterns (see Table 8.3), some of which are deterministic and some of which are stochastic, requiring different metrics for quantifying nestedness (Wright et al., 1998; Ulrich et al., 2009). All of the explanations for nested subsets can be seen as variations of ordered colonizations or extinctions along environmental or biological gradients (area, isolation, habitat) of the target areas. Frequently, these mechanisms cannot be distinguished by just establishing the statistical pattern of nestedness. Inferences of causation ideally require independent lines of verification beyond manipulations and analyses of the original presence/absence matrix (Ulrich et al., 2009). Although nestedness can be driven by a number of processes, it appears that differential extinction plays a major role in producing nested structure in many habitat island data sets (Wright et al., 1998). Knowledge of nested subset structure might therefore provide a basis for predicting the ultimate community composition of a fragmented landscape, particularly if it is possible to attribute patterns to particular causes (Fischer & Lindenmayer, 2005; Fleishman et al., 2007). Feeley’s (2003) study of bird communities inhabiting recently isolated land bridge islands in Lago Guri, Venezuela, showed how nestedness calculations can provide useful insights. Lago Guri is a large hydroelectric reservoir created in 1986 in east-central Venezuela. The inundation of an area of hilly terrain expanding over 4,000 km2 resulted in the fragmentation of oncecontinuous forest into hundreds of land bridge islands (e.g. Terborgh et al., 2001). Feeley found that the resident forest-interior bird communities displayed a significantly nested distributional pattern that was hypothesized to be the result of species’ differential extinction rates. In an earlier study of forest birds, Blake (1991) also found a significant degree of nestedness, particularly among birds breeding in the forest interior and among species wintering in the tropics. By contrast, species breeding in forest-edge habitat showed more variable distribution patterns. These findings concur with those of Patterson (1990) from São Paulo, Brazil (original data from Willis, 1979). Patterson reported significant nestedness amongst sedentary bird species but, when

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transient species were also included, the system as a whole was found to be non-nested. These results are indicative of a large number of studies of nestedness, which show the outcome of nestedness analyses to be variable across different systems and for different ecological groups of species, but which show that significant nestedness is a common pattern. Such analyses often indicate that species that are restrictive habitat specialists, including many of high conservation value, do require larger, more species-rich patches (Fleishman et al., 2007; Whittaker & Fernández-Palacios, 2007). In theory, a nestedness analysis can contribute a simple answer to the SLOSS question, as a strong degree of nestedness implies that most species could be represented by conserving the richest (largest) patch. According to Atmar & Patterson (1993), the widespread occurrence of nested subsets speaks for the value of larger protected areas. However, Boecklen (1997) and Fischer & Lindenmayer (2005) convincingly showed that this argument is only valid for perfectly nested subsets, which are very rare in nature. Even for highly significantly (but not perfectly) nested subsets, the total species numbers from subsets of many smaller sites are often higher than the respective number of species from a single larger site of the equivalent total area (Ulrich et al., 2009). On the other hand, a low degree of nestedness may be considered as indicative that specific habitat patches are sampling distinct species sets, and thus an array of reserves of differing size and internal richness may be required to maximize regional diversity in such circumstances (e.g. Kellman, 1996). Broadly speaking, a nestedness index can provide one compositional descriptor and can perhaps aid identification of risk-prone species. However, it should not be given primacy in conservation planning. Identifying a community as nested at a certain point in time, has limited predictive ability as to the probability of the community maintaining the same sets of species (or even a single species) over time (Simberloff & Martin, 1991). The isolates may be subject to turnover and/or species attrition in new ways dictated by the changing biogeographical circumstances of the landscape in which the fragments occur. As Worthen (1996, p. 419) put it, nestedness is not a ‘magic bullet’, ‘ … no single index should be expected to distil the informational content of an entire community, let alone predict how it will react to habitat reduction or fragmentation’.

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According to Ulrich et al. (2009), there are three key steps in a nestedness analysis: 1 calculation of a metric to quantify the pattern of nestedness in a matrix; 2 comparison with an appropriate null model or randomization test to assess the statistical significance of the metric; 3 inference of the mechanism that generated the pattern of nestedness. Unfortunately, on all three points, no consensus has yet been reached among scientists, which has hindered a general understanding of the frequency, causes, and consequences of nestedness (Whittaker & FernándezPalacios, 2007; Ulrich & Gotelli, 2007; Almeida-Neto et al., 2008). This lack of consensus, along with the growing number of applied and theoretical studies of nestedness, therefore calls for a critical review of the state of the art and for perspectives for future research. This review should take into account other patterns related to but distinct from nestedness patterns, such as island assembly theory (Diamond, 1975b) and species incidence (see also Section 8.3).

Figure 8.11 Bird species richness–area relationships in the littoral forests of southeastern Madagascar, including regression lines and r2 values. Two classifications of species richness were considered: total species richness (closed circles) and forest-dependent species richness (open circles). Linear regressions: unbroken lines; break-point regression: dashed line. The break-point regression procedure followed Lomolino & Weiser (2001). All regressions are significant (P < 0.01). From Watson et al. (2004).

8.4.1 Edge effects Where two habitats abut, they often intermingle, forming a zone of species overlap (and of locally higher diversity) – a pattern termed an ecotone. In the case of many protected areas, habitat alteration often produces quite sharp ecotones or edges, but it is often the case that species numbers are elevated in these edge habitats (Kellman, 1996). However, many of these species are dependent on the matrix habitat rather than on the habitat within reserves. Such species are unlikely to be those that are most in need of protection. Watson et al. (2004) studied birds in littoral forest habitat islands and surrounding habitats in southeastern Madagascar. Core forest locations were found to be richer than edge or matrix habitats, with some 68 per cent of the forest dependent species found to be edge-sensitive. Frugivorous species and canopy insectivores were generally edge-sensitive, while sallying insectivores preferred edges. The vegetation structure at remnant edges contributed to edge-sensitivity. The relationship between fragment area and overall species richness conceals the fact that forest-dependent species were generally lacking from fragments of less than 10 ha (Figure 8.11).

Wilcove et al. (1986) suggest that reserves of less than 100 ha cannot support viable populations of forest songbirds due to high densities of nest predators such as blue jay (Cyanocitta cristata), weasel (Mustela erminea) and racoon (Procyon lotor) around forest edges. Laurance (2000) suggests that edge effects can occur on even large spatial scales. For example, Curran et al. (1999) found that recruitment of canopy trees in the 90,000 ha Gunung Palung National Park in western Borneo collapsed because vertebrate seed predators flooded into the park from surrounding degraded areas. A core-area model proposed by Laurance (2000) illustrates the impacts of edge effects on nature reserves ranging from 1,000 to 100,000 ha (Figure 8.12). These examples illustrate that the relationship between a reserve and its surrounding matrix is not subject to easy generalization. There are species that share both zones and, just as there are matrix species that may impact negatively upon core reserve species, there may also be reserve species which exploit resources in the matrix. Therefore, the heterogeneous

Conservation planning in a changing world

Figure 8.12 A core-area model illustrating the impacts of edge effects on nature reserves ranging from 1000 to 100,000 ha. The curves show the percentage of the reserve’s total area that is influenced by edge effects that penetrate to distances of 100 m (dotted line), 500 m (dashed line) or 2 km (solid line) inside the reserve. For an edge effect that penetrates to 5 km (not shown), the reserve would need to be approximately 650,000 ha in size to ensure that half of its area is free from edge effects. Source: Laurance (2000).

nature of habitats within reserves needs to be taken into account when understanding patterns of species distribution and habitat suitability.

8.4.2 Habitat corridors Habitat connectivity can be achieved by ‘stepping stones’ or ‘corridors’ of suitable habitat linking larger reserves together. In addition to forest peninsulas or hedgerows, other linear landscape features such as rivers, roads, and railways may act as conduits for the movement of particular species. However, for others they may represent barriers or hazards (Reijnen et al., 1996). Therefore, habitat corridors act as differential filters, enabling the movement of some species but being of little value, or presenting an impediment, to others (Table 8.4). A useful illustration of how corridors can be beneficial comes from the study by Saunders and Hobbs (1989; from Whittaker & Fernández-Palacios, 2007)

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of Carnaby’s cockatoo (Calyptorhyncus funereus latirostrus) from the Western Australian wheat belt – an area of 140,000 km2 in the south-west of the state, 90 per cent of which has been cleared for agriculture. The Carnaby’s cockatoo is one of Australia’s largest and most striking parrots and was once the most widely distributed cockatoo in the region. The widespread clearance of the native forest has removed extensive areas of their foraging and breeding habitat, replacing it with annual crops of no value to the species. In more recently cleared areas, however, wide verges of native vegetation have been left uncleared along the roads. These act to channel the cockatoos to other areas where food is available. Cockatoos have not persisted in areas of earlier clearances that were carried out without these connecting strips because, once they run out of a patch of acceptable habitat, it takes a long time for the flock to find another patch of native vegetation. The big reduction in suitable habitat across the region is fairly recent, and the cockatoo is not yet in equilibrium with the new regime (and indeed is considered to be an endangered species). So, it is not clear yet if the degree of connectivity and remaining area of woodland habitat are sufficient for the long-term persistence of this cockatoo. Some scientists argue that the requirement of corridors for faunal movement may have been overstated and that corridors may not be required for many taxa (see discussion in Simberloff et al., 1992). While movement along corridors is frequently assumed to occur, there have been relatively few studies which have shown that corridors are actually required for movement (Hobbs, 1992). Some studies of marked or radiotagged animals, however, have provided clear indication that certain species use corridors for movement (e.g. Dmowski & Kozakiewicz, 1990; Merriam & Lanoue, 1990), as do observations such as those above for the Carnaby’s cockatoo. In historical biogeography, the term ‘corridor’ is used for very broad connecting areas between regions, which are assumed to provide relatively unfettered movement between then. However, in considering habitat corridors at finer scales, within landscapes, and given that each species has its own requirements for habitat, its own ability to move and its own behaviour, few corridors can be considered all-purpose (Dawson, 1994). Rather, like other elements of the landscape matrix, habitat corridors act as filters. Many rare and threatened species are unlikely to benefit from

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Table 8.4 Advantages and disadvantages of habitat corridors (adapted from Noss, 1987). Potential advantages of corridors

Potential disadvantages of corridors

1 Increase immigration rate to a reserve, which could: a increase or maintain species richness and diversity, as predicted by the equilibrium theory of island biogeography; b increase population sizes of particular species and decrease probability of extinction (rescue effect) and/or permit establishment of extinct local populations; c prevent inbreeding depression and maintain genetic variation within populations. 2 Provide increased foraging area for wide-ranging species. 3 Provide predator-escape cover for movement between patches. 4 Provide a mix of habitats for different activities or stages of their life-cycles. 5 Provide alternative refugia from large disturbances (a ‘fire escape’). 6 Provide ‘green belts’ to limit urban sprawl, abate pollution, provide recreational opportunities and enhance scenery and land values.

1 Increase immigration rate to a reserve, which could: a facilitate the spread of epidemic diseases, insect pests, exotic species, weeds, and other undesirable species into reserves and across landscapes; b decrease the level of genetic variation among populations or subpopulations, or disrupt local adaptation and coadapted gene complexes (‘outbreeding depression’). 2 Facilitate spread of fire and other abiotic disturbances (‘contagious catastrophes’). 3 Increase exposure to wildlife hunters, poachers and other predators. 4 Riparian strips, often recommended as corridor sites, might not enhance dispersal or survival of upland species. 5 Cost and conflicts with conventional land preservation strategy to preserve endangered species’ habitat (when inherent quality of corridor habitat is low).

corridors, because their specialist habitats are unlikely to be found throughout the length of most corridors. For some populations, corridors may even act as ‘sinks’, drawing out individuals from the main habitat area, perhaps into dangerous places with higher risks of predation, but not returning individuals to supplement the main source area. Alternatively, they may be fairly neutral in their ecological cost-benefit, but perhaps be quite expensive to purchase and set up if not already existing in a landscape. On the other hand, some corridors are essential in providing links between preferred habitats for animals that undertake regular seasonal migrations. In the longer term, it has been argued that climate change is likely to drive substantial shifts in the distribution of species, and that the resulting species migrations will be impeded by the human sequestration of land to agriculture and other purposes. Therefore, on these grounds, it would seem prudent to plan more or less continuous habitat corridors that straddle major climatic/elevational gradients where this is feasible (e.g. Bush, 1996, 2002 – Box 7.4).

8.4.3 Landscape context – matrix effects The presence of a species within a reserve will depend not only on the suitability of habitat within the reserve, but also on the species’ ability to use the intervening landscape matrix. If this matrix is hospitable, species can also move between reserves (Gustafson & Gardner, 1996). Therefore, although reserves are important, an increasing emphasis is now being placed on the role of the quality of the landscape matrix between reserves. For example, Baum et al. (2004) have demonstrated that corridors and stepping stones are more effective if surrounded by a hospitable matrix. In tall-grass prairie ecosystem in the central USA, they showed that the effectiveness of corridors and stepping stones for promoting dispersal of the planthopper Prokelisia crocea among patches containing prairie cordgrass Spartina pectinata (the sole host plant for the planthopper) depended strongly on the intervening matrix habitat. In a low-resistance matrix (one that facilitates high rates of inter-patch dispersal), where both stepping stones and corridors promoted high connectivity, the

Conservation planning in a changing world number of planthopper colonists increased by threefold relative to patches separated by matrix habitat only. The effectiveness of stepping stones and corridors was significantly lower in a high-resistance matrix (one that provides only low rates of interpatch dispersal), with stepping stones failing to improve connectivity for the planthoppers relative to controls (Baum et al., 2004). To test whether the type of interpatch matrix can contribute significantly to patch isolation, Ricketts (2001) conducted a mark–recapture study on a butterfly community inhabiting meadows in a naturally patchy landscape. The relative resistances of the two major matrix types (willow thicket and conifer forest) to butterfly movement between meadow patches were estimated. For four of the six butterfly taxa (subfamilies or tribes) studied, conifer forest was 3–12 times more resistant than willow thicket. For the two remaining taxa (the most vagile and least vagile in the community), resistance estimates for the two differing matrix types were not significantly different, indicating that responses to matrix differ even among closely related species. These results suggest that the surrounding matrix can significantly influence the ‘effective isolation’ of habitat patches, rendering them more or less isolated than patch size and/or isolation would indicate (as Figure 8.13). In a study conducted in grasslands in western Victoria, Australia, Williams et al. (2006) assessed how both the spatial attributes of remnant patches (area and isolation) and the landscape factors (extent of urbanization and maximum inter-fire interval) influence the persistence of native plant species. They found that, on average, 26 per cent of populations of native species distributed across 30 remnants became locally extinct between the 1980s and 2001. While area and isolation had little effect on the probability of local extinction, urbanization and longer maximum interfire intervals corresponded with increased extinction risk (Williams et al., 2006). Most nature reserves are now surrounded by human-dominated landscape matrix, which is often inhospitable to many species. For example, intensively cultivated agriculture landscape is unable to support forest-dwelling species. Modification of this agricultural matrix, therefore, may provide opportunities for reducing patch isolation and thus the extinction risk of populations in fragmented landscapes. Tropical agro-forestry systems, where crops are grown under the shade of native tree species, often

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Figure 8.13 A modified species incidence function for a hypothetical species in a series of habitat islands. The occupancy of the species depends primarily on the area and isolation of the habitat island but also varies between Landscape A and Landscape B as a function of the quality of the matrix habitat. Black circles indicate occupied habitat islands and white cells unoccupied habitat islands. The grey remnants and solid line indicate that a species would inhabit these remnants when in a landscape with matrix composition ‘B’ (favourable) but would not in matrix composition ‘A’ (less favourable; dashed line). From Whittaker et al. (2005) and based on original ideas developed by Mark V. Lomolino and James E. Watson.

provide matrix habitats suitable for a substantial proportion of native species, although typically not for certain habitat-specialist forest dwellers (Bhagwat et al., 2008). Therefore, it can be argued that for successful conservation within reserves, a wholelandscape approach is needed that accounts for maintaining suitable matrix habitat.

8. 5 EMER GEN T GU I DELI N ES FOR CON S ER V AT I ON Theories are nets cast to catch what we call ‘the world’: to rationalize, to explain, and to master it. We endeavour to make the mesh ever finer and finer. (Karl R. Popper, 1959, p. 59) MacArthur & Wilson (1967) and MacArthur (1972) have described ‘habitat patches’ such as farmer’s woodlots surrounded by fields and recent fire burns, as

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‘islands’, but they carefully differentiate them from true islands. MacArthur (1972, p. 105) pointed out that true islands are ‘separated by a vacuum insofar as land birds and insects are concerned’, whereas habitat islands are ‘separated by other habitats filled with birds and insects’, thus the spill-over of organisms from adjacent habitats is a primary factor for habitat islands. Island biogeography theory and the subsequent theories and applications it has inspired and influenced have made an important contribution to conservation biogeography. The theory has inspired much thinking about the importance of the size and connectivity of protected areas in the maintenance of species diversity, and it has stimulated an avalanche of research on fragmented ecosystems. However, generalizations derived from this theory have given rise to models that are too simplistic (e.g. Laurance, 2008). Recent advances in island theory demonstrate that we are moving towards a new synthesis, identifying and incorporating aspects of the island systems that were not considered in the past. For example: i Within oceanic island biogeography, efforts have recently been made to adjust the MacArthur–Wilson (1967) model to accommodate the dramatic changes in the carrying capacity and environmental characteristics of islands that occur through the life history (ontogeny) of an oceanic island itself (see e.g. Whittaker et al., 2008, 2010). ii Application of genetic analyses are producing a more nuanced grasp of species and gene flow between insular and mainland habitats. iii Scale-dependency of isolation and fragmentation effects are beginning to be quantified. iv Efforts have been made to incorporate matrix effects and to consider the implications of longer term changes within habitat islands post-isolation. v Assumptions of initial equilibrium in prefragmentation landscapes have been challenged (for discussion and exemplification of the foregoing, see Whittaker & Fernández-Palacios, 2007). These considerations suggest that the dynamic process at the heart of the island equilibrium theory needs to be embedded in a much more dynamic model of the physical environment (a point argued more generally in Chapter 7). We have selected four key areas that we consider important for a more successful application of island theory to conservation biogeography. These include examination of (1) the life history of habitat islands,

(2) threshold effects, (3) assembly rules, and (4) the role of the matrix for conservation in habitat islands. 1 Life history of habitat islands: As Whittaker et al. (2005) comment: ‘It is disappointing that we still know so little about the power and timescale of “species relaxation”.’ Here we suggest that the consideration of the life history (ontogeny) of habitat islands could be particularly insightful in revealing the patterns and processes shaping species richness, species assembly and disassembly. Although a number of theoretical frameworks have been put forward to describe the sequential process of species relaxation after habitat loss and fragmentation (Section 8.2.2), the temporal scale of habitat loss and fragmentation has received the least attention. This has restricted our knowledge on how, for example, the abiotic characteristics of a fragment (e.g. net primary productivity) and rates of nutrient cycling change through time after its isolation, and how this affects the fragment’s capacity in maintaining biodiversity. Habitat conversion is almost always a non-random process (e.g. Raheem et al., 2009). In forest landscapes, for example, the most accessible and productive areas tend to be deforested first. Thus, the remaining fragments show a non-random spatial distribution with respect to age, because the geographical distribution of older fragments (i.e. isolated earlier) is different from that of those isolated later. Moreover, other environmental factors, ranging from anthropogenic disturbance (e.g. hunting) to physical gradients (e.g. topography and climate) may be correlated with fragmentation and forest loss (Laurance et al., 2002). We believe that the integration of research on the ontogeny of habitat islands will help us towards estimating more accurately the rates at which species extinctions are likely to occur. The time-lags and ‘extinction debt’ involved in such extinction processes are still poorly explored and in need of much attention (Box 8.2; Tilman et al., 1994). By focusing further work on the above questions, we will be able to approach more analytically questions related to the time-lag for relaxation and extinction debt. 2 Thresholds: Taxonand system-dependent thresholds, beyond which species losses accelerate (see Ewers & Didham 2006; Whittaker & FernándezPalacios, 2007; Suding & Hobbs, 2009) have received very limited attention. Analyses of critical value ranges, where even small changes in environmental variable(s) will lead to large changes in the system, will help us towards understanding relaxation as a result

Conservation planning in a changing world of habitat loss and fragmentation (Simberloff & Martin, 1991; Laurance 2002). A highly relevant island phenomenon is the socalled ‘small island effect’ (see Lomolino & Weiser, 2001; Triantis et al., 2006). The main feature of the phenomenon is the absence of the commonly found relationship of island area and species richness below a certain island size (dashed line in Figure 8.11). The particular threshold of this effect appears to vary depending on the taxon and archipelago selected, but it generally appears to occur only with islands of a very small size and diversity. In practice, within the limits of the small island effect, species richness is independent from the direct effects of area and is mainly driven by the effects of habitat diversity. Hence, it would be interesting to assess the existence of such thresholds in habitat island data sets for which the usual explanatory variables – such as area and isolation – are not important (see Prugh et al., 2008) and other variables – such as island age, productivity, energy and environmental heterogeneity – are important. The consideration of such variables, although challenging, is necessary if we are to build up a more predictive science of species richness variation across true and habitat island systems. In a fragmented landscape, species can either become extinct or go through changes in life history traits that will adapt them to the changed living conditions. Another issue related to spatial thresholds of fragmented landscapes that has received limited attention is how the evolutionary dynamics of species change in response to landscape transformation. Adaptation in a fragmented landscape may influence measurable features of the phenotype of a species, e.g. body size. In island studies, it is well established that islands favour the change of species body size, compared to their mainland counterparts; usually small species become larger (gigantism) and large species smaller (nanism) – a phenomenon termed the ‘island rule’ (Lomolino, 1985). These changes lead to a more effective exploitation of the available resources in the context of the limited available space on islands. The absence of the full collection of competitors and predators found on the mainland contributes towards these size changes (see Lomolino, 2005; Lomolino et al., 2006; but see Meiri et al., 2006). In illustration of these effects within a habitat island context, Schmidt & Jensen (2003) studied the body size changes within the entire Danish mammalian community during the last 175 years. They found that the

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rate of body length change was increased for both smaller and larger mammals, while it was lower for the medium-sized species. Following the general trend of the island rule, small mammals have generally increased, whereas large mammals have decreased in length. Schmidt & Jensen suggested that the major, but not the only, driver of these changes was habitat fragmentation. Based on island studies, Losos & Schluter (2000) have identified that for Anolis lizards in the Great Antilles, below a certain island size threshold there is little or no cladogenesis. The identification of such size thresholds not only in the short term, but over evolutionary timescales, could be quite insightful for conservation biogeography (e.g. Triantis et al., 2008). This corresponds to the plea of Gunderson & Folke (2003), who called upon conservation biologists to work towards the ‘science of the long view’ and to integrate insights from other disciplines in the search for new predictive and transcalar models in time and space (see also Lomolino, 2006). 3 Assembly rules (phylogeny): Island biotas are not simply random draws from regional species pools. Instead, they typically exhibit compositional structure: some species, species combinations, or species types, are found more frequently, and some less frequently, than might be expected by chance. This idea was presented in Jared Diamond’s island assembly theory (Diamond, 1975b; reviewed in Whittaker & FernándezPalacios, 2007). Related to island assembly theory is an increasing number of studies appearing to show deterministic patterns of evolution on islands, i.e. independent evolutionary diversification events, producing on different islands the same set of habitat specialists adapted to use different parts of the environment (see Losos et al., 1998; Chiba, 2004; Gillespie, 2004; Losos & Ricklefs, 2010). The incorporation of phylogenetics into community ecology will offer key insights into the assembly and structure of communities (see Webb et al., 2002; Emerson & Gillespie, 2008), with insular systems having a pivotal contribution to make within this research programme. Extinction and extinction risk are often phylogenetically non-random (Purvis, 2008). Nonrandomness when species are faced with a similar threat intensity indicates that some species are more extinction resistant than others (e.g. Purvis et al., 2000). Hence the use of phylogenies for identification of those traits that are associated with a high extinction risk in declining species, e.g. high trophic level,

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low population density, slow life history, small geographical range, ‘ecological naivety’, is currently one of the great challenges of conservation biology. Studies on islands, nature’s test tubes and the location of a high proportion of globally threatened species, will certainly offer significant insights in this research program. 4 Matrix: The number of species held in a reserve (or reserve system) is actually less important than the conservation of those species which cannot survive outside the remnants (e.g. Newmark, 1991). Some recent efforts have been made to move beyond an exclusive focus on (forest) fragments and towards understanding the role of such habitat islands within mixed-use landscapes. This switch in emphasis comes under varying headers. For example, Watson et al. (2005) show that the incidence functions of woodland bird species in three different landscapes in the Canberra area of Australia differ significantly, seemingly as a function of differences in properties of the landscape matrix within which the woodlands are embedded. Hence, Watson et al. (2005) join others (e.g. Ewers & Didham, 2006) in calling for greater attention to ‘matrix effects’. J.B. Hughes et al. (2002) adopted a slightly different approach within their study in southern Costa Rica, focusing on the extent to which native forest species make use of the surrounding countryside. They found that some 46 per cent of bird species foraged often kilometres away from extensive areas of native forest. Although they stress that not all species can be so readily accommodated outside large tracts of native forests, their work supports the importance of developing ‘countryside’ landscapes that are biodiversityfriendly and penetrable by native fauna (as Harris, 1984). Daily and colleagues (e.g. Daily et al., 2001, 2003) coin the term ‘countryside biogeography’ for this switch in attention from remnants per se to the way in which remnants function within whole landscapes. This switch in emphasis is similar to that promoted by Rosenzweig (2003) under the heading ‘reconciliation ecology’. But whether we label it ‘matrix effects’, ‘countryside biogeography’ or ‘reconciliation ecology’, the common element is a realization that effective conservation must include consideration of what happens outside reserves. The way we shape the countryside, whether we farm intensively or extensively, whether we retain hedgerows and trees within mixed landscapes, can all have profound implications for regional diver-

sity and for abundances of wildlife (e.g. see Gascon et al., 1999; Gates & Donald, 2000). Conservation requires pragmatic decision-making. As we continue to fragment landscapes, island effects may inform such decision-making, but should not be oversimplified. There is no single message, and no single island effect; indeed, insularity may sometimes bring positive as well as negative effects (Lockwood & Moulton, 1994). Island effects may be weak or strong. The implications of insularity vary, depending on such factors as the type(s) of organism involved, the type(s) of landscapes involved, the nature of the environmental dynamics, the biogeographical setting and the nature of human use and involvement in the system being fragmented. In closing this chapter, we return to the basic question posed in the introduction: is it realistic to expect habitat islands to behave according to the same principles as real islands? Our answer is yes, but caution is needed in the island theories and models we are using. Island systems of generally restricted spatial extent and most importantly similar age and intrinsic rates of change in time to habitat islands (e.g. Terborgh et al., 2001, 2006; Cody, 2006) will probably continue to offer more relevant principles for the understanding of processes such as relaxation in habitat islands and their more effective preservation. As island biogeography moves towards new syntheses and theories, we anticipate that this body of work will become increasingly helpful for understanding and conserving our natural world.

FOR DI S CU S S I ON 1 In what circumstances are scattered protected areas of modest size better than a few large ones? 2 What is the relationship between the SLOSS debate and nestedness? 3 How important is it to take account of underlying biogeographical structure within a region when applying island models to projecting species extinctions? 4 Is there any optimal slope (z) value for models projecting species losses based on species–area relationships, and what is the relevance of such models if no account is taken of efforts made to mitigate these losses? 5 How important is connectivity between patches for maintaining species diversity in a landscape?

Conservation planning in a changing world 6 In what ways can the landscape outside protected areas be managed to support biological diversity? 7 How far do you agree that the problem for oceanic islands is the loss of their isolation, while within continents the reverse is the case? S U G G ES T E D R E AD I NG Bhagwat, S.A., Willis, K.J., Birks, H.J.B. & Whittaker, R.J. (2008) Agroforestry: a refuge for tropical biodiversity? Trends in Ecology & Evolution, 23, 261–267. Brook, B.W., Sodhi, N.S. & Ng, P.K.L. (2003) Catastrophic extinctions follow deforestation in Singapore. Nature, 424, 420–423. Brook, B.W., Traill, L.W. & Bradshaw, C.J.A. (2006) Minimum viable population sizes and global extinction risk are unrelated. Ecology Letters, 9, 375–382. Emerson, B.C. & Gillespie, R.G. (2008) Phylogenetic analysis of community assembly and structure over space and time. Trends in Ecology & Evolution, 23, 619–630.

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Ewers, R.M. & Didham, R.K. (2006) Confounding factors in the detection of species responses to habitat fragmentation. Biological Reviews, 81, 117–142. Ladle, R.J. (2009) Forecasting extinctions: uncertainties and limitations. Diversity, 1, 133–150. Laurance, W.F. (2008) Theory meets reality: how habitat fragmentation research has transcended island biogeographic theory. Biological Conservation, 141, 1731–1744. Prugh, L.R., Hodges, K.E., Sinclair, A.R.E. & Brashares, J.S. (2008) Effect of habitat area and isolation on fragmented animal populations. Proceedings of the National Academy of Sciences USA, 105, 20770–20775. Rosenzweig, M.L. (2003) Win-win ecology: how the earth’s species can survive in the midst of human enterprise. Oxford University Press, New York. Williams, M.R., Lamont, B.B. & Henstridge, J.D. (2009) Species–area functions revisited. Journal of Biogeography, 36, 1994–2004. Whittaker, R.J. & Fernández-Palacios, J.M. (2007) Island biogeography: ecology, evolution, and conservation, 2nd edn. Oxford University Press, Oxford.

CHAPTER 9 Biological Invasions and the Homogenization of Faunas and Floras Julian D. Olden1, Julie L. Lockwood2, and Catherine L. Parr3 1

School of Aquatic and Fishery Sciences, University of Washington, Seattle, USA Ecology, Evolution, and Natural Resources, Rutgers University, New Brunswick, NJ, USA 3 Environmental Change Institute, School of Geography and the Environment, University of Oxford, Oxford, UK 2

9 . 1 T HE B I OGE OGR AP HY OF S PECI E S I NV AS I ONS In considering the distribution of organic beings over the face of the globe, the first great fact which strikes us is, that neither the similarity nor the dissimilarity of the inhabitants of various regions can be accounted for by their climatal and other physical conditions … A second great fact which strikes us in our general review is, that barriers of any kind, or obstacles to free migration, are related in a close and important manner to differences between the productions of various regions. (Charles Darwin, 1859, pp. 395–396)

9.1.1 The invasion process One of the fundamental elements of life on Earth is change. Species appear through time via evolution and disappear by the natural actions of environmental change (e.g. volcanic eruptions, changing sea levels, glaciation). Species have also regularly shifted their geographical ranges in response to biological and physical forces, sometimes becoming less common and other times becoming more widespread. In general,

however, the large majority of species are not distributed broadly, because individuals of most species have limited dispersal capabilities. These limitations on dispersal ability have produced the interesting phenomenon that many, perhaps even most, species do not occupy all of the areas of the world in which they could quite happily thrive. Instead, they are restricted to certain regions, where they are able to interact with only those species with which they cooccur. The limited geography of species is responsible, in part, for the fantastic array of diversity that presently carpets the Earth, as it provides opportunity for convergent evolution in disparate unconnected regions. With the range expansion of modern humans, initially out of Africa, then across the globe, came the possibility of human-mediated dispersal of a large variety of other species. By this, we mean that humans provided the conduit for individuals of some species to disperse much farther abroad than they could naturally. Species were moved within, or on, humans as parasites or disease organisms, in their household goods as hitchhikers, as their livestock or working animals, as their crop plants, as their pets, and as commodities themselves. There is written evidence that intentional movements of species by humans traces back to ancient

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

Conservation planning in a changing world times, such as the introduction of the tamarind tree (Tamarindus indica) into China by way of commerce along the Shu-Yan trade route that linked China to India 8,000 years ago (Yan et al., 2001). Some species apparently have nearly circumglobal distributions because of ancient trade activities, with many of these examples only recently coming to light thanks to the power of molecular analyses to locate the evolutionary origins of now very widespread species (e.g. Wares et al., 2002). There is ample historical evidence that the number of species that were moved out of their native ranges and introduced to somewhere novel via human actions increased as the world began to become ever more interconnected (Elton, 1958). As this number grew, the need to understand how this process occurs, and to differentiate natural species’ range expansions from those mediated by humans, became critical. Without making this distinction, it becomes difficult to untangle the mechanisms that are driving historical biodiversity changes, to understand the role of new arrivals in driving evolutionary dynamics and, more practically, to stem the flow of species that cause ecological or economic harm (see below). Before continuing, however, it is very important to recognize that a multitude of names have been given to species that are introduced to a novel location via human actions – such as ‘exotic’, ‘invasive’ or ‘alien’ species (Lockwood et al., 2007). We use the term ‘invasion’ to refer to the process whereby species expand their geographical distribution outside of their natural dispersal range via the actions of humans, while we refer to populations that have become otherwise established outside the bounds of their native ranges as ‘non-native’. A more lucid understanding of the invasion process may be achieved if it is considered as a stepwise progression of events, whereby individuals of some species are moved out of their native ranges, released into a novel location, establish self-sustaining populations there and then spread to new locations (Figure 9.1; Sakai et al., 2001). Fundamental to this process is that not all individuals successfully pass through all these stages. The tens rule of Williamson (1996) states that only ≈10 per cent of transported individuals are released into a foreign location, ≈10 per cent of these introduced species will go on to survive and successfully breed (i.e. establish a new population) and ≈10 per cent of these established species will expand their geographical

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Figure 9.1 Generalized stages common to all species invasions. A species must successfully transition through each sequential stage, and the proportion of species that proceed from one stage to the next is less than the previous one (depicted by arrow width).

ranges and become pests. These estimates were based, in large part, on non-native animals and plants of Britain. More recently, Jeschke and Strayer (2005) investigated all freshwater fish, mammal and bird species native to Europe or North America that have been introduced outside their native range. They found that the frequencies of transitions across all three of the above stages averaged 6.1 per cent, 56.0 per cent and 59.7 per cent, respectively. Regardless of the specific percentages for each stage, it is apparent that only a fraction of the species that are moved by people, either on purpose or by accident, will complete all stages of the invasion process. A considerable amount of research within invasion biology has

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therefore focused on attempts to understand which factors differentiate between those species that successfully progress through all invasion stages and those that do not (Lockwood et al., 2007).

9.1.2 Human-assisted versus prehistoric invasions A valid and persistent question is the extent to which modern trends in species invasions differ from those that occur naturally. This question is especially relevant to students of biogeography because range expansions are a very clear component of palaeoecological and historical biodiversity patterns (Vermeij, 2005). Do modern invasions warrant the attention currently given to them by scientists? How different are the mechanisms, spatial patterns and rates of modern versus prehistoric invasions? Can we use prehistoric trends to help predict the consequences of modern biological invasions? Human-assisted dispersal of non-native species differs from natural dispersal events in several important aspects (J.R.U. Wilson et al., 2009). Ricciardi (2007) detailed the differences between prehistoric and human-assisted invasions, which we summarize below and in Table 9.1. The most obvious differences are in the number and frequency of ‘dispersal’ events. Natural dispersal events are characteristically rare, both in the number of species being transported and in the temporal frequency with which species disperse. By contrast, modern human-assisted dispersal events happen constantly and involve a wide variety of species, which

show a much wider array of biological traits than those species that are likely to experience natural longdistance dispersal. The rate at which non-native populations are establishing around the world is consistently several orders of magnitude larger than fossil-derived estimates for natural dispersal events at the same locations. For example, the invasion rate of terrestrial species for the Hawaiian Islands was approximately 30 species per million years (0.00003 per year) prior to human settlement, but increased to 20,000 species per million years (0.02 per year) after the arrival of the Polynesians and to approximately 20 per year during the past two centuries (Ricciardi, 2007). In other words, contemporary rates of biological invasions are nearly one million times higher than the prehistoric rate for Hawaii before human influence. The number of individuals of each species being transported is also vastly different between natural and human-assisted invasion events. Natural dispersal events typically involve a few individuals of a species finding their way out of the native range and attempting to establish a self-sustaining population in the novel locale. Occasionally the number of individuals in these natural events can be quite high – as for instance, during biotic interchanges involving episodic events of mass dispersal. For example, the opening of the transpolar corridor between the Pacific and Atlantic oceans and the formation of the Panamanian land bridge between North and South America during the Great American Interchange permitted a massive flux of species between formerly isolated regions (Vermeij, 2005; Lomolino et al., 2006). By contrast, humanassisted dispersal events are commonly characterized

Table 9.1 A comparison of key characteristics of prehistoric versus human-assisted invasions. Modified from Table 1 of Ricciardi (2007). Characteristics

Prehistoric invasions

Human-assisted invasions

Frequency of long-distance dispersal event Number of species transported per event Propagule size per event Number of mechanisms and routes of dispersal Temporal and spatial scales of mass transport events Degree of homogenizing effect Potential for interactions with other stressors

Very low Low* Small* Low Episodic (short-distance) Regional Low

Very high High Potentially large High Continuous (long-distance) Global Very high

* Except during biotic interchange events.

Conservation planning in a changing world by the release of hundreds to thousands of individuals of a species into one novel locale, although there is much variation around this number. Finally, human-assisted invasions serve to connect two or more locations that are geographically very distant from one another, whereas natural dispersal events tend to link sites that are comparatively close together or otherwise linked naturally. Quite simply, patterns of modern dispersal unite parts of the world solely by social and economic ties, as opposed to biophysical pathways such as prevailing wind directions, jet streams or ocean currents, as would happen for natural dispersal events (Box 9.1).

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9.1.3 Economic and ecological impacts of invasion The interest in human-assisted invasions has grown rapidly over the past two decades, which is attributable to three factors (Lockwood et al., 2007): • First, as the world economy globalizes, there are increased trade and social connections between geographical localities, and along with these connections come the introduction of non-native species (Perrings et al., 2005; Hulme, 2009). Thus, the sheer number of non-native populations establishing worldwide has increased substantially in recent times.

Box 9.1 The human imprint on modern day species dispersal patterns The Earth is now better connected via human transport than ever before. In recent decades, human activities have greatly increased the frequency and spatial extent of species introductions across the globe through both intentional and unintentional actions. These include ballast-water discharge from international shipping; bait-bucket releases associated with recreational fishing; the global pet trade; intentional translocations of wildlife for recreation purposes; biological control; and inadvertent releases from aquaculture and horticulture activities. The following two case studies illustrate how modern biotas are connected via social and economic networks and by sea and air.

Ship traffic In marine and estuarine systems, the dominant invasion pathway worldwide is the ballast water of commercial ships (Carlton & Geller, 1993; Drake & Lodge, 2004). Ocean-going vessels must achieve proper stability to minimize drag (and thus maximize speed) and to reduce the likelihood of capsizing in rough seas. To achieve this, early ships strategically filled ballast compartments within the hull with soil, rocks or scrap metal – essentially, anything with some weight that could be easily loaded into a ship at dock. Today, ships pump water into ballast tanks, and a typical commercial bulk vessel might carry over 30,000 metric tonnes of ballast water during an inter-oceanic voyage. Ballast water is usually taken from the harbour in one port and subsequently may be discharged in a recipient port through openings in the ship’s hull. The number of non-native species that are transported via ship ballast has increased with the rise in global commerce and the consequent upsurge in the number of ships travelling the world’s oceans and major waterways (Figure B9.1a). Current estimates suggest that a global fleet of approximately 35,000 commercial vessels transports an annual volume of about 3.5 × 109 metric tonnes of ballast water, containing some 7,000–10,000 species (mostly marine) at any one time (Wonham et al., 2005). Even if only a small fraction of these species establish non-native populations, it is easy to see that ballast water is a primary mechanism by which aquatic invasions are occurring. By tracking the number of ships that visit ports worldwide, Drake and Lodge (2004) were able to map ‘hotspots’ of marine invasions and, via network modelling, to determine which ports are likely to have increased rates of invasions in the coming years (Figure B9.1a). These hotspots are clearly the product of economic and social influences on global trade and are in marked contrast to what we might expect given natural dispersal patterns of marine species via oceanic currents.

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Figure B9.1a (a) The frequency of commercial shipping traffic along shipping routes around the world, ranging from low (blue) to high (red). From Halpern et al. (2008). (b) Global hotspots for biological invasion from ballast water, ranging from low (blue) to high (red). From Drake and Lodge (2004). (See Plate B9.1a for a colour version of these images.)

Airline traffic International air travel has been recently pinpointed as a significant factor in the movement of economically damaging pest species and infectious diseases (Tatem, 2009). Among others, the Mediterranean fruit fly Ceratitis capitata has been consistently imported in airline baggage (Liebhold et al., 2006), plant pathogens are often found in air cargo (McCullough et al., 2006) and diseasecarrying mosquitoes have survived long haul flights in aircraft cabins (Lounibos, 2002). Far-removed regions with similar climates have now been suddenly linked by a busy flight schedule, which has resulted in an elevated risk of foreign invasions. This risk, however, depends greatly on the time of year. Tatem and Hay (2007) identified an ‘invasion window’ across the global air network from June to August, when climatic conditions in regions linked by long-haul routes are most similar to one another and the higher number of flights increases the chances of exotic species hitching a ride to somewhere new. With expected increases in global trade and travel (Perrings et al., 2005; Hulme, 2009), opportunities for such extreme hitchhiking through the world airline transportation and shipping network look set to increase further (see trend in Figure B9.1b).

Figure B9.1b Trends in global shipping cargo volumes and air freight, 1970–2005. From Hulme (2009).

Conservation planning in a changing world • Second, as the number of non-native populations increases, scientists find it increasingly hard to ignore them. It is important to recognize that many of these species present unique opportunities to test various ecological, evolutionary and biogeographical concepts and theories. Certainly the basic insights gained from the study of modern invasion events are substantial (Sax et al., 2007). • Third, some of the non-native populations that have established have gone on to impart substantial economic and ecological cost (Simberloff et al., 2005; Pimentel et al., 2006). As detailed above and shown in Figure 9.1, not all species that are dispersed via human actions have negative impacts within their new environment. The definition of what constituents ‘impact’ is somewhat problematic for at least two reasons: 1 There are scientific and societal influences on the perception of impact (not to mention that the effects of invasive species are often subtle and difficult to observe). 2 After impact is perceived, there is a variety of ecological factors that determine the level of impact produced (Lockwood et al., 2007). Let us move past this issue by simply conceding that human perception and valuation are an integral part of the integration stage of the invasion process (Figure 9.1). It is important to recognize that the proportion of species that do cause harm as compared to those that are simply moved out of their native range is quite low. Nevertheless, these few species will eat, parasitize and compete with native species, often driving the latter extinct or into very low population numbers (Elton, 1958; Clavero & García-Berthou, 2005; Strayer et al., 2006). Some non-native populations invade natural areas such as parks or wildlife reserves and disrupt native species communities (Simberloff et al., 2005). In these instances, the value of the natural area in terms of its ability to conserve biodiversity may be reduced if the non-native is not controlled or eradicated. Many species threaten human economic interests, notable examples including the zebra mussels (Dreissena polymorpha) that clog utility companies’ water intake valves (MacIsaac, 1996); emerald ash borers (Agrilus planipennis) that devastate urban and commercial forests (Poland & McCollough, 2006); and monk parakeets (Myiopsitta monachus), whose bulky nests can cause electric power line failures (Avery et al., 2002). A substantial number of non-native species have adverse impacts on human health by transmitting diseases (Lounibos, 2002; Tatem, 2009), the most obvious

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of which is the widespread distribution of Norway rats (Rattus norvegicus). This rodent species was regularly, and inadvertently, transported with human colonists as they expanded across the globe. They serve as the reservoir and vector for a variety of particularly troublesome human diseases, the most well know being bubonic plague. In general, scientists reserve the term ‘invasive’ for these few non-native species that cause ecological or economic harm. It is an open question as to whether these few invasive species have characteristics that make them unique amongst the world’s species, but there is a clear need to be able to identify them as potentially harmful long before they have the chance to become invasive.

9. 2 B I OT I C H OMOGEN I Z AT I ON The regional connectivity of the world is stronger and more varied than ever before and, consequently, there are very few places where non-native species have not become established. Looking back over human history, it is apparent that changes in species diversity are frequently the result of the widespread invasion of ubiquitous non-native species into areas containing rare, and often unique, native species (Elton, 1958, Ricciardi, 2007). If the same non-native species are being introduced to multiple locations, then there is potential for disparate regions to become more similar in their species composition through time, a process known as biotic homogenization. There are certainly well-known invaders that can be found nearly everywhere. These days, for example, you can land at nearly any airport in the world and, while waiting for your next flight, watch house sparrows (Passer domesticus) cavorting on the tarmac. This species is native to Eurasia, but it has realized a very broad geographical distribution via human-mediated introductions. For many years, the biodiversity crisis has been focused on the loss of species through global extinction. Although this is clearly of prime importance, at sub-global scales the loss of populations through local extirpation, combined with the invasion of already common non-native species, may be the more dramatic reconfiguration of modern biodiversity. In fact, changes in diversity patterns at fine and coarse scales of analysis can be either concordant or, alternatively, can be decoupled and even conflicting.

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For example, Pautasso (2007) conducted a metaanalysis of the relationship between human population size and change in the plant and animal species richness of study areas. The study reported negative changes in richness at small spatial scales of analysis (or small extent) but positive changes at larger spatial scales. The introduction of non-native species by humans is typically integral to such changes. In essence, anthropogenic changes driving habitat loss, fragmentation, species invasions and ecosystem transformation may result in declining local richness but, across larger landscapes and regions, relatively few native species may become entirely extinct, while nonnatives boost the richness above natural baseline levels. Changes such as these, in the inventory richness of smaller areas nested within larger regions, may also be accompanied by changing patterns in differentiation diversity, i.e. in the degree of compositional turnover between localities – also known as ‘beta diversity’. A change in beta diversity can, in fact, occur either through a reduction in the total number of species in the region (regional species richness or sometimes ‘epsilon diversity’) or through a change in the species similarity between areas. Basically, if a similar suite of species is shared across the areas in a region, beta diversity will be quite low. If very different species occur in different areas, beta diversity will be high. Biotic homogenization is thus a term describing the process of reducing differentiation diversity between regions, but it may be accompanied by varying patterns of change in inventory richness at different scales of analysis. See Box 1.2 for an explanation of terminology. Put another way, biotic homogenization is described as the process by which regionally distinct native communities are gradually replaced by locally expanding, cosmopolitan, non-native communities (McKinney & Lockwood, 1999). Some have likened the process of biotic homogenization to the now global distribution of fast-food restaurants, coffee houses and big-box retailers (Olden et al., 2005). The more connected we are as a society, the more likely we are to see the trans-global distribution of both species and businesses. In circumstances where invasive species impact negatively on locally co-occurring native species, rare and endemic native species may be lost, resulting in rapid loss of differentiation diversity. However, it is also important to recognize that the reverse can also occur and that, in cases, the combined effects of invasions and extirpations can be to increase the mean differentiation

diversity across a study region, a phenomenon termed ‘biotic differentiation’ by Olden and Poff (2003).

9.2.1 The process of biotic homogenization In the simplest sense, human activities that increase rates of species invasions and extirpations are the ultimate cause of biotic homogenization. However, biotic homogenization can arise when only invasions occur without the concurrent loss of species, or conversely where only species extirpations occur. In other words, species additions or replacements need not occur for regions to become homogenized or even differentiated over time (Olden & Poff, 2003). To illustrate this point, we provide a simple graphical example showing how the number and manner in which non-native species establishment and native species extirpations occur may lead to very different levels of homogenization or differentiation (Figure 9.2). In the absence of any extirpation, the establishment of the same non-native species at two separate localities will lead to increases in the similarity of the invaded communities. Conversely, the establishment of a different non-native species at each locality will decrease community similarity. Although this example is useful to illustrate the simplest way biotic homogenization can occur, both empirical data and theoretical modelling suggests that the process is both complex and sensitive to the spatial and temporal scale of investigation (Olden, 2006).

9.2.2 Different manifestations of biotic homogenization Biotic homogenization is considered an overarching process that encompasses either the loss of taxonomic, genetic or functional distinctiveness over time (Olden et al., 2004). Taxonomic homogenization, which we used to introduce the concept of homogenization above, has been the primary focus of previous research and is commonly referred to as biotic homogenization. However, imposing a narrow definition of biotic homogenization does not truly reflect the multidimensional nature of this process. Consequently, it is useful to think of biotic homogenization as a broader ecological process by which formerly disparate biotas lose biological distinctiveness at any level of organization, including in their genetic and functional characteristics.

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Figure 9.2 Illustration of how species invasions and extinctions can cause either biotic (taxonomic) homogenization in scenario 1 or differentiation in scenario 2, depending on the identity of the species involved. A pair of communities (shaded ovals) for each scenario is illustrated, where extirpation events are represented by the disappearance of a species icon over a time step, whereas introduction events are represented by the arrow and appearance of a species icon. Importantly, both scenarios share the same species pool (6 native butterflies, 2 introduced butterflies) and species richness through time is identical for both scenarios. From Olden and Rooney (2006).

Let us spend a moment exploring these two additional ways in which biotic homogenization can be manifested. Genetic homogenization refers to a reduction in genetic variability within a species or among populations of a species. It can occur through at least three mechanisms:

• First, the intentional translocation of populations from one part of the range to another enhances the potential for intraspecific hybridization (i.e. hybridization between different sub-species within a species), with the end result being the assimilation of gene pools that were previously differentiated in space (Stockwell et al., 1996).

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• Second, introductions of species outside of their original range(s) increases the likelihood of a founder effect and reduced levels of genetic variability, as well as setting the stage for interspecific hybridization (i.e. hybridization between different species within the same genus) (Rhymer & Simberloff, 1996). • Third, if extirpations were a cause for faunal homogenization, then one consequence might be bottleneck(s) in local populations of the impacted species, along with lowered effective population size(s) (Lee, 2002). Functional homogenization refers to an increase in the functional similarity of biotas over time resulting from the replacement of ecological specialists by the same widespread generalists. It occurs primarily because patterns of species invasions and extirpations are not random, but instead are related to particular biological traits that commonly predispose native species to extirpation and non-native species to successful establishment. The end result is an increase in the functional convergence of biotas over time associated with the establishment of species with similar ‘roles’ in the ecosystem (e.g. high redundancy of functional forms or traits) and the loss of species possessing unique functional ‘roles’ (McKinney & Lockwood, 1999; Olden et al., 2004). For example, Winter et al. (2008) examined how the presence of non-native plant species in Germany affected the distribution of a genetic trait, namely ploidy level (referring to the number of homologous sets of chromosomes in a biological cell), at two spatial scales. It is commonly accepted that polyploidy species should have a greater ability to colonize or invade new habitats due to greater genetic variability. Interestingly, this study found evidence for functional differentiation at fine spatial scales ( 0.60) panels. Jtotal refers to floral similarity based on native and non-native species composition. Floral similarity is based on Jaccard’s coefficient of similarity (J), which ranges from 0 (no species in common) to 1 (all species in common). From Figure 1 of Qian and Ricklefs (2006).

41 nations located worldwide. They found that between 1965 and 2005, ungulate assemblages had become two per cent more similar for countries globally and eight per cent more similar at the coarsest resolution within South Africa. Interestingly, species introduced from other continents, as opposed to those introduced from within Africa, were found to have different effects on patterns of homogenization. Homogenization was most affected by translocations of species from neighbouring localities (extra-limital species) (4.6 per cent increase in similarity), whereas introductions of ungulates from more distant areas (extra-regional species) tended to differentiate assemblages (3.8 per cent decreased in similarity). Quite simply, non-native species introduced from distant regions are more likely to establish in only a few localities, resulting in differentiation. Similar findings have also been reported for plants and freshwater fishes in the United States (McKinney, 2005; LaSorte & McKinney, 2006). Levels of homogenization were found to increase with increasing resolution (see Table 9.2) and with time. In the South African study, from 1971 to 2005, homogenization by extra-limital introductions increased rapidly after

initially having a smaller homogenizing effect than the differentiating effect of extra-regional introductions (Figure 9.6).

9. 4 EN V I R ON MEN T AL AN D H U MAN DR I V ER S OF B I OT I C H OMOGEN I Z AT I ON Environmental change ultimately promotes the geographical expansion of some species and the geographical reduction of others, leading to biotic homogenization (McKinney & Lockwood, 1999). Habitat loss, pollution, climate change or other sources of disturbance often precede, and in a sense prepare, the environment for changes in beta diversity over time. The research highlighted above, in addition to a number of other studies in the literature, has provided compelling evidence linking human-induced environmental change to biotic homogenization across taxonomic groups. Collectively, this research has shown that human activities on the landscape are often characterized by greater increases in taxonomic similarity, suggesting that

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Figure 9.6 Temporal trends in ungulate homogenization as a result of extra-regional and extra-limital introductions in South Africa, at the quarter-degree grid cell resolution, between 1971 and 2005. Redrawn from Figure 4 of Spear and Chown (2008).

humans are playing a central role in promoting the homogenization process by introducing new species and favouring the persistence of non-native species over native species. For freshwater ecosystems, Scott and Helfman (2001) reported that cosmopolitan species’ richness increased and endemic species’ richness decreased in response to increased watershed deforestation and density of buildings and roads in Tennessee, USA. At a larger spatial scale, Marchetti et al. (2001) observed that measures of human occupancy and aquatic habitat alteration, including the density of dams and aqueducts in the watershed, were associated with increased similarity of zoogeographical provinces in fish communities in California, USA. However, at a finer spatial scale, Marchetti et al. (2006) found a negative relationship between change in community similarity and the proportion of the watershed in development (including commercial, industrial, urban and suburban) – or, in other words, more developed watersheds showed greater biotic differentiation. Olden et al. (2008) found that geographical patterns of homogenization in Australia were highly concordant with levels of disturbance associated with human settlement, infrastructure and land use. These results

suggest that human settlement may directly increase the likelihood of intentional or accidental non-native species introductions, and disturbance associated with physical infrastructure and land-use change may promote the establishment of these species by disrupting environmental conditions. Wetland degradation has also led to the homogenization of aquatic and invertebrate communities in Michigan, USA (Lougheed et al., 2008). Specifically, habitat homogenization at both the local and landscape scales were found to shift community structure from a species-rich and spatially heterogeneous community dominated by floating-leaved plants in undeveloped wetlands, to nutrient-rich wetlands dominated by ubiquitous duckweed (Lemnaceae). Urban/rural gradient studies have provided important insights into associations between urbanization and bird and plant homogenization. Blair (2004) found that temporal changes in bird community composition varied in a similar fashion along an urban/rural gradient in the oak woodlands of northern California and the eastern broadleaf forests of Ohio, USA. The degree of taxonomic overlap in the bird communities increased from approximately 5 per cent in the least developed sites to approximately 20 per cent in the

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most urbanized sites – an outcome of the replacement of local endemic species (often urban-sensitive species) by ubiquitous non-native species (urban-adapted species). By contrast, Clergeau et al. (2006) found that avifaunal similarity of town centres in Europe was actually lower than in less urbanized habitats – a result that may have been connected to the larger size of towns and, thus, greater types of potential habitat in this study system. The results from this study also suggested that urbanization might cause homogenization by decreasing the abundance of ground-nesting bird species and bird species that preferred bush/shrub habitats. Schwartz et al. (2006) reported floristic homogenization of urbanized counties in southern California, whereas they found no change in more rural areas of northern California. The study of Kühn and Klotz (2006), on the other hand, found no overall relationship between patterns of homogenization and urbanization across Germany. In summary, although urbanization undoubtedly plays a role in shaping patterns of biotic homogenization, the exact nature and generality of this relationship is still unclear (McKinney, 2006). Environmentally mediated interactions between species may also be an important driver of biotic homogenization. Holway and Suarez (2006) examined native ant communities in scrub and riparian habitats of mediterranean California to test the hypothesis that the invasion of Argentine ant (Linepithema humile) has caused biotic homogenization. By comparing invaded and un-invaded sites across similar habitats, the authors showed that sites invaded by Argentine ants have lower beta diversity compared to un-invaded sites. Specifically, functional homogenization of ant communities occurred via shifting community dominance to smaller-bodied workers with lower thermal tolerance and a reduced diversity of behaviours (i.e. nesting habits, dispersal strategies and foraging behaviours). Because Argentine ant abundance in seasonally-dry mediterranean environments is positively correlated with soil moisture, the authors hypothesized that the homogenizing effects of the Argentine ant are facilitated by inputs of urban and agricultural water run-off that acts to create mesic soil conditions. This observation supports the notion that anthropogenic modifications to the environment indirectly cause biotic homogenization by creating opportunities for the invasion of the Argentine ant, as opposed to threatening the persistence of native ants directly.

9. 5 B I OT I C H OMOGEN I Z AT I ON AN D CON S ER V AT I ON Biotic homogenization is an important dimension of the modern biodiversity crisis, with significant ecological, evolutionary and social implications (McKinney & Lockwood, 1999; Olden et al., 2004; 2005). It extends beyond the narrow focus on elevated extinction rates to incorporate the other side of the equation: the establishment of non-native species. Biotic homogenization conjures the prospect of Kunstler’s (1993) The Geography of Nowhere, in which biotic distinctiveness is gradually dissolving over time. Consequently, a major challenge within conservation biogeography is to identify and understand present-day patterns of biotic homogenization to guide policy aimed at mitigating its future effects (Rooney et al., 2007). Clearly, the most effective conservation of biodiversity involves reducing and, where possible, preventing the two processes generating biotic homogenization – species invasions and extinctions. The conundrum is determining the best way to achieve this goal. Because the key factors facilitating homogenization include people and habitat transformation (through extinctions or the establishment of non-native species), a first step towards achieving biodiversity conservation goals is to focus efforts in areas subject to human activities and to reduce human-related impacts. Unfortunately, there is a strong correlation between human population density and species richness, and the areas of high biotic diversity that are under the greatest threat are often in the most populated areas (Chown et al., 2003; McDonald et al., 2008). Indeed, at a finer scale of analysis, designated conservation areas may often attract people to them through perceived benefits of employment, market access and foreign aid (Wittmeyer et al. 2008). The increased external threat from accelerated human population growth does not bode well for the native biota in these areas, which consequently face the risk of increased homogenization. In the past, purposeful homogenization was undertaken within countries such as Australia and some Pacific island territories by acclimatization societies within colonist human societies who, for a variety of reasons, wanted to surround themselves with familiar, colourful or (regarding birds) tuneful species. Even today, some conservation organizations encourage the intentional movement or translocation of species, which may also have the unintended consequence of promoting homogenization.

Conservation planning in a changing world This act is a problem when species are introduced and become established outside of their historical distribution, or where the genetic consequences (e.g. interspecific hybridization) are not considered. For example, in parks across southern Africa there has been a trend to introduce the same suite of species across nature reserves. Fuelled by tourism and the public’s desire to see large mammals (especially predators), spotted hyena (Crocuta crocuta), wild dog (Lycaon pictus) and antelope such as roan (Hippotragus equinus) have been introduced and have established within areas where they did not historically occur, or to areas that are now unsuitable due to small park sizes. In fact, Spear and Chown (2008) demonstrated that it is extra-limital introductions that are driving the homogenization of ungulate assemblages in South Africa (Figure 9.6). They warn that the potential for changes in local diversity and ecosystem functioning as a consequence of translocations should not be underestimated. These concerns contrast with other conservation actors arguing for various forms of rewilding, or for assisted migrations of species as a climate-change mitigation strategy (see, e.g. Chapter 3; Donlan, 2007). The concept of biotic homogenization and differentiation may provide a useful tool in conservation planning (Rooney et al., 2007). Much attention in conservation has focused on reserve selection and choosing the best network of reserves to maximize biodiversity coverage. Such efforts have largely focused on species number, endemism and complementarity as the metrics that should be optimized (Chapters 6 and 7; Pressey et al., 1993). Complementarity exists when an area has some biodiversity components that are unrepresented in other areas. It may thus be possible to use biotic homogenization to monitor whether complementarity goals are being met. For example, if a network of reserves becomes more similar over time due to the loss of unique species, this reduces complementarity (Rooney et al., 2007). Importantly, any assessment of complementarity related to conservation planning should be restricted to indigenous species only. The inclusion of non-native species could show increased biotic homogenization when, in reality, the full set of native species that the reserve network was designed to conserve still occur. This idea has much potential, but there are a few caveats. For example, when dealing with a minimum set complementarity (each area contains distinctive species) goal, all areas may lose the same number of

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unique species over time, and neither complementarity nor the level of biotic homogenization would change, yet the true state of biodiversity loss will not be reflected.

9. 6 N OV EL AS S EMB LAGES Novel assemblages, sometimes referred to as novel or emerging ecosystems, are communities that consist of extant species which have not occurred previously in the same combinations found today (Hobbs et al., 2006). Increased homogenization of biotas associated with the massive and accelerating movement of species within and between regions/provinces is likely to contribute substantially to the creation of novel or noanalogue assemblages. Although, technically, any area that has lost native species or gained non-native species is novel in some respect, some current assemblages have been transformed to such an extent that they are verging on becoming entirely new assemblages (Williams & Jackson, 2007). Certainly, in terms of system functioning, many ecosystems have already become ‘novel’. One of the best examples comes from the San Francisco Bay, California, which has the dubious distinction of being the most invaded aquatic region on Earth, with more than half its fish and most of its bottom-dwelling organisms representing non-native species (Cohen & Carlton, 1998). The total dominance (number of species and biomass) of non-native species has transformed the bay from a pelagic (mid-water) system to a benthic (bottom) one and productivity has declined. Invasive species such as Corbula amurensis (Asian clam), Sphaeroma quoyanum (a burrowing isopod from Australia and New Zealand) and Spartina alterniflora (smooth cordgrass) have become among the most important species in the bay in terms of both biomass and their role in controlling biological processes in the bay (Cohen & Carlton, 1998). Although the process of homogenization can create novel assemblages, global climate change is increasingly likely to magnify this effect. Thus, any prediction of where novel assemblages will form needs to take into account not only non-native species introductions, but also global climate change and the individualistic responses of species (native and non-native) to environmental change (Chapters 4, 7). Recent models suggest there will be substantial regions of the world with novel climates by 2100 (particularly in tropical and sub-tropical regions) and also that some extant

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Figure 9.7 A conceptual diagram showing how nonanalogue combinations of species arise in response to novel climates. The set of climates in existence at two periods are represented as open ellipses. Novel climates are the portions of the 21st century envelope that do not overlap 20th century climates, and disappearing climates are the portions of the 20th century envelope that do not overlap 21st century climates. Species co-occur only if their fundamental niches simultaneously intersect with each other and the current climatic space. Future climate change may cause a variety of ecological responses, including shifts in species’ distributions (species 1–3), community disaggregation (species 1 and 3), new communities forming (species 2 and 3), and extinction (species 4). From Figure 1 of Williams et al. (2007); copyright (2007) National Academy of Sciences, USA.

create novel environmental conditions or, as Saxon et al. (2005) refers to them, ‘environmental domains’. The disappearance or contraction of present environmental domains and the appearance of new domains will have profound consequences for most species and the identity of communities today. Climate change is expected to alter the effectiveness of environmental filters; to alter the likelihood of species establishing; to change pathways of species introductions; and to affect the impact of non-native species (Rahel & Olden, 2008). The combination of novel assemblages and altered biophysical conditions will result in new systems that have unknown functional characteristics, and whose processes and interactions are hard to predict (Hobbs et al., 2006). Given the dynamic nature of species’ distributions, current homogenization patterns and trends are likely to change too. It is very difficult to predict the make-up of novel assemblages, given that it is almost impossible to know which species will co-occur, whether they will interact and how altered climatic regimes will influence any interaction. Importantly, many of these communities may be more, or less, similar across locations than the native assemblages they replaced. In other words, homogenization is not the only outcome of the massive movement of species across the globe. Perhaps the only certainty is that conservation efforts will have to intensify to tackle the threat of anthropogenically-assisted novel assemblages, and society will be faced with some tough decisions as to what biodiversity it values.

FOR DI S CU S S I ON climate types will have disappeared (Williams et al., 2007). Because climate is a primary control on species’ distributions and ecosystem processes, novel 21st century climates may promote the formation of novel species associations and other ecological surprises. On the other hand, the disappearance of some extant climates increases the risk of extinction for species with narrow geographical or climatic distributions, as well as the risk of disruption of existing communities (Figure 9.7). Of greater concern, perhaps, is the combined effect of altered climate and other abiotic environmental characteristics (such as topography or soil type) which

1 How do natural patterns of species invasion differ from anthropogenically assisted species invasions, and with what consequences? 2 In the light of social demands and economic development, what are the most likely timescales and scenarios of introduction, establishment and spread of non-native species in the future? 3 What are the ecological consequences of faunal and floral homogenization? 4 What are the temporal dynamics of taxonomic and functional homogenization? 5 What are the primary environmental and biological drivers of biotic homogenization at different spatial and temporal scales?

Conservation planning in a changing world 6 How will rates and patterns of biotic homogenization respond to shifting pathways of species introductions and future environmental change? 7 What novel species assemblages are likely to emerge in response to climate change? 8 What might be the consequences of novel ecosystems for biodiversity, ecosystem functioning, and human societies?

S U G G ES T E D R E AD I NG Elton, C.S. (1958) The ecology of invasions by animals and plants. Methuen, London. Hobbs, R.J., Arico, S., Aronson, J., Baron, J.S., Bridgewater, P., Cramer, V.A., Epstein, P.R., Ewel, J.J., Klink, C.A., Lugo, A.E., Norton, D., Ojima, D., Richardson, D.M., Sanderson, E.W., Valladares, F., Vilà, M., Zamora, R., & Zobel, M. (2006) Novel ecosystems: theoretical and management aspects

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of the new ecological world order. Global Ecology and Biogeography, 15, 1–7. McKinney, M.L. & Lockwood, J.L. (1999) Biotic homogenization: a few winners replacing many losers in the next mass extinction. Trends in Ecology & Evolution, 14, 450–453. Olden, J.D. (2006) Biotic homogenization: a new research agenda for conservation biogeography. Journal of Biogeography, 33, 2027–2039. Rahel, F.J. (2002) Homogenization of freshwater faunas. Annual Review of Ecology and Systematics, 33, 291–315. Riccardi, A. (2007) Are modern biological invasions an unprecedented form of global change? Conservation Biology, 21, 329–336. Sax, D.F., Stachowicz, J.J., Brown, J.H., Bruno, J.F., Dawson, M.N., Gaines, S.D., Grosberg, R.K., Hastings, A., Holt, R.D., Mayfield, M.M., O’Connor, M.I., & Rice, W.R. (2007) Ecological and evolutionary insights from species invasions. Trends in Ecology & Evolution, 22, 465–471. Strayer, D.L., Eviner, V.T., Jeschke, J.M., & Pace, M.L. (2006) Understanding the long-term effects of species invasions. Trends in Ecology & Evolution, 21, 645–651.

PART 4 FUTURE DIRECTIONS

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

CHAPTER 10 Prospects and Challenges Richard J. Ladle1,2 and Robert J. Whittaker1 1

School of Geography and the Environment, University of Oxford, Oxford, UK Department of Agricultural Engineering, Federal University of Viçosa, Brazil

2

10. 1 WHY W E NE E D C ONS E R V A T I ON BI O G EO GR AP HY The death of a species is a more remarkable event than the end of an imperial dynasty. (Orton 1869, p. 540, commenting on the demise of the Great Auk.) The world is in the midst of an unparalleled period of biotic change driven by human alterations of the natural environment. Even with the considerable gaps in our basic knowledge of global biodiversity, there is still good evidence for an increase in species extinctions relative to natural background rates (Lawton & May, 1995). This human-induced crisis began towards the end of the Pleistocene as modern humans began to spread out from Africa, bringing their tool-using and ecosystem-transforming habits to other land masses (Wilson, 1992; Lomolino et al., 2010). Notwithstanding that there remains huge uncertainty surrounding the magnitude and significance of current patterns of species extinction (Ladle, 2009), much of this uncertainty relates to poor knowledge of baseline diversity levels, rather than to whether species extinction rates have been elevated by humans. For example, in 2000, at a discussion hosted by the National Academy of Sciences of the United States of America on the ‘Future of Evolution’ the expert panel unanimously agreed that current extinction rates are 50–500 times background and are still increasing (Woodruff, 2001).

Additional to the recent historical patterns of elevated rates of extinction driven by habitat conversion, hunting and biotic homogenization is the spectre of rapid anthropogenic climate change, which has the potential to cause dramatic shifts in the distributions of species and ecosystems before the end of the current century (Chapter 7). Each of these processes is difficult enough to model on its own, but there are, of course, interactions and synergies (multiplying effects) between them. For example, as habitats shift and transform (and sometimes disappear completely), exotic species will shift in their distributions alongside native species, with both new arrivals and old members of the regional biota invading new territories, forming up assemblages and communities with no modern or past analogues (Chapters 3, 7 and 9). Conservation biogeography will not be a source of cure-alls, but it can provide the tools and concepts that are needed to make scientifically informed choices about what and where to protect, the consequences of different policies, or of not acting at all (Table 10.1). This is essential, because effective biodiversity conservation requires that governments and other policy-making bodies make rational decisions about land use (or ocean use) and management that are based on the most accurate and up-to-date information. Biogeographers also have an important role to play by contributing to the debate about the ultimate goals of conservation through education and public engagement (Ladle, 2008; Devictor et al., 2010). As

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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Prospects and challenges

Table 10.1 Prominent areas of current research in conservation biogeography as identified by Richardson & Whittaker (2010; their Table 1). The biogeography of degradation (habitat fragmentation, homogenization, urbanization and other humaninduced impacts) Processes (colonization, climate as a fundamental determinant of distribution, dispersal, disturbance, extinction, persistence, range expansion, resilience, speciation) Inventory, mapping and data issues (atlas data, breeding bird surveys, citizen science, detectability/ discovery probabilities, herbaria and other collections, sampling intensity and biases) Species distribution modelling (bioclimatic modelling, habitat suitability analysis, model performance, niche-based models, presence-only data vs. presence–absence data, dispersal kernel analysis) Characterizing biotas (conservation status, diversity indices and patterns, ecoregions, endemism, rarity, range size, species–area relationships, threatened species, identification of alternative baselines from longterm ecological data) Conservation planning (complementarity, congruence, conservation units, ecosystem services, gap analysis, global conservation assessments, irreplaceability, reserve networks, surrogates) Methods (molecular methods, palaeoecology, remote sensing, scenario development) Related fields (global change biology, invasion ecology, bioinformatics, molecular phylogenetics, network analysis, reintroduction ecology, risk analysis, behavioural ecology, population viability analysis) Overarching themes: niche (fundamental vs. realized), novel climates/ecosystems, scale issues, uncertainty, Linnean shortfall, Wallacean shortfall)

new biogeographical knowledge enters the public domain on issues such as extinction rates, range shifts, species invasions and habitat transformation, public attitudes may shift, putting pressure on policymakers and presenting both challenges and opportunities to conservation organizations. However, it is important to remember that the relationship between knowledge, attitudes and behaviours is complex and multi-faceted, and that the act of providing new information may not, in itself, be sufficient to influence attitudes and actions. Explaining, contextualizing and framing biogeographical information for diverse constituencies is therefore arguably as important as generating the data.

1 Barriers to scientific development: • Filling biogeographical shortfalls • Improving the accuracy and specificity of forecasts 2 Barriers to application: • Turning theory into practice • Education and communication • Social values and lifestyles In the following sub-sections, we will address each of these challenges, the prospects for overcoming them and how doing so may impact on the discipline and contribute to the long-term conservation of biodiversity.

10. 2 T HE C HAL L E NGE S

10.2.1 Filling the Wallacean and Linnean shortfalls

Conservation biogeographers face a range of different challenges over the next few decades if they are to make a significant and lasting impact on the conservation of global diversity (Richardson & Whittaker, 2010). The five challenges identified below reflect two different categories of barriers to the development of the discipline:

As we have repeatedly stressed within this book, data on species identities and distributions are central to research and practice in conservation biogeography. It is therefore essential that these data are of appropriate quality, i.e. that they are fit for purpose. For example, global conservation prioritization frameworks (Chapter 5) such as Conservation

Future directions International’s hotspots scheme require data on both richness and geographical distribution (to assess endemism) but, being a coarse global analysis, the data required do not need to be of the same resolution as needed for within-country protected area planning frameworks. Even so, the CI hotspots analysis is premised on the assumptions that the plant species richness of large geographical areas can be deemed reasonably complete and that the distributions of species is generally known well enough to judge the endemism of each area. In addition, the scheme requires baseline assessments of ‘original’ and present natural vegetation cover – data that are typically crudely specified at regional scales, and which often become increasingly poorly specified at finer scales of analysis (Chapter 3). These problems are general to strategic conservation planning; the application of selection algorithms and other methods for the optimal design of representative protected area networks also require detailed information on species numbers, identities and geographical distributions (Chapters 4 to 6). Moreover, the accuracy of forecasts about the future distributions (and possible extinction) of species under climate change or any other sort of environmental change are also critically constrained by the quality of their input data (Whittaker et al., 2005; Ladle, 2009). Given the central role of robust biogeographical data in the development of conservation biogeography, efforts to address the Wallacean and Linnean shortfalls are likely to be of continuing importance for a long time to come, although there are several important initiatives in play that hold the promise for rapid advances in the coming years and decades (Chapter 4). One such area is the production of a definitive global species list that can be used to resolve problems such as synonomy. The ‘Catalogue of Life’ (CoL: www. catalogueoflife.org), which aims to become a comprehensive catalogue of all known species of organisms on Earth, now has 1.1 million species on its annual checklist (Thomas, 2009). Species occurrence records are also rapidly accumulating, most notably through the Global Biodiversity Information Facility (GBIF: www. gbif.org), which provides access to 189 million species occurrence records to date. More ambitious and data rich bioinformatics projects are also under way, such as the much vaunted ‘Encyclopedia of Life’ (www. eol.org) project (initiated in 2007) detailed in Chapter 4. Most recently, some conservationists have suggested that the IUCN Red List system needs to be expanded into a project dubbed the ‘Barometer of Life’, which

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focuses on the conservation threat status of each species (Stuart et al., 2010). These global initiatives will, at a conservative estimate, take at least a decade to complete (Thomas, 2009), although this may be either overly pessimistic or optimistic, depending on technological advances and socio-economic trends. The hopeful view is that advances in automated species identification and remote sensing hardware will rapidly accelerate the current rate of data acquisition. A less rose-tinted assessment is that limited funding and the shifting geopolitical climate may critically impede the progress of many initiatives. Although global bioinformatics initiatives such as the GBIF provide important data for research and practice, in general they lack specific tools and applications to assist real-world decisions about the conservation, management and the sustainable use of biodiversity (see also Section 10.2.3 below). Biodiversity information is also increasingly being seen as important for decision-makers in other sectors such as agriculture, fisheries, and tourism. Several countries, realizing the importance of providing high quality biodiversity data to their decisionmakers in a form that is accessible and useful, have responded by creating more spatially focused regional or national biodiversity information management initiatives. These typically tend to concentrate more upon providing the information and tools required for policy, governance and management across different sectors. For example, the recently implemented InterAmerican Biodiversity Information Network (IABIN: www.iabin.net) project aims to: 1 develop an internet-based decentralized network to provide access to scientifically credible biodiversity information that currently exists in individual institutions and agencies in the Americas; 2 provide the tools necessary to draw knowledge from that wealth of resources, which in turn will support sound decision-making concerning the conservation and sustainable use of biodiversity. The eventual impact of these various initiatives on the sum total of global biogeographical knowledge should be immense. By processing and collating the data that already exist in scattered and largely inaccessible form, they pave the way to better forecasts and more informed environmental decision-making about conservation, natural resource management, agriculture, sustainable development, etc. There may also be

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enormous benefits for our conceptual understanding of biogeographical processes and relationships. However, in the excitement generated by the online availability of data sets that, until recently, would take years of study to assemble, it is important not to lose sight of the need for continued efforts to determine taxonomies and phylogenies, to document community ecology and auto-ecology, and to engage in continued field sampling to generate the rigorous, systematic and unbiased data sets required for pure and applied biogeographical analysis. Indeed, one problem inherent in the increasing availability of digital data sets is that it becomes increasingly easy for scientists to undertake sophisticated but poorly founded analyses of inadequately sampled systems with which they are insufficiently familiar, as they have not personally taken years of painstaking efforts to assemble the data. To guard against this prospect, it is important that those building such databases make particular efforts to record meta-data describing the properties of the data sets, and that biogeographers undertake what we may broadly term sensitivity analyses involving the use of null or simulation models, analyses of the spatial structure of their data, application of data quality filtering, etc., to tease out the real structure in their data from the bias and noise that frequently exist (e.g. see Hopkins 2007; Hortal et al., 2007; Feeley & Silman, 2010). Moreover, high-quality biodiversity information systems and biogeographical analyses are not, in themselves, the solution to global conservation problems. The hard decisions about where to invest resources and what aspects of biodiversity and ecosystem function should be prioritized will still be dependent upon a mixture of historical legacy, political necessity and societal consent (see Chapters 2 and 3).

10.2.2 Improving models, simulations and forecasts We previously highlighted four generic themes which, we argued, required concerted attention within conservation biogeography, these were: 1 scale dependency; 2 inadequacies in taxonomic and distributional data; 3 sensitivity analyses to develop improved understanding of the effects of model structure and parameterization (e.g. in relation to predicting future species extinction rates); and

4 development of more refined theoretical models, e.g. of species–area effects (Whittaker et al., 2005; Richardson & Whittaker, 2010). As discussed in these pages, much recent progress has been made in these areas, but they are likely to remain important research foci. We will briefly discuss each in turn. First, scale dependency is of central importance in diverse aspects of conservation biogeography. For instance, our assessments of conservation baselines are strongly influenced by the temporal frame of reference (Chapter 3), yet there remain many areas of the world where we lack clearly resolved long-term perspectives on ecosystem dynamics (e.g. Birks, 2005; Willis et al., 2007). Similarly, assessment of the spatial patterns of diversity, the location of hotspots and the outcomes of strategic conservation planning exercises have also been shown to exhibit scale-dependency (e.g. Lennon et al., 2001; Araújo et al., 2005a). The effects of anthropogenic influences generally, changes arising from the introduction of non-native species (e.g. Olden, 2006; Foxcroft et al., 2009) and the criteria applied to assess the extinction risk assigned to plant or animal species also are sensitive to scale parameters of the system (e.g. Martín, 2009). Second (as highlighted in Section 10.2.1), as more and more genetic, taxonomic and distributional data are becoming available for analysis, we need to develop increasingly sophisticated means of determining which components of diversity variation are artefacts of collecting intensity or of analytical failings, as opposed to the ‘real’ underlying biogeographical pattern. In addition, we are in the midst of a phylogenetic revolution, fuelled by fast, cheap molecular DNAsequencing technologies, which promises not only to continue to refine our knowledge of species identities but also to continue to provide exciting advances in our understanding of evolutionary relationships across even large clades. This in turn opens up new possibilities and challenges in the use of evolutionary distinctiveness indices in conservation planning analyses (e.g. Forest et al., 2007; Cadotte & Davies, 2010). Third, and a key repeated theme, is the need to deploy sensitivity analyses to a wide range of issues in modelling future processes and patterns of diversity change – for example, in respect of the spread of alien species (Gallien et al., 2010; Smolik et al., 2010); the role of spatial autocorrelation in species’ distribution models (Veloz, 2009); and the forecasting of species’

Future directions responses to future climate and land-use change (Araújo et al., 2008; Diniz-Filho et al., 2009). Fourth, there is a tendency for many of us to seek to draw recommendations from particular case studies as we publish them but, as in many areas of human endeavour, a single case study may make for poor guidance. Therefore we highlight a continuing need for efforts to develop and synthesize emerging findings into improved theoretical frameworks and to update conservation biogeography theory for the purpose of revising guidelines to practitioners.

10.2.3 Turning theory into practice Biogeographical science has had a foundational role within conservation biology, as demonstrated by the early attention paid to deriving guidelines for protected area design from island biogeography (Chapter 8). Currently, biogeography is in an exciting phase in which there is greatly enhanced potential to apply the subject to problems in the conservation of biodiversity (Whittaker et al., 2005; Richardson & Whittaker, 2010). An important step in this process is to encourage the teaching of conservation biogeography at university level – a key motivation in the writing of this book. However, even if there is a rapid increase in biogeographical understanding, translating this knowledge into guidelines, protocols, tools and applications that are useful at every level of conservation decisionmaking presents significant challenges. Indeed, many would argue that generating the scientific information is the easy bit, and the difficulties really start in communicating geographically precise information about the current and future status of biodiversity in a form that is accessible and able to influence policy (Kalliola et al., 2008). If societies are to realize the full benefits of conceptual advances and an increase in knowledge, it is therefore vital that new understandings are quickly converted into practical tools. This process can be facilitated in a number of ways: • First, conservation biogeography, like many other ecologically-related disciplines, will progress faster and have more impact if scientists and practitioners make their data freely available and accessible. This will be a challenge. There remains a culture of ‘data hoarding’ in ecological and environmental science

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(Parr & Cummings, 2005; Thomas, 2009). This is understandable from the perspective of those who wish to make further use (i.e. produce more publications) of hard-won, painstakingly compiled data sets prior to releasing them for others to plunder. Costello (2009) has recently argued that many of the perceived barriers will be removed if the idea of ‘data sharing’ is replaced by that of ‘data publication’. However, to make this a reality, online data publication systems and, of more relevance to conservation biogeography, biodiversity information systems (see Section 10.2.1 above) will need to develop mechanisms and protocols for data curation, supporting documentation, citation and data access. • Second, a greater investment needs to be made in producing freely available and user-friendly tools and applications to analyse and visualize biogeographical data. An excellent example of this is the SAM (Spatial Analysis in Macroecology) software that allows biogeographers to correct statistically for the influence of spatial clustering of data points (Rangel et al., 2006). Interestingly, the often powerful and pervasive influence of this sort of clustering (known technically as spatial auto-correlation) on standard statistical tests has only recently been generally recognized (Legendre, 1993), and many studies still ignore it in their analyses. It is therefore particularly encouraging that Rangel and his colleagues have developed and made this powerful statistical package freely available (download from: www.ecoevol.ufg.br/sam). Other similar initiatives are clearly to be welcomed. • Third, greater efforts need to be made to mainstream biodiversity into other sectors of environmental governance. In this context, ‘mainstreaming’ can be defined as the integration of conservation goals and sustainable use of biodiversity into sectors that impact biodiversity (mainly outside of protected areas). Successful mainstreaming of biodiversity requires that high-quality biodiversity data are made available to key decision-makers in forms that they can use. Conservation biogeography can play a vital role in this by providing improved frameworks and concepts that allow biodiversity information to be meaningfully organized and structured. Moreover, conservation biogeographers have the knowledge and skills to produce models and visualization tools that can transform raw data on occurrences and distributions into products that are useful and of genuine practical importance.

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Advances are already being made in this direction; for example, Richardson et al. (2009) recently developed a multidimensional decision-making framework for managed relocation of species as a response to anthropogenic climate change – one of the most radical and potentially socially divisive conservation strategies currently being considered. Their heuristic tool incorporates both ecological and social criteria and, critically, produces outputs that can be depicted in graphical 2-D space, making it easier for diverse stakeholders to interpret and use. In the broader context, effective mainstreaming requires that environmental stakeholders and wider society are made aware of the importance of biodiversity and its current status. This can be achieved through better education and communication.

10.2.4 Education, communication and public engagement In a recent review of the representation of ‘biogeography’ in UK newspapers, websites, and blogs, Ladle (2008, p. 390) concluded that, ‘[the term biogeography has] very little presence in the public sphere, and that this probably reflects a lack of understanding of the subject area and its relevance in contemporary environmental debates’. A lack of popular representation and understanding can often be traced back to a lack of exposure during formal education. It is perhaps revealing that a search for ‘biogeography’ teaching resources in the UK government’s schools National Curriculum website carried out in August 2008 (while the site was still publically available) returned no results, whereas searching for ‘biodiversity’ resources produced 219 separate documents. This suggests that in the UK, at least, biogeography is hardly seen as a discipline at all in schools, let alone one that has a contemporary importance. It should be noted, however, that while the term biogeography is poorly represented, biogeographical subjects are richly covered across the electronic and popular media, especially those related to conservation and the future of the natural world. Unfortunately, the translation of such information into the public domain via these media often misinterprets or sensationalizes findings. An example of this was the coverage of a large study by Thomas et al. (2004), collating bioclimatic envelope modelling efforts for 1,103 species from several

different taxa and from several different regions of the world. The simulations indicated that a potentially substantial proportion (15–37 per cent being the favoured range, but perhaps as low as 5.6 per cent or above 50 per cent) of these species would be ‘committed to extinction’, based on loss of habitat driven by climate change scenarios for 2050 (see Box 7.3 for details). Despite the authors clearly describing these percentages as being ‘an estimate of proportions of species committed to future extinctions as a consequence of climate change over the next 50 years’, and ‘not the number of species that will become extinct during this period’, the global news media almost universally misreported the story. In the UK, 26 out of 29 newspaper reports were factually incorrect (Table 10.2), with the most frequent misrepresentation being that one million species would go extinct by the year 2050 (Ladle et al., 2004). The source of the remarkable claim of one million species being threatened with extinction was the press release issued by the lead author, and it may be derived by picking a value within the wide range of possible values reported in the paper, say 25 per cent, then assuming this proportionate loss can be extrapolated to all species on the planet and then making the further assumption that there may be around 4 million species of land plants and animals on Earth (Ladle et al., 2004). If Thomas had chosen a perfectly reasonable value for global species diversity of, for example 10 million then, by this reasoning, 2.5 million species would have been considered ‘committed to extinction’. Unsurprisingly, the figure of a million threatened species was widely misunderstood, as illustrated by a letter sent to a leading UK newspaper from a staff member of a high-profile national conservation NGO. The letter stated that, ‘The recent report from Professor Chris Thomas and his research team gives further evidence of the fragility of a million known species, as well as probably several million that are still unknown’ (“Revealed: how global warming will cause extinction of a million species”, J. Purvis, The Independent, 8 January, p. 19). The writer was clearly unaware that the million species were substantially made up of the ‘still unknown’. It will be appreciated by readers of this book that there are, of course, several other important assumptions inherent in the extrapolations involved here, not least that the models can be relied upon in their forecasts of species range losses and consequent extinctions, and that the species in the Thomas et al. (2004)

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Table 10.2 Results of a survey of the coverage of the Thomas et al. (2004) article on climate change and extinction by UK national and regional newspapers. Primary reports refer to direct reporting of the study, while secondary reports include letters and commentary. None of the six claims were made in the original scientific study, whose authors went to some length in the original paper or associated press release to stress the three important qualifications listed in the table. Table from Ladle et al. (2004). Primary reports

Secondary items

Claims Million or more species extinct Species extinct by 2050 Quarter of all life forms extinct Quarter of all land animals/plants extinct Third of all life forms extinct Third of all land animals/plants extinct

14 10 1 7 2 1

7 7 2 1 0 0

Qualifications Based on millions of unidentified species Only a few actually extinct by 2050 Phrase ‘committed to extinction’ used

0 2 1

0 0 0

analyses are broadly representative of all terrestrial species. In fact, the selected 1,103 species were all endemics within their respective study areas, mostly having rather restricted geographical ranges and thus comparatively narrow climate envelopes and more vulnerability to climate change (see Box 7.3). Despite the huge degree of uncertainty associated with the projections within the paper, the story made the front pages. The paper has been extremely highly cited since, featuring not just in scientific discourse but also in such places as the Stern Review on the economic implications of climate change, commissioned by the UK government, in which it was used to support an estimate of at least ten per cent species extinctions globally in response to a 1°C global temperature increase. The climate change/extinction story was also actively discussed on the internet, but here the representation was far more variable. Several sites ranked highly on popular search engines were critical of the underlying science, while traditional media sites (e.g. newspapers, newswires, etc.) were, like the print newspapers, generally uncritical and factually incorrect (Ladle et al., 2005). Such polarization of representations and oversimplification, to the point of misrepresentation, has in our view the potential to damage the

credibility of environmental science. It is equally clear that conservation biogeographers will struggle to counter misrepresentations if, like most scientists, they restrict their pronouncements and debates to academic journals and conferences. On the other hand, efforts by scientists to communicate via traditional media often run into the difficulty that these media are not receptive to representing complex and highly technical content without a great deal of simplification, and they are generally highly resistant to scientists checking and correcting the way their information is represented. An approach that one of us has advocated is the engagement of scientists in blogs (also known as weblogs), which are web-based interactive forums in which topical issues can be freely debated by anyone with access to the internet who is interested in the issue (Ashlin & Ladle, 2006). However, uncontrolled, unregulated debating sites on controversial scientific issues often appear to generate more heat than light, and it is by no means clear that they provide the most effective means of scientists engaging in public dissemination and outreach. Devictor et al. (2010) highlight a rather different means of engagement between conservation biogeographers and the public, through the medium of citizen science (Figure 10.1). Citizen science programmes

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Prospects and challenges

Figure 10.1 The role citizen science may play within conservation biogeography, highlighting: (a) the differences between those schemes in which the scientists dictate the programme, and more participative approaches; (b) five key factors in the success of citizen science programmes. Modified slightly from Devictor et al. (2010 – their Figure 1).

refer to data sets collected by the participation of the general public, and they have been more formally defined as ‘a method of integrating public outreach and scientific data collection locally, regionally, and across large geographical scales’ (Cooper et al., 2007). Typically, they involve members of the general public following a simple, standardized sampling protocol to collect data across a large number of sampling points at a particular time of year, repeated over a number of years (e.g. Table 10.3, Figure 10.1). Such data sets are not without their limitations – so, for instance, those species that are particularly difficult to detect and identify are likely to be under-sampled. However, the sheer scale of some of the data sets means that a variety of cross-validation techniques can be employed, and useful results can be extracted. Indeed, Devictor et al. (2010) comment that they know of more than 200 scientific publications resulting wholly or in part from citizen science data sets, with

contributions made to understanding, for example: mechanisms driving species responses to land-use changes; species traits most affected by global warming; impacts of acid rain on birds; changes in plant phenology; and protected area efficiency. Beyond the direct value to science, citizen science programmes have a key role to play in demystifying science, in reinforcing and extending environmental education and in engaging citizens actively in the endeavour of nature conservation. This engagement might extend from the typical top-down approach, whereby scientists dictate the form of the project and are solely responsible for the interpretation, to more participatory, bottom-up forms of engagement, in which: ‘… citizen science could be of great help to promote a conservation biogeography based on local ecological knowledge in several socio-economic contexts (i.e. not limited to the most developed countries)’ (Devictor et al., 2010, p. 360).

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Table 10.3 Some examples of citizen science programmes, (a) highlighting the scale of some prominent schemes; and (b) highlighting the nature of the involvement, skills and training involved and the means of communication of findings (from Devictor et al., 2010, their Tables 1 and 3). (a) Website (all after http//www.)

Issue

Type

Name

Characteristic

a Spatial scale

Regional

Appalachian Mountain Watch

Mountains

outdoors.org/ conservation/ mountainwatch

USA

National

French Garden Butterfly Monitoring

France

noeconservation.org

France

Continental

Spider WebWatch

North America

spiderwebwatch.org

USA, Canada

Worldwide

e-bird

worldwide

ebird.org/content/ ebird

World

Long-term data set

Christmas Bird Count

since 1900

audubon.org/bird/ cbc/

USA, Canada

Medium-term data set

UK Butterfly Monitoring Scheme

since 1976

ukbms.org

UK

Few years old programmes

FrogwatchUSA

since 1998

nwf.org/

USA

High

Nest Watch

>25,000 people

birds.cornell.edu

USA

Very High

Big Garden Bird Watch

>40,0000 people

rspb.org.uk/ birdwatch/

UK

Issue

Type

Name

Characteristic

Website (all after http//www.)

Country

a Skill

Beginner

French Garden Butterfly Observatory

no skill required

noeconservation.org

France

Intermediate

Nocturnal Owl Survey

ability to identify few species

bsc-eoc.org/volunteer

Canada

Confirmed

French Butterfly Monitoring

ability to identify many species

mnhn.fr/vigie-nature

France

Occasional observation

e-bird

no commitment

ebird.org/content/ ebird

World

One-day event per year

Bailly Birdathlon

during a 24-hour period in May, to find as many bird species as possible

bsc-eoc.org/support/ birdathon

USA, Canada

b Temporal scale

c Sample size

Country

(b)

b Time required

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Prospects and challenges

Table 10.3 Continued (b)

Issue

c Education

d Communication

Website (all after http//www.)

Country

ten minutes, one evening a week throughout the summer

mos.org/fireflywatch

USA

UK Butterfly Monitoring Scheme

26 counts a year and 5 hours by count

ukbms.org

UK

Special training tools

Frog Watch

frog calls’ records

naturewatch.ca

Canada

Field training / workshop

Anglers Monitoring Initiative

1 day workshop obligatory

riverflies.org

UK

Special activities for children

Great Lake Worm Watch

downloads of educative games

nrri.umn.edu/worms

USA

Special tools for teachers

RoadKill

12 interactive activities proposed

http:// roadkill.edutel.com/

USA

Special tools for university students

Project BudBurst

exercises of data collection

budburst.ucar.edu/

USA

Tools to analyse its own data

e-bird

access to maps and data important to the observer

ebird.org/content/ ebird

World

Press room

Feeder Watch

free downloads of reports, videos and press articles

birds.cornell.edu/pfw

USA, Canada

Pollution and conservation alert

Anglers Monitoring Initiative

update of species conservation status

riverflies.org

UK

Results online

All of them to varying extents (interactive maps, graphs, reports …)

Type

Name

Characteristic

Regular commitment on a season

Firefly Watch

High commitment

Future directions 10.2.5 Reconciliation ecology and a biogeography of the countryside Conservation philosophy, science, and practice must be framed against the reality of human-dominated ecosystems, rather than the separation of humanity and nature underlying the modern conservation movement. (Western, 2001, p. 5458) Rosenzweig (2001, 2003) has argued persuasively that the most fundamental conservation challenge facing us is to learn ‘how to share anthropogenic habitats with wild species’ (2001, p. 5409). In other words, we need to discover ways to transform and diversify what are often impoverished anthropogenic habitats so that they have the potential to harbour more species. Daily et al. (2001, 2003) make much the same case, arguing for recognition of the importance of making anthropogenic habitats as wildlife-friendly as possible, both for the ecosystem services (e.g. pollination) provided by wild species in these landscapes and because protected area systems alone cannot save enough of wild nature. Daily and colleagues issue their plea under the term ‘countryside biogeography’, while Rosenzweig uses the label ‘reconciliation ecology’ in recognition of the inevitable trade-offs and compromises that need to be made in such an undertaking. Rosenzweig’s case is firmly based on biogeographical principles; he reasons that, because reduction in area of suitable habitat is the key driver of reduction in diversity, we need to halt or reverse this trend (Rosenzweig, 2003; and see Chapter 8). In recent decades, a big part of the effort to do so has been channelled through the creation of protected areas, but at roughly 12 per cent of the terrestrial surface of the Earth under some form of protection it is likely to become increasingly difficult to argue for significant further increases in the terrestrial protected area estate. We may be approaching the limits of this strategy, at least in terms of the more restrictive categories of protected area (e.g. IUCN categories 1–4; Table 2.2). Rosenzweig’s argument is that by changing the way we transform habitats and how we manage already transformed habitats, we can potentially achieve our socio-economic objectives without the accompanying reduction in species ranges. Both countryside biogeography and reconciliation ecology derive directly from an understanding of

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biogeography at the landscape scale, and in particular of the logic of species–area relationships, the role of the transformed matrix (i.e. ‘countryside’), habitat corridors, dispersal and metapopulation dynamics (see Chapter 8). The conservation biogeography toolbox of conceptual frameworks, models and forecasts will thus be central to the success of the reconciliation ecology agenda. There are already many ongoing initiatives that are aligned with this agenda, most notably in modifications to agricultural practices that are extending the geographical area available to many plants and animals, or which allow the passage of species between habitat fragments. Ultimately, the success of reconciliation ecology in ‘pushing’ nature back up the species–area curve will depend on the broad support of many sectors of society for the goals and values encompassed. For large-scale habitat modifications, it will be necessary to engage with a whole range of disciplines and professions, including planners, developers, anthropologists and social scientists, who can both help design effective strategies and also ‘broker the deal’ with the human inhabitants of these landscapes.

10. 3 LOOK I N G T O T H E FU T U R E In Chapter 1 we defined conservation biogeography as ‘the application of biogeographical principles, theories, and analyses, being those concerned with the distributional dynamics of taxa individually and collectively, to problems concerning the conservation of biodiversity’ (original source: Whittaker et al., 2005). However, as outlined in Chapter 2, conservation is a reflection of societal values. Thus, although the models and insights of conservation biogeography may help us better conserve biodiversity, the type of biodiversity we seek to conserve depends upon the values held by the various organizations and institutions that determine conservation policy. These bodies are, in turn, influenced by the values of wider society as a whole. Over time, values change – just consider the change in typical western attitudes to fur coats in the last 50 years – and, while there is no guarantee that society will continue to endorse the current objectives of the conservation movement, it will represent a great failure if conservationists cannot convince society at large that wild nature matters to the human condition and warrants more careful custodianship. Conservation biogeographers have the potential to play an important

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Prospects and challenges

role in ensuring public support for nature conservation if they can succeed in converting their research findings into compelling narratives – modern parables that are in tune with the changing social values of nature (Bowman, 2001). It is an exciting time to be a biogeographer: the discipline is awash with new tools, techniques, and accessible, large-scale data sets, with the promise of more systematic local, regional, and global information systems to come in the near future. Conservation

biogeography, the intersection of a venerable academic discipline with some of the world’s most pressing environmental problems, surely has a vital role to play in informing future debates about the conservation of biodiversity. We hope that this book has given you the sense of a discipline in a state of intellectual ferment, offering promise and potential, as scientists from around the globe tackle some of the biggest issues to face the natural world.

Glossary of Terms

Adaptive ecosystem management: A form of conservation management that integrates scientific knowledge of ecological relationships within a complex socio-political and values framework towards the general goal of protecting native ecosystem integrity over the long term (Grumbine, 1994). Anthropocene: A term used by some scientists to refer to the most recent period of Earth’s history, where humans began to have a significant global impact on the environment and climate system. It has no precise start date, but is often considered to start in the late 18th century. Bioclimate envelope models: See Species distribution models. Biodiversity: A contraction of ‘biological diversity’. Biodiversity has many definitions, one prominent one being ‘[t]he variability of life from all sources, including within species, between species, and of ecosystems’ (Matthews et al., 2001). Some commentators have noted that biodiversity definitions are often closer to subjective ‘value judgement’ concepts such as quality of life than an objective measure of an environmental property. Biodiversity hotspot: An area high in selected biodiversity attributes, such as species richness or endemism; sometimes a biodiversity attributes analysis is combined with a threat criterion to provide a composite hotspot/threatspot analysis. The most prominent ‘hotspots’ scheme in conservation is of the latter form, being that developed by the international NGO Conservation International, whose technical definition of a hotspot is a geographical area that contains at least 0.5% or 1,500 species of vascular plants as endemics, and which has lost at least 70% of its primary vegetation (Myers et al., 2000).

Biological invasion: The process whereby species expand their geographical distribution outside of their natural dispersal range via the actions of humans. (cf. non-native species, naturalized species, invasive species). Biome: A major type of natural vegetation that occurs wherever a particular mix of climatic and edaphic conditions is encountered; or we may equate it with the notion of a major ecosystem type. The latter usage translates rather better into the marine realm than ‘natural vegetation type’. Biophilia: The innate emotional affiliation of humans to other forms of life, which predisposes humanity to value life and living systems. Biotic homogenization: The process by which the genetic, taxonomic or functional similarities of regional biotas increase over time. Climate envelope models: see Species distribution models. Complementarity: The principle that in designing reserve networks to maximize the total number of species ‘saved’ with least effort (expenditure), you should seek sites that complement one another rather than simply designating the sites that are individually most diverse. Conservation area network: A network of areas that perform a conservation function, whether they are strictly protected or not. See Protected area network. Conservation biogeography: ‘[T]he application of biogeographical principles, theories and analyses, being those concerned with the distributional dynamics of taxa individually and collectively, to problems concerning the conservation of biodiversity’ (Whittaker et al., 2005, p. 3).

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

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Glossary of terms

Conservation biology: The area of applied research designed to inform management decisions concerning the conservation of biodiversity. The roots of conservation biology lie largely within the mid-20th century, although it was not until the 1970s and early 1980s that it was formally identified as an academic sub-discipline with dedicated journals and textbooks. Conservation values: Beliefs and ideas about nature and wildlife that inform assessments of worth. Convention on Biological Diversity (CBD): The CBD was signed on 5 June 1992 at the United Nations Conference on the Environment and Development in Rio de Janeiro. The Convention has three main aims: the conservation of biodiversity; the sustainable use of its components; and sharing the benefits from the commercial and other utilization of genetic resources in a fair and equitable way. At the heart of the CBD is the acknowledgement that biodiversity is essential for human existence and that the utilitarian (use) value is the key to effective conservation. Cultural landscape: A landscape dominated by anthropogenic habitats, within which may be embedded fragments of less impacted land with higher conservation value. Decision support tools: Computer-based information systems intended to help decision-makers compile and analyse data to help solve conservation problems. Earth Summit: An informal term for the United Nations Conference on Environment and Development (UNCED) held in Rio de Janeiro, Brazil, on 3–14 June 1992. Ecological extinction: Species that are extinct in the wild but which have an extant captive bred population, or are present in the wild but at such low densities that they no longer interact to a meaningful degree with other species in the community. Ecological relict: ‘A community or fragment of one that has survived some important change, often to become in appearance an integral part of the existing vegetation’ (Clements, 1934, p. 42). Ecoregions: ‘Regions of relative homogeneity in ecological systems or in relationships between organisms and their environments’ Omernik (1987, p. 123). Ecosystem resilience: The capacity of ecosystems to absorb disturbance; and their ability to reorganize when a critical threshold is exceeded.

Ecosystem services: The processes or products of natural ecosystems and species that provide a utilitarian value, e.g. natural processes such as pollination and watershed protection which, if removed, would have economic consequences. Edge effects: The ecological influence of the altered physical and biotic properties that typically characterize the edge of a habitat type, and which extend towards the core of the habitat patch. The term is most frequently applied within applied island biogeography when discussing characteristics of habitat islands. Endemism: A species (or other taxonomic entity) is endemic when it is naturally confined to a defined geographical area. Some authors have sought to delimit the term as applying at a particular scale, or with particular physical units of analysis; herein we do not so delimit it. See Range restricted species. Environmental surrogate: A physical or climatic variable used as a proxy to derive ecological classifications. Environmental surrogates can incorporate some biotic variables such as vegetation. Extinction: Used to refer to the disappearance of a species from an area (also termed extirpation) or globally. Various ways the term can be used are described in Table 4.2. Extinction debt: The anticipated eventual species loss from an area due to habitat loss and fragmentation. Fortress conservation: The practice of completely excluding local people from protected areas. The term is often used more broadly to apply to any conservation practice that excludes local people from access to, or exploitation of, natural resources. Habitat island: Areas isolated from other reserves by anthropogenically transformed habitats (sometimes named the ‘matrix’) that are generally unsuitable for the species of conservation concern. Historical (phylogenetic) biogeography: The broad branch of biogeography that concerns itself with historical interpretations of biogeographical patterns. The term can have quite specific meanings, e.g. the derivation of cladograms for areas based on the phylogenies of the organisms inhabiting these areas. Homology/homologous: Characters that are similar in different taxa because they are shared through a common ancestor.

Glossary of terms

Invasive species: A species that expands its population from the site of original arrival into intact or semi-intact vegetation (regardless of demonstrated impacts). Island rule: The general tendency for an ordered size changes of island vertebrate species in relation to mainland congeners, such that large-bodied species tend to get smaller and vice versa. Keystone species: A species that has a disproportionate effect on its environment relative to its biomass. Such organisms typically have a strong influence on many other organisms within an ecosystem and may play an important role in determining the structure of the ecological community. Linnean extinction: Extinctions of undiscovered species inferred from the species–area relationship and estimates of species diversity for a given ecosystem or region. The assumed losses of these inferred species have been termed Centinelan extinctions by Wilson (1992). Linnean shortfall: The discrepancy between the number of species that have been formally described by taxonomists and the number of species that are thought to exist. Named after the father of modern nomenclature, Caroleus Linnaeus (a Latinized form of Carl von Linné). Local extinction: The complete loss of a population of a species within a clearly defined geographical area, but where extant free-living populations still exist outside that area. Macroecology: A top-down and multi-scale approach to analyses of the structure of biotas, focused on the emergent outcomes of statistical analysis of key properties (e.g. species abundance, distribution and diversity) in order to understand the processes involved in structuring ecological systems. It is often, but not necessarily, concerned with analyses of coarse-scale data sets of large spatial extent. Mass extinction event: A major episode of extinction involving many different taxa and occurring fairly suddenly in the fossil record. Matrix (or habitat matrix): Anthropogenically transformed habitats that are generally unsuitable for the species of conservation concern within which habitat islands (which sometimes are also conservation areas) are embedded. Mesopredator release: The increasing number of smaller omnivores and predators due to the absence of larger predators – a phenomenon that can result from habitat fragmentation.

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Metapopulation: A population of geographically separated subpopulations interconnected by patterns of gene flow, extinction and recolonization. Minimum viable population (MVP): The minimum size of a population that will ensure its long-term survival – for instance, providing a 95% probability of persistence for the next 100 or 1,000 years. Monophyletic: A group (taxon or gene) that includes an ancestral group and all of its descendant groups, also referred to as a clade. Naturalistic fallacy: The assumption that, prior to invasion by non-indigenous species, ecosystems must have been natural and pristine, and that this is the ecosystem state that should be preserved. Naturalized species: Species that have selfsustaining populations at the original point of establishment or only in highly modified habitats. Naturdenkmal: Roughly translated as nature monument, a concept that was articulated by the Berlin-based forester Hugo Conwentz in the first decade of the 20th century to mean a place for the study and contemplation of nature, motivated by the belief that monuments of nature have value to human civilization, culture and identity. Nature conservation: A social movement working to develop or reassert certain values in society concerning the human/nature relationship. Non-analogue community: Used in palaeoecology to refer to past communities having a species composition for which there is no contemporary equivalent, such systems providing evidence of the essentially individualistic nature of community assemblage processes. Projected into the future, the term is sometimes used to describe hypothesized future communities assembled under the influence of anthropogenic change processes. Non-native species: Populations that have become established outside the bounds of their native ranges through the action of humanmediated transport. Paraphyletic group: An incomplete evolutionary unit in which one or more descendants of a particular ancestor have been excluded from the group. Phylogeography: The study of the genetic and geographical structure of populations and species. Phylogenetic diversity (PD): A measure of biodiversity that incorporates taxonomic difference between species, e.g. based on estimating the length

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Glossary of terms

of evolutionary pathways that connect a given set of taxa. Phytosociology: A sub-discipline of plant community ecology that seeks to describe and understand plant species co-occurrences at the level of local communities. Phoenix extinction: A species that is extinct in the wild, but for which genetic material is available in the form of stored material or a closely related conspecific or congeneric variety/breed/hybrid, allowing the possibility of a future reintroduction of the species or a functionally equivalent form. Population bottleneck: A severe, but perhaps fleeting, reduction in population number that is often associated with reduced genetic diversity and heterozygosity, and reduced adaptability of that population. Population viability analysis: A sophisticated family of models that use demographic and environmental information to provide an estimate of the probability of a population surviving over a given time period (often 1,000 years). Protected area: An area of land or sea designated in order to conserve or protect attributes of nature valued by society, groups or individuals. Many have been created largely for their biodiversity value and/ or are now managed to maintain or enhance their biodiversity value, often alongside other management goals and considerations. Protected area networks: A number of protected areas within a given geographical area that have been chosen to fulfil a shared conservation goal (e.g. minimizing future biodiversity loss). The constituent protected areas may, or may not, be ecologically connected, although this is often assumed. Examples include the European Natura 2000 network. See Conservation area networks. Range restricted species: Species that occupy a very limited geographical range due to specialized habitat preferences. Reciprocally monophyletic: A term used in phylogenetics to describe the point at which, for a given gene, alleles within closely-related taxa or clades are monophyletic with respect to each other. Red List: A list of species that are considered to be threatened with extinction. The best known example is the IUCN Red List, which provides information on the taxonomic, conservation status and distribution of taxa that are facing a high risk of global extinction.

Redundancy: This term can have varied meanings, one of which is the strategy of choosing multiple protected areas that contain the same species (or other attribute) of conservation interest in order to decrease the long-term probability of its extinction, i.e. to provide a certain level of resilience within a protected area network. Representation principle: The idea that conservation schemes should seek to conserve systems of sites that are representative of a set of community types, major ecosystem types, biogeographical zones, species, etc. Restoration ecology: An attempt to ‘move a damaged system to an ecological state that is within some acceptable limits relative to a less disturbed system’ (Falk et al., 2006). Rewilding: Refers to ‘action on the landscape level with a goal of reducing human control and allowing ecological and evolutionary processes to reassert themselves’ (Klyza, 2001). Sink population: A breeding group that does not produce enough offspring to maintain itself over several generations without immigrants from other populations. Sink habitat: A habitat in which local mortality exceeds local reproductive success for a given species. Source population: A breeding group that produces enough offspring to be self-sustaining, and that often produces excess offspring that augment the population of other areas nearby. Source habitat: A habitat in which local reproductive success exceeds local mortality for a given species. Species distribution models: Species distribution models relate field observations of the presence/ absence of a species to environmental predictor variables, based on statistically or theoretically derived response surfaces, for prediction and inference. The predictor variables are often climatic but can include other environmental variables. Species relaxation: The decline of species number towards an eventual, hypothetical and lower equilibrium, due to the dominance of extinction over immigration after fragmentation. Synanthropic species: Also known as urban exploiters, theses are species that are able to thrive in the novel conditions created as a landscape is urbanized. Systematic conservation planning: A discipline aiming to maximize the efficiency and effectiveness of protected area network design. The key

Glossary of terms

technique of systematic conservation planning is to create a range of hypothetical alternative networks that enable planners to engage with the complexity of multi-sectoral spatial planning. Systematic conservation planning is concerned with the optimal application of spatially-explicit conservation management actions to promote the persistence of biodiversity and other natural features in situ. Trophic cascade: The chain of knock-on extinctions observed or predicted to occur following the

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loss of one or a few species that play a critical role (e.g. as a pollinator) in ecosystem functioning. Wallacean extinction: Species that have not been documented for many years, but in which final extinction is uncertain because populations might survive in areas that have not been surveyed within the potential distributional range. Wallacean shortfall: The inadequacy of scientific knowledge of the geographical distributions of species.

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index

adaptive cycles 37 alien species 27 anthropogenic biomes 81 anthropogenic extinction 47, 58 archipelagic scale of SAR 195 Area of Outstanding Natural Beauty 93 assembly rules 221–222 assisted migration 185–188 Atlas Florae Europaeæ 57 Avise, John 72 azonal protected areas 95 hotspots concept 96–98 important areas 97 Azores extinction debt 204–206 Barcode of Life Initiative 53 baselines defining and using 38 derived from long-term ecology 39–41 derived from relict pristine systems 38–39 rapid environmental change 42 rewilding 41–42 bigfoot 70–72 bioclimatic envelope modelling (BEMs) 69–70 biodiversity 47 fundamental taxonomic units 62 Evolutionary Significant Units (ESUs) 64–65 other units 65 species versus other geneticallybased units 62–64 incomplete knowledge 47–48

knowledge shortfalls extinction estimate shortfall 58–62 Linnean shortfall 49–54 Wallacean shortfall 54–58 mapping function 76 biomes, ecosystems and communities 76–81 ecoregions 82–83, 87–91, 96, 101–103, 111, 113–117 mapping, reasons for 48 explaining species types in geographical areas 49 reconstruction of historical development 49 region classification based upon biotas 49 marine realm 83–91 predicting change 176–177 modelling current distributions 177–180 modelling range shifts 180–183 spatial distributions biogeographical regions 75–76, 77 endemism 74–75 mapping species 65–72 phylogeography 72–74 biodiversity conservation 6 Biodiversity Information Standards 53 biogeographical provinces 105, 106 biogeographical realms 105 biological species concept (BSC) 63 biomes 78–81 modelling current distributions 177–180

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

biotic homogenization 229–230 conservation 240–241 environmental and human drivers 238–240 manifestations 230–232 novel assemblages 241–242 patterns 232 birds 235–237 fishes 232–235 mammals 237–238 plants 237 process 230 biotic realms 105 biounits 105 Birdlife International 103, 107 Birds Directive (EU) 23, 117 body sizes on islands 221 Boone and Crockett Club (B&CC) 20 carrying capacity 33 Catalogue of Life 53 centres of richness and endemism (CORE) 74 cheese cutter spatial division 97, 98 cherry picking spatial division 97, 98 citizen science 254 climate 33 climate change 185–188 climatic envelope modelling (CEMs) 69–70 climax theory 33 Coastal Zone Management (CZM) 132 Commons Preservation Society 18 communication of biogeographic ideas 252–256 communities 78 community conservation areas 22

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Index

Community Conserved Areas (CCAs) 22 Community-Based Natural Resource Management (CBNRM) 22 complementarity 95, 138, 151, 241 compositionalism 31–32 conservation biogeography areas of current research 248 challenges 248 education, communication and public engagement 252–256 filling Wallacean and Linnean shortfalls 248–250 improving models, simulations and forecasts 250–251 reconciliation ecology 257 turning theory into practice 251–252 definition 3–4 emergence 4–7 future directions 257–258 need for 247–248 scope 7 diversity 8 scale 8 taxonomic units 62 Evolutionary Significant Units (ESUs) 64–65 other units 65 species versus other geneticallybased units 62–64 conservation biology 4–7 general characteristics 6 Conservation International (CI) hotspots 74–75, 97–98, 109–113 conservation movement 5 foundational values 16 conservation practice aims 31 social values 26 attitudes to non-native species 26–28 restoration and rewilding 28–29 conservation resource allocation problem 154 contagion hypothesis 174 Convention Concerning the Protection of World Cultural and Natural Heritage 23 Convention on Biological Diversity (CBD) 10, 18, 22–23 Conwentz, Hugo 19 cookie cutter spatial division 97, 98 correlative species distribution 178 corridors 217–218

country parks 18–19 countryside biogeography 257 critical seascapes 132 cryptozoogeographical distributions 70–72 Dasmann–Udvardy scheme 104–106 decision support tools for planning 152–155 demographical hypothesis 172 differentiation diversity 8 Directive on the Conservation of Wild Birds (EU) 117 dispersal events 226 distribution maps 55–56 diversity scale 9 dynamic arenas 115 Earth Summit 18 ecological biogeography 5 ecological extinctions 59, 60, 61 ecological surprises 34 ecoregions 82 ecosystems 76–78 adaptive management 42–44 balance versus flux 32–34 composition and function 31–32 flux 34–37 predicting the future 168–169 edge effects 216–217 education of biogeographic ideas 252–256 efficiency 139–140 plan development 151–152 Elton, Charles 34 Encyclopedia of Life (EOL) project 52, 53 Endemic Bird Areas (EBAs) 103, 106–108 endemism 8, 74–75 environmental change, rapid 42 environmental surrogates 144–146 Equilibrium Theory of Island Biogeography (ETIB) 190, 191 design guidelines 192 eukaryotes world distribution 51 European Distributed Institute of Taxonomy (EDIT) 53 Evolutionary Significant Units (ESUs) 64–65 extinction debt 200–203 Azores 204–206 extinction estimate shortfall 58–62 extinctions ecological 59, 60, 61 Linnean 58–59, 60

local 59, 60, 61 phoenix 59, 60, 61 predicting 181–183, 198 global species loss 199 time to extinction 202 rates 58 true 59, 60, 61–62 Wallacean 59–61 feedback 34, 37 flexibility 140 plan development 152 focal scale 8 focal species 141 forest reserves 94 forests, High Conservation Value approach 99–100 fragmentation 212 functional homogenization 232 functionalism 31–32 fundamental niche 179 game reserves 17–18, 94 gap analysis 124–125 genetic homogenization 231–232 geographical extent 8 Geographical Information System (GIS) 124 geographical range collapse 170–176 dynamic nature 172–174 geographical scale 7 Global Biodiversity Information Facility (GBIF) 53, 249 habitat corridors 217–218 habitat islands 137, 190 ecosystem collapse 203–208 life history 220 habitat/special management area, IUCN category 24 Habitats Directive (EU) 23 habitats modelling current distributions 177–180 Hierarchical Patch Dynamics Paradigm (HPDP) 34–35 hierarchy of processes 36 High Conservation Value approach 99–100 high seas protected areas 132–134 historical biogeography 4 homogenization of species 229–230 conservation 240–241 environmental and human drivers 238–240 manifestations 230–232 novel assemblages 241–242

Index

patterns 232 birds 235–237 fishes 232–235 mammals 237–238 plants 237 process 230 hotspots 74–75, 96–98, 103, 107, 109–113, 126, 130–132 human population density and species richness 240 human-assisted dispersal of species 226–227 airline traffic 228 ship traffic 227–228 Important Bird Areas (IBA) programme 117–119 incidence functions 210–211 instrumental values 14 Inter-American Biodiversity Information Network (IABIN) 249 Intergovernmental Panel on Climate Change (IPCC) 42 International Council for Bird Preservation (ICBP) 106–107, 117–118 International Institute for Species Exploration 53 International Union for the Conservation of Nature and Natural Resources (IUCN) 23–25 Biogeographical Regions 104–106 protected area categories 24 interprovincial scale of SAR 195 intraprovincial scale of SAR 195 intrinsic values 14 invasive species 27, 224 biogeography economic and ecological impacts 227–229 human-assisted versus prehistoric 226–227 process of invasion 224–226 irreplaceability of sites 152 island biogeography 190–194 guidelines for conservation 219–222 implications of habitat loss and fragmentation 194 ecosystem collapse 203–208 relaxation and extinction debt 199–203 species–area relationships 194–199 nestedness 213–216 edge effects 216–217 habitat corridors 217–218 landscape context 218–219

species incidence metapopulation dynamics 211–213 minimum viable populations 208–211 island species–area relationships (ISARs) 196–197 Jurassic Coast World Heritage site 23 Key Biodiversity Areas (KBAs) 119–121 keystone species 141 landscape scale 7 large marine ecosystems (LMEs) 130–131 latitudinal changes 176 Legacy Infrastructure Network for Natural Environments (LINNE) 53 Leopold, Aldo 21 Linne, Karl von 49–50 Linnean extinctions 58–59, 60 Linnean shortfall 49–54 filling 248–250 local extinctions 59, 60, 61 logistic curve 33 London Convention on African Wildlife 20 Major Ecosystem Types (METs) 48 major habitat type (MHT) 116 marine 131 management unit (MU) 65 mapping function 76 biomes, ecosystems and communities 76–81 ecoregions 82 protocol 65–72 purposes 102 Marine Ecoregions of the World (MEOW) 125 marine realm biodiversity 83–91 conservation initiatives 127–129 protected areas (MPAs) Coastal Zone Management and critical seascapes 132 connectivity networks 133 global representation system 123–126 high seas protected areas 132–134 large marine ecosystems 130–131 origins and expansion 122–123 reefs at risk 126–130

299

status 121–122 WWF Global 200 131–132 Systematic Conservation Planning 147–148 mathematical modelling current distribution of species, habitats and biomes 177–180 range shifts 180–183 Systematic Conservation Planning 152–155 matrix effects 218–219, 222 maximal coverage conservation prioritization problem 153 maximum sustainable yield 33 mesoecosystems 115 metapopulation dynamics 211–213 minimum set problem 153 minimum viable area (MVA) 209–210 minimum viable population (MVP) 193, 208–211 mitochondrial DNA (mtDNA) 65 Muir, John 20–21 national parks 21, 94 IUCN category 24 National Vegetation Classification (NVC) 39, 79–80, 100–101 Natura 2000 (EU) 23 natural disturbances 41 natural monument, IUCN category 24 natural sacred sites 17 Naturdenkmal 94 nature conservation 6 nature movements 19 nature reserves 19 nestedness 213–216 non-analogue communities 40 non-native species, attitudes to 26–28 novel assemblages 241–242 operational geographical units (OGUs) 75 operational model for pragmatic conservation planning 157 palaeoecology 4, 40 Partnership for Enhancing Expertise in Taxonomy (PEET) 52, 53 persistence (adequacy) 139, 163–164, 188, 191 dynamic conservation planning 183 biotic and abiotic processes 183–185 climate change and assisted migration 185–188 long-term ecological insights 184 socio-economic factors 185

300

Index

past, present and future 164–167 geographical range collapse 170–176 integrating evolutionary considerations 166–167 interpreting recent trends 169–170 predicting future ecosystems 168–169 plan development 146–151 predicting biodiversity change 176–177 modelling current distributions 177–180 modelling range shifts 180–183 phoenix extinctions 59, 60, 61 phylogenetic difference 143 phylogenetic diversity (PD) 166 phylogeography 49, 72–74 phytosociology 39 Pinchot, Gifford 18 Planetary Biodiversity Inventories (PBI) 53 pollen record analysis 171 population changes from human activities 169–170 population scale 7 population viability analysis (PVA) 209 populations 33 pristine systems, relict 38–39 prokaryotes, world distribution 51 protected area with sustainable use of natural resources, IUCN category 24 protected areas 14–16 community conservation areas 22 current trends and future directions 134–135 factors influencing establishment 15 framework typology 95–97 biogeographical versus ecological approaches 100–102 spatial classification 97–100 strategic goals 102–104 growth 15 international categorization system 23–25 marine schemes Coastal Zone Management and critical seascapes 132 connectivity networks 133 global representation system 123–126 high seas protected areas 132–134 large marine ecosystems 130–131 origins and expansion 122–123 reefs at risk 126–130 status 121–122 WWF Global 200 131–132

national parks 21 nature movements and nature reserves 19 origins 93–95 resource and game reserves 17–18 sacred sites 16–17 social purpose classification 25 state and country parks 18–19 terrestrial schemes 104 Conservation International (CI) hotspots 109–113 Endemic Bird Areas 106–108 Important Bird Areas (IBA) programme 117–119 IUCN Biogeographical Regions scheme 104–106 Key Biodiversity Areas (KBAs) 119–121 WWF Ecoregions scheme 113–117 wilderness areas 20–21 wildlife sanctuaries and refuges 19–20 protected landscape/seascape, IUCN category 24 public engagement with biogeographic ideas 252–256 Rabinowitz’s seven forms of rarity 67–69 range restriction 67–69 range shifts collapse 170–176 modelling 180–183 rapid environmental change 42 rarity of species 67–69 realized niche 179 reconciliation ecology 257 Red List 67–69, 128 reefs at risk 126–130 relaxation 199–201 relict pristine systems 38–39 representation (principle) 96–98, 101, 103–105, 123–125, 131 representativeness 138–139, 191 plan development 140 environmental surrogates 144–146 species-based surrogates 140–144 reserve designations, international 22–23 resource areas 17–18 restoration of habitats 28–29 rewilding 28–29, 41–42 sacred sites 16–17 sasquatch 70–72 scientific benchmark sites 94 self-organization 35 Sierra Club 20–21

Single Large Or Several Small (SLOSS) reserves debate 193 single-island endemics (SIEs) 197 sinks 218 Sites of Special Scientific Interest (SSSIs) 22, 39 small island effect 221 social purpose classification of protected areas 25 social values 13–14, 29–30 conservation practice 26 attitudes to non-native species 26–28 restoration and rewilding 28–29 international reserve designations 22–23 international system for categorizing protected areas 23–25 protected areas 14–16 community conservation areas 22 factors influencing establishment 15 growth 15 national parks 21 nature movements and nature reserves 19 resource and game reserves 17–18 sacred sites 16–17 state and country parks 18–19 wilderness areas 20–21 wildlife sanctuaries and refuges 19–20 Society for the Preservation of the Wild Fauna of the Empire (SPWFE) 20 socio-economic factors in dynamic conservation planning 185 reserve network design 186–187 spatial scale 8 Special Areas of Conservation 23 Special Protection Areas (SPAs) 23 species major concepts 63 mapping 65–72 modelling current distributions 177–180 species accumulation curves (SACs) 196 species–area relationships (SAR) 194–199 scales 195–197 species-based surrogates 140–144 species density 8 species description cumulative curves 51 species distribution models 69–72, 181–183 species diversity 8 species richness 8 human population density 240 species turnover 8

Index

state parks 18–19 stochastic variation 34 Strategic Adaptive Management 43 strict nature reserve, IUCN category 24 Systematic Conservation Planning 95, 136–137 concepts and principles efficiency 139–140 flexibility 140 persistence (adequacy) 139 representativeness 138–139 consultation and implementation 155–156 decision support tools 152–155 definition 138 development 140 achieving efficiency 151–152 achieving flexibility 152 achieving persistence 146–151 achieving representation 140 future directions 156–157 attainability of persistence 159 better economics and socio-economics 158 changing assets with time 158 dealing with uncertainty 158–159

dynamic problem 158 investment 159 mix of actions 158 threats 159 taboos 17 taxonomic units of conservation biogeography 62 Evolutionary Significant Units (ESUs) 64–65 other units 65 species versus other genetically-based units 62–64 Thoreau, Henry David 20 threatened species 47, 142, 202–203 threshold behaviour 35 thresholds 220–221 Thresholds of Potential Concern (TPC) 42–44 time to extinction 202 transition 35–37 true extinctions 59, 60, 61–62 umbrella species 141 urban/rural gradient studies 239–240

301

Vera, Franz 41 Wallace, Alfred Russel 48–49, 54–55 Wallacean extinctions 59–61 Wallacean shortfall 54–58 filling 248–250 Watt, Alexander 34 wetland degradation 239 wilderness areas 20–21 IUCN category 24 Wilderness Society 21 wildlife conservation 6 wildlife sanctuaries and refuges 19–20, 94 World Wildlife Fund (WWF) 22 Ecoregions scheme 111, 113–117 Global 200 programme 131–132 Yellowstone National Park 20 Yosemite 20 zonal protected areas 95 functional approach 96–97 representation–compositionalist approach 96

Plate 3.4 Landscape stability in alternative steady states. Superimposed lines show possible non-linear change from a nondegraded ‘steady state’ before 1430 cal yr BP, through a 600-year transition period leading to the modern degraded ‘steady state’ after 800 cal yr BP. T1 and T2 represent likely positions of major thresholds in the system. The dashed arrows from T2 show possible future trajectories of landscape recovery. From Dearing (2008).

Conservation Biogeography Edited by Richard J. Ladle and Robert J. Whittaker © 2011 Blackwell Publishing Ltd. ISBN: 978-1-444-33503-3

(a)

(e)

(b)

(f)

(c)

(g)

(d)

(h)

Plate 4.2 Steps in producing hypothetical distribution maps, illustrated (panels a–d) for a species with a restricted distribution (Inga plumifera), and (panels e–h) for a species with a widespread distribution (Inga capitata). (a) and (e) the degree squares with confirmed occurrences. (b) and (f) the contours of the predicted probability of occurrence, using a probability of occurrence in adjacent degree squares of 0.5 and allowing this effect to accumulate for 5 degree squares. (c) and (g) the hypothetical distribution deduced by accepting a probability of occurrence of greater than 0.5 in any degree square. (d) and (h) the degree squares for each species. Summing these values across all species modelled in the exercise allows the estimation of the total number of species hypothetically occurring in any one degree square. From Hopkins (2007).

(a)

(c)

(b)

(d)

Plate 4.3 Known and unknown plant diversity of the Amazon Basin, based on species occurrence data for 1,584 monographed species. Major rivers of the Amazon Basin are shown in grey. Deeper shades of blue indicate higher numbers of species per 0.5 degree grid cell; yellow the lowest values; brown areas represent land > 1000 m. (a) The distribution of information of species occurrences. (b) The distribution of the expected diversity as predicted by a bootstrap model that compares the contents of the checklists within a circle with a radius of five degree squares of the focal square. (c) The distribution of the diversity that can be explained by modelling the distributions of the 1,584 species as predicted by assuming that each has a likelihood of occurrence of 50% in degree squares adjacent to those where they are already known to occur, and this additive effect extends within a radius of five degree squares. (d) The modelled distribution of incompleteness of knowledge, derived as the difference between layers b and c. Major rivers of the Amazon Basin are shown in grey. From Hopkins (2007).

Plate 4.14 A map of anthropogenic biomes, attempting to show the present day ecological reality rather than the hypothetical or potential ‘natural’ biomes of F.E. Clements and others. Source: Alessa & Chapin (2008), after Ellis & Ramankutty (2008).

Plate 4.15 Partial illustration of eco-biogeographical classification schemes in the central Indo-Pacific, assessed during the compilation of the MEOW – Marine Ecoregions of the World (Spalding et al., 2007). (a) shows two schemes: expert-derived map of biogeographical zones and subzones (pastel shades; Kelleher et al., 1995), and bathymetry, hydrography and productivity-based Large Marine Ecosystems (blue lines and cross-hatching; Sherman & Alexander, 1989). (b) shows three schemes: prevailing wind and chlorophyll-based (pastel blocks; Longhurst, 1998), expert-derived map of the Coral Triangle and its ecoregions, based on biological and physical characteristics (green lines; Green & Sheppard, 2005), and coral distribution-based (blue lines; Veron, 2000). Some biogeographical regionalization schemes consider the entire area as part of a single Indo-Pacific province, lacking internal divisions (e.g. Briggs, 1974; Hayden et al., 1984), but this may reflect the limited data available for what is an extremely biotically diverse and complex biogeographical region. (c) shows the final MEOW classification: provinces (colour-coded) with ecoregion subdivisions. The scheme is largely based on Green & Sheppard (2005) in the east, and on Longhurst (1998) and Kelleher et al. (1995) in the west, with minor adjustments based on expert regional advice.

Plate 5.9 The ‘Global 200’ ecoregions, being those deemed ‘most important’ by WWF. Source: www.panda.org/who_we_ are/history/wwf_conservation_1961_2006/

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Plate B7.1a Genus richness and phylogenetic diversity in the Cape flora: (a) genus richness (ten quantile intervals from yellow to deep red); (b) phylogenetic diversity (PD) per cell calculated using absolute age estimates in million years (colour code as for a); (c) residuals from a loess regression of PD on genus richness. Cells with negative residuals are indicated in blue, and those with positive residuals are shown in red (shading increments of half a standard deviation); (d) the distribution of unusual PD values, as assessed by comparing the observed PD in each cell with 10,000 PD values calculated by random selection of the same number of genera from the Cape flora. Cells with significantly lower PD (P < .0.05, two-tailed) than expected are shaded in blue. Figure from Forest et al. (2007).

Plate 7.2 Geographic range change in the grasshopper warbler (Locustella naevia) in Britain. (a) Probability of occurrence in 1968–1972, estimated using the pattern of contagion among records, with blue being minimum (non-zero) probability of occurrence and warm colours high probabilities (b) The probability of extinction by 1988– 1991, where blue represents low probabilities of extinction and warm colours high probabilities. Actual occurrences in 1988–1981 are shown by black dots. Occurrence records without probability values represent range expansions. Cells of high probabilities of extinction and lacking occurrence records are concentrated towards the margins of the 1968–1972 geographic range, although other species showed markedly different patterns (and contrast with Box 7.2). Reproduced from Araújo et al. (2002b).

Plate B7.5a Four sets of complementarity areas, for plants, breeding birds, mammals and amphibians and reptiles combined, using a 50 km grid resolution, based on a maximum coverage set algorithm and with a target of identifying the best performing 10% of grid cells. The complementarity areas, shown by the open cells, are overlapped on a map of urban land use change intensity for 2020–2050 based on one of four IPCC land use change scenarios examined: the A1FI scenario, a fossil fuel-intensive world of rapid economic growth, low population growth and rapid introduction of new and more efficient technologies. The complementarity area solutions depicted on the maps are illustrative and hypothetical networks of high priority areas for each taxon. Based on Figure 2 from Araújo et al. (2008).

Plate B9.1a (a) The frequency of commercial shipping traffic along shipping routes around the world, ranging from low (blue) to high (red). From Halpern et al. (2008). (b) Global hotspots for biological invasion from ballast water, ranging from low (blue) to high (red). From Drake and Lodge (2004).