about island press Island Press is the only nonprofit organization in the United States whose principal purpose is the publication of books on environmental issues and natural resource management. We provide solutions-oriented information to professionals, public officials, business and community leaders, and concerned citizens who are shaping responses to environmental problems. In 2004, Island Press celebrates its twentieth anniversary as the leading provider of timely and practical books that take a multidisciplinary approach to critical environmental concerns. Our growing list of titles reflects our commitment to bringing the best of an expanding body of literature to the environmental community throughout North America and the world. Support for Island Press is provided by the Agua Fund, Brainerd Foundation, Geraldine R. Dodge Foundation, Doris Duke Charitable Foundation, Educational Foundation of America, The Ford Foundation, The George Gund Foundation, The William and Flora Hewlett Foundation, Henry Luce Foundation, The John D. and Catherine T. MacArthur Foundation, The Andrew W. Mellon Foundation, The Curtis and Edith Munson Foundation, National Environmental Trust, National Fish and Wildlife Foundation, The New-Land Foundation, Oak Foundation, The Overbrook Foundation, The David and Lucile Packard Foundation, The Pew Charitable Trusts, The Rockefeller Foundation, The Winslow Foundation, and other generous donors. The opinions expressed in this book are those of the author(s) and do not necessarily reflect the views of these foundations.
about the society for ecological restoration international The Society for Ecological Restoration International is an international nonprofit organization composed of members who are actively engaged in ecologically sensitive repair and management of ecosystems through an unusually broad array of experience, knowledge sets, and cultural perspectives. The mission of SER International is to promote ecological restoration as a means of sustaining the diversity of life on Earth and reestablishing an ecologically healthy relationship between nature and culture. SER International, 1955 W. Grant Road, Suite 150, Tucson, AZ 85745. Tel. (520) 622-5485, Fax (520) 622-5491, E-mail
[email protected], www.ser.org.
about the center for plant conservation The nonprofit Center for Plant Conservation (CPC) works to build a national network of community-based institutions (botanic gardens, arboreta, museums) providing professional, hands-on assistance to prevent extinction and achieve recovery for imperiled plants native to the United States. Over 20 years the activities of the CPC have grown beyond securing seed and living collections off site to include educational outreach and scientific research about imperiled plants as well as active efforts to restore those taxa most in need to the wild. The network has 32 institutions with over 80 restoration projects and collectively secures material of over 600 species in the National Collection of Endangered Plants. Hosted by the Missouri Botanical Garden in St. Louis, the national office coordinates development of best practices, maintains a Web site for professionals and the public (www.centerforplantconservation.org), supports an extensive database, informs policymakers, and works to provide stable resources through the Friends of CPC support group.
ex situ plant conservation
Society for Ecological Restoration International The Science and Practice of Ecological Restoration James Aronson, editor Donald A. Falk, associate editor Wildlife Restoration: Techniques for Habitat Analysis and Animal Monitoring, by Michael L. Morrison Ecological Restoration of Southwestern Ponderosa Pine Forests, edited by Peter Friederici and Ecological Restoration Institute at Northern Arizona University Ex Situ Plant Conservation: Supporting Species Survival in the Wild, edited by Edward O. Guerrant Jr., Kayri Havens, and Mike Maunder
Ex Situ Plant Conservation Supporting Species Survival in the Wild Edited by
Edward O. Guerrant Jr., Kayri Havens, and Mike Maunder Foreword by Peter H. Raven
Society for Ecological Restoration International Center for Plant Conservation
island press Washington Covelo London
Copyright © 2004 Island Press All rights reserved under International and Pan-American Copyright Conventions. No part of this book may be reproduced in any form or by any means without permission in writing from the publisher: Island Press, 1718 Connecticut Avenue, NW, Suite 300, Washington, DC 20009. ISLAND PRESS is a trademark of The Center for Resource Economics. Library of Congress Cataloging-in-Publication Data. Ex situ plant conservation : supporting species survival in the wild / edited by Edward O. Guerrant, Jr., Kayri Havens, and Mike Maunder ; foreword by Peter H. Raven. p. cm. (The science and practice of ecological restoration) Includes bibliographical references and index. ISBN 1-55963-874-5 (alk. paper)—ISBN 1-55963-875-3 (pbk. : alk. paper) 1. Germplasm resources, Plant. 2. Plant diversity conservation. I. Guerrant, Edward O. II. Havens, Kayri. III. Maunder, Mike. IV. Series. 639.99—dc22 2003026435 British Cataloguing-in-Publication data available. No copyright claim is made in the work of Christina Walters and Leigh Towill, employees of the federal government. Cover: Lobelia gloria-montis (Champanulaceae), a Hawaiian endemic plant from the upland swamps of Maui. Reprinted with permission from the spectacular pictorial essay on the decline of Hawaiian biodiversity, “Remains of a Rainbow,” by David Littschwager and Susan Middleton. The future for many endemic Hawaiian plants, and other species around the world, will depend on the careful use of ex situ techniques. Printed on recycled, acid-free paper Design by Teresa Bonner
Manufactured in the United States of America 10 9 8 7 6 5 4 3 2 1
For Charlie Lamoureux, passionate student of Hawaii’s native flora and its conservation, whose legacy continues to inspire the plant conservation community in Hawaii and worldwide. and For the scientists at the Vavilov Institute who, during the Siege of Leningrad, gave their lives protecting irreplaceable plant collections
contents
foreword Peter H. Raven
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preface
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acknowledgments
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introduction Ghillean T. Prance
PART I. The Scope and Potential of Ex Situ Plant Conservation 1. Ex Situ Methods: A Vital but Underused Set of Conservation Resources
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Mike Maunder, Kayri Havens, Edward O. Guerrant Jr., and Donald A. Falk
2. In Situ and Ex Situ Conservation: Philosophical and Ethical Concerns Holmes Rolston III
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3. Western Australia’s Ex Situ Program for Threatened Species: A Model Integrated Strategy for Conservation Anne Cochrane
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4. The Role of Federal Guidance and State and Federal Partnerships in Ex Situ Plant Conservation in the United States Kathryn L. Kennedy
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5. Ex Situ Support to the Conservation of Wild Populations and Habitats: Lessons from Zoos and Opportunities for Botanic Gardens Mark R. Stanley Price, Mike Maunder, and Pritpal S. Soorae
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PART II. Tools of the Trade 6. Principles for Preserving Germplasm in Gene Banks
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Christina Walters
7. Classification of Seed Storage Types for Ex Situ Conservation in Relation to Temperature and Moisture Hugh W. Pritchard
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8. Determining Dormancy-Breaking and Germination Requirements from the Fewest Seeds Carol C. Baskin and Jerry M. Baskin
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9. Pollen Storage as a Conservation Tool Leigh E. Towill
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10. Tissue Culture as a Conservation Method: An Empirical View from Hawaii Nellie Sugii and Charles Lamoureux
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11. Ex Situ Conservation Methods for Bryophytes and Pteridophytes Valerie C. Pence
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PART III. The Ecological and Evolutionary Context of Ex Situ Plant Conservation 12. Population Responses to Novel Environments: Implications for Ex Situ Plant Conservation
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Brian C. Husband and Lesley G. Campbell
13. Population Genetic Issues in Ex Situ Plant Conservation Barbara Schaal and Wesley J. Leverich
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14. Integrating Quantitative Genetics into Ex Situ Conservation and Restoration Practices Pati Vitt and Kayri Havens
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15. Effects of Seed Collection on the Extinction Risk of Perennial Plants Eric S. Menges, Edward O. Guerrant Jr., and Samara Hamzé
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16. Hybridization in Ex Situ Plant Collections: Conservation Concerns, Liabilities, and Opportunities Mike Maunder, Colin Hughes, Julie A. Hawkins, and Alastair Culham 17. Accounting for Sample Decline during Ex Situ Storage and Reintroduction Edward O. Guerrant Jr. and Peggy L. Fiedler
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PART IV. Using Ex Situ Methods Most Effectively 18. Realizing the Full Potential of Ex Situ Contributions to Global Plant Conservation
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Mike Maunder, Edward O. Guerrant Jr., Kayri Havens, and Kingsley W. Dixon appendix 1. Revised Genetic Sampling Guidelines for Conservation Collections of Rare and Endangered Plants Edward O. Guerrant Jr., Peggy L. Fiedler, Kayri Havens, and Mike Maunder appendix 2. Guidelines for Seed Storage Christina Walters appendix 3. Guidelines for Ex Situ Conservation Collection Management: Minimizing Risks Kayri Havens, Edward O. Guerrant Jr., Mike Maunder, and Pati Vitt
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appendix 4. Ex Situ Plant Conservation Organizations and Networks Kevin M. James
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about the contributors index
485 491
foreword Peter H. Raven
Plants are fundamental to all human life. They are a profoundly undervalued resource that provides food, shelter, medicines, and biomass, the substrates for life. In natural and altered communities they provide irreplaceable ecosystem services, maintaining our atmosphere, protecting topsoil, and purifying wastes. Plants enhance our daily lives through their beauty and symbolism. In short, without plants life on Earth as we know it would cease to exist. Plants hold the genetic keys to enhanced quality of life today and will help us determine whether life will be worth living tomorrow. We are facing the largest extinction crisis in 65 million years, a crisis caused largely by human population growth and consumption patterns. If present trends continue—and we could choose to take many actions that would mitigate this outcome—two out of every three species of plants, animals, and microorganisms on Earth could be gone by the end of this century. For the estimated 300,000 plant species, however, we can work together to make the picture much brighter because 85,000 of these species are estimated to be in cultivation already and because plants are easily maintained as seeds or in tissue culture or grown with human protection. We can use these methods and protect the natural areas where they occur to ensure their survival for future generations. The management of genetic lineages of plants in such artificial conditions, often as a prelude to their reintroduction in wild or managed ecosystems, is the subject of this book. Plant diversity is not evenly distributed across the planet; there are regions of extraordinary diversity, called hotspots, where flora is particularly rich and the need for conservation investment the highest. These are often regions where the conservation need outstrips the capacity, xiii
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particularly of any one agency or organization, to protect and restore threatened species. In the United States, the states of Hawaii, California, and Florida exemplify the global pattern of biodiversity loss. The largest proportion of America’s plant extinctions will occur in these fragile areas, but they can be prevented. Indeed, it can reasonably be argued that the United States, as the world’s richest nation, should not tolerate the loss of a single plant species. Although in situ habitat protection is the top priority because conserving natural communities and their intricate network of relationships allows individual species to adapt and evolve, in situ conservation by itself is not sufficient to preserve all species. With intact, high-quality habitats increasingly rare and with natural lands threatened by invasive species and pollution, conservationists need to integrate habitat restoration and species management with habitat protection. The need for large-scale habitat restoration and species reintroduction is acute. Appropriate plant stock for restoration and botanic and horticultural expertise are needed; this is a fundamental role that botanic gardens and other ex situ providers can play. As part of an integrated conservation program, ex situ conservation is a pragmatic response to an expanding crisis. As ex situ plant conservation organizations, botanic gardens have many roles beyond serving as repositories of plant material to supply restorations. They can be, and increasingly are, centers of research and venues for formal and informal education. Much of the basic information about the characteristics, distribution, and status of plants is developed at botanic gardens and similar institutions, and this information is fundamental to effective conservation efforts. Increasingly, botanic gardens are developing applied plant conservation research programs focusing on the science of small population management, plant reintroduction, germplasm preservation, and related fields. Botanic gardens also are active in all levels of botanic education, from children’s programs to graduate degree programs. They also serve as shop windows for plant science by demonstrating the importance and beauty of plants to millions of visitors per year. This role is expanding as botanic gardens take on responsibilities for landscape conservation. The world’s botanic gardens have a role to play in helping to secure important plant habitats and ecosystems. Promoting effective integration, including building new networks, optimizing the effectiveness of existing networks, and building effective relationships between land management, academic, and ex situ communities,
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is vital for successful plant conservation. The Center for Plant Conservation model for networking in the United States is a successful example of how to bring people and organizations with various resources and expertise together. This volume builds on two previous Center for Plant Conservation books on rare plant genetics and reintroduction. It examines the value and limits of ex situ methods and provides concrete recommendations to improve and integrate ex situ programs in mainstream plant conservation. This has been an overlooked area, and this book brings a new rigor to the practice of ex situ conservation by reviewing both the scientific and policy issues.
preface
This volume forms part of a logical trilogy about the practice and theory of ex situ plant conservation that has emerged from the Center for Plant Conservation (CPC), a network of botanic gardens and arboreta involved in ex situ (off-site) plant conservation and the application of integrated conservation strategies (Falk 1987, 1990). This book, like the first two CPC books, Genetics and Conservation of Rare Plants (Falk and Holsinger 1991) and Restoring Diversity: Strategies for the Reintroduction of Endangered Plants (Falk et al. 1996), was born of necessity. To have any chance of bequeathing to our descendants a world that retains a large proportion of the plant diversity we have inherited, we must act now and do so effectively. Together, these three volumes represent an attempt by the CPC community to clarify and improve the practice and theory of ex situ conservation as an integral part of plant conservation. They are intended to provide scientifically based, pragmatic, practical guidelines and recommendations to those engaged in ex situ plant conservation. These guidelines are in a sense a catalyst of their own obsolescence, representing what we know today. We hope they will encourage new research directions and lead to the incorporation of new knowledge as the field of ex situ conservation grows. The discipline of ex situ wild plant conservation is still very young. Nevertheless, the basic structure has become clear. An effective ex situ conservation project begins with the collection of a genetically appropriate and representative sample. Ultimately, the conservation value of these samples will be realized, or not, in their natural habitats. Ex situ samples are a means to an end, a tool for enhanced survival prospects in the wild. Therefore, we must also know how to use them to reestablish populations in xvii
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native habitats. Between collection and use, we need to store and manage the samples, as growing plants or dormant seed, in good condition for potentially very long periods of time. It quickly became clear to the CPC that although the basic strategy of using ex situ resources to complement in situ management is straightforward, the technical, theoretical, and practical aspects of effective ex situ conservation are not as simple. It is one thing to know we must collect genetically representative samples, store them alive and in good condition for long periods of time and be able to germinate and propagate them, and reintroduce them into the wild to restore diversity. It is quite another to know how best to accomplish these formidable tasks. As a pioneer in ex situ conservation of threatened plant species, the CPC soon realized that even the first step in the process, collecting a genetically representative sample, was not adequately understood. In 1989 the CPC convened a scientific conference in which a number of experts were brought together to discuss important issues in the development of the CPC’s now well-known genetic sampling guidelines (CPC 1991). The guidelines form the appendix of Falk and Holsinger’s Genetics and Conservation of Rare Plants (1991). The pattern was set, and the next step was to address reintroduction in a similar way. In 1993 the CPC convened a second international conference to address the underlying components that would need to be considered to develop reintroduction guidelines. The reintroduction guidelines form the appendix to a book that addresses a wide range of issues relating to reintroduction (Falk et al. 1996). What remained to be addressed were the parts in the middle: storing samples in good condition and being able to germinate, propagate, and cultivate them. Attempting to fill that gap in our understanding is what inspired a third international conference convened in 1999 by the CPC and others, notably the Royal Botanic Gardens, Kew; Berry Botanic Garden; and particularly the Chicago Botanic Garden, which generously hosted and cofunded the symposium as the 1999 installment of their annual Janet Meakin Poor Research Symposium Series. The purpose of that symposium was to assemble experts to address the parts that go into this chronologically third, albeit logically middle, part of the trilogy. This volume follows the other two in form as well as substance. The book centers around chapters on the diverse components of maintaining samples between collection and reintroduction while learning how to manage the taxa sampled. The majority of the chapters are organized into two
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main sections. The first focuses on the technical aspects of storing collections for long periods of time and associated issues such as seed germination. The second concerns some larger ecological, genetic, and evolutionary issues that must be considered between collection and use and what may happen to plants when they are used. These two sections are bracketed front and back by two smaller but no less important sections. Up front is a general introduction to what ex situ conservation is and could be. The last section starts with a summary chapter that looks to the future of what ex situ conservation may become and what we need to do to accomplish our goals. The book finishes with four appendixes. Following the example set by the first two CPC books, the first three appendixes offer practical recommendations. In Appendix 1 the editors with P. L. Fiedler revisit original genetic sampling guidelines. They incorporate 10 years of experience and place a greater emphasis on the specific purposes for which a collection is made. Appendix 2, by Christina Walters, explains how best to prepare and store seed for the long term. She emphasizes the complex relationship between the temperature and humidity at which seed is dried and the relative humidity they will experience when stored frozen at various temperatures. Appendix 3, also by the editors with P. Uitt, addresses the challenges associated with maintaining a living, growing conservation collection. Finally, Appendix 4, by Kevin James, is a summary of some of the major organizations around the world that are engaged in ex situ conservation, most of which contributed to the symposium on which this volume is built. A major departure from the previous volumes is that although the CPC is based and operates in the United States, this volume explicitly takes a more global view. The rich tapestry of plant life is unraveling not just in the United States but around the globe. Indeed, many of the problems of biodiversity loss are greater elsewhere than they are in the United States, in places that generally have fewer economic resources available to address them. Ex situ plant conservation is not a single monolithic method but a diverse family of techniques that can be applied in many different ways to many different situations. A major challenge for us all is to take an expansive enough view so that humankind can successfully bridge the disparity between where the greatest needs are found and where the most resources are held. To succeed, we need to bring to bear all available tools. Ex situ resources are an essential part of integrated conservation strategies that seek to conserve biodiversity in the wild. We must all think and act in ways that benefit the planet as a whole. We, and our descendants, all depend on
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healthy ecosystems. How well we conserve the earth’s biota today will affect the quality of life for humanity for all time. References CPC (Center for Plant Conservation). 1991. Genetic sampling guidelines for conservation collections of endangered plants. Pages 225–238 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Falk, D. A. 1987. Integrated conservation strategies for endangered plants. Natural Areas Journal 7:118–123. Falk, D. A. 1990. Integrated strategies for conserving plant genetic diversity. Annals of the Missouri Botanical Garden 77:38–47. Falk, D. A., and K. E. Holsinger (eds.). 1991. Genetics and Conservation of Rare Plants. New York: Oxford University Press. Falk, D. A., C. I. Millar, and M. Olwell (eds.). 1996. Restoring Diversity: Strategies for the Reintroduction of Endangered Plants. Washington, DC: Island Press.
acknowledgments
We extend our profound thanks to the many people who gave their time and creativity to this project. This volume, and the conference on which it was based, were possible because so many dedicated people contributed their talent and effort to the cause. We would especially like to thank the Center for Plant Conservation (CPC). The network of institutions that make up the CPC, and its National Office, have promoted and supported ex situ plant conservation since 1984. Through the research and experience of individuals at CPC gardens, the practice of ex situ plant conservation continues to be improved and refined. The CPC family, including executive director Kathryn Kennedy, the network’s conservation officers, and the scientific advisory council, helped shape the conference and this volume in so many ways. We thank the Chicago Botanic Garden and its president, Barbara Carr, for hosting the “Strategies for Survival: Ex Situ Plant Conservation” conference in 1999. Barbara taught us, “If you’re going to do something, you should do it right,” and generously made the facilities and resources of the garden available to us. The hard work of the garden’s conservation science staff (Pati Vitt, Susanne Masi, and Justin Epting) and the education and events staffs (Candice Shoemaker, Linda Jones, Holly Estal, Ed Valauskas, and Suzanne Boué) and many garden volunteers kept the symposium running smoothly. Anukriti Sud of Bloom, Inc., provided a Web-based discussion forum for participants. The conference was sponsored by the Chicago Botanic Garden; Berry Botanic Garden; Royal Botanic Gardens, Kew; Center for Plant Conservation; and Botanic Gardens Conservation International. Financial support was provided by the Janet Meakin Poor xxi
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research symposium endowment of the Chicago Botanic Garden, the U.S. Fish and Wildlife Service, the U.S. Environmental Protection Agency, the U.S. Department of Agriculture Forest Service Midewin National Tallgrass Prairie, Chicago Wilderness, the Lincolnshire Garden Club, and the Baird Foundation. A big thank you goes to all of the authors who contributed chapters to this volume for their excellent work and their patience as we assembled it. We thank reviewers of the conference program, book prospectus, chapters, and appendixes, including Tim Bell, Paulette Bierzychudek, Marlin Bowles, Bill Brumback, Vickie Caraway, Carol Dawson, David DeKing, Kingsley Dixon, Ehsan Dulloo, Christopher Dunn, Florent Engelmann, Holly Forbes, Elizabeth Friar, David Galbraith, David Given, Vernon Heywood, Kent Holsinger, Peter Wyse Jackson, Tom Kaye, Kathryn Kennedy, Sawsan Khuri, Charlie Lamoureux, Joyce Maschinski, Susanne Masi, Kimberlie McCue, Linda McMahan, Jeanette Mill, Suzanne Nelson, Peggy Olwell, Brian Parsons, Hugh Pritchard, Robin Probert, George Rabb, Johnny Randall, Andrea Raven, Peter Raven, Kathy Rice, Bill Rottschaefer, Barbara Schaal, Roger Smith, Pati Vitt, Stuart Wagenius, Michael Wall, Christina Walters, Louise Egerton Warburton, Peter White, Dieter Wilken, Diana Wolf, and Mary Yurlina. Your insights and ideas greatly improved this work. We also offer our appreciation to the institutions that provided support and allowed us time to complete the book: Berry Botanic Garden, Chicago Botanic Garden, Fairchild Tropical Garden, National Tropical Botanical Garden, and Royal Botanic Gardens, Kew. We have been thoroughly impressed with the Island Press staff who have helped us with this project from start to finish. We especially acknowledge the support of Barbara Dean and Barbara Youngblood for their thoughtful comments and their help in shaping this volume. Carol Anne Peschke and Cecilia González provided valuable assistance in copyediting and production, for which we thank them. Finally, we thank family, friends, and especially each other for encouragement and support on the long road from initial concept to publication. It has been a pleasure to take this journey together.
introduction Ghillean T. Prance
The late Stephen J. Gould had agreed to write this essay, but unfortunately this great zoologist, evolutionary biologist, and interpreter of science died before he had time to complete it. I am sorry not to have had the opportunity to read another essay by him. Although I cannot possibly emulate Gould’s wonderful style of writing, I find myself wondering what an evolutionary biologist and zoologist would have written about ex situ plant conservation. The greatest drawback to ex situ conservation is that in most cases it halts or distorts the natural process of evolution. Evolution was Stephen Gould’s particular specialty, and he wrote many articles about the detailed interactions between organisms to illustrate this process. The process of evolution is modified when we are forced to store plant species in seed banks or even grow them in botanic gardens away from their natural range and specific ecosystem. Perhaps we are causing a gap in the evolutionary process that will eventually be regarded as another period of dormancy to support the theory of punctuated equilibrium that Gould (2002) was so instrumental in developing. In any case, it is an honor to have the opportunity to remember here this great theoretical biologist and defender and interpreter of the natural world. Most conservationists readily admit that in situ conservation, the conservation of habitats and ecosystems with their constituent populations of species, is the highest priority. This approach is certainly supported by the Convention on Biological Diversity (CBD), of which Article IX states that parties shall use ex situ techniques “as far as possible and as appropriate, and predominantly for the purpose of complementing in situ methods” (Glowka et al. 1994, p. 52). The emphasis of the convention is definitely xxiii
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to recommend in situ rather than ex situ conservation wherever possible. If this is the case, why do we need a book about the science of ex situ conservation? The sad truth is that ex situ conservation is becoming ever more important as a tool to help maintain biodiversity. Human-caused habitat loss and degradation and invasive species are accelerating the loss of species (Tilman and Lehman 2001). In addition, many habitats are vulnerable to alteration through human-caused climate change, and these changes are occurring at a pace that is beyond the dispersal ability of many plant species (Crumpacker et al. 2001). Therefore, a book that helps to develop ex situ conservation as a practical science is of vital importance to conservation. As I write this on the beautiful island of Kauai in Hawaii, sitting in the National Tropical Botanical Garden, I am surrounded by examples of the practical challenges of ex situ conservation. The majority of the threatened rare species in Hawaii exist only as a very small population of questionable viability. It is to be hoped that some of these, such as Cyanea pinnatifida (Cham.) F. Wimmer (Campanulaceae), which was reduced to a single individual, can be rescued. The success of ex situ conservation has already been demonstrated with the nene, or Hawaiian goose (Branta sandivicensis). After captive management and reintroduction this bird has made a remarkable recovery from what many skeptics thought was an impossibly small population. Today I do not have to visit the Wildfowl and Wetlands Trust in England to see nene, but each time I visit Koke’e or Kilauea on Kauai I see these beautiful birds that have been rescued from the path of extinction. In the case of Cyanea pinnatifida (Campanulaceae) there is some hope because hundreds of individuals have been propagated from the single founder by the Lyon Arboretum in Honolulu. The case of Hibiscadelphus woodii (Malvaceae) Lorence & W. L. Wagner is less hopeful. Of the four wild individuals discovered, which were accessible only through the use of climbing ropes, only one remains alive, and no one has been able to propagate it. Similarly, Kanaloa kahoolawensis (Fabaceae) Lorence & K. R. Wood had a wild population of only two individuals when it was discovered in 1992, only one of which survives, but there are now two individuals in cultivation. Hawaii, like so many devastated oceanic islands, is the ultimate challenge to species conservationists, whether proponents of in situ or ex situ methods. Growing near to me are some pots full of the attractive pachycaul member of the Campanulaceae, Brighamia insignis (Campanulaceae) A. Gray.
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The creamy yellow flowers have a long, narrow corolla tube typical of hawkmoth-pollinated flowers. Unfortunately, it is thought that the sphingid moth that pollinates this plant is extinct or near extinction. No other moth is able to pollinate this species. Artificial pollination can be done easily here in the botanic garden, but in the wild it often entails rappelling down cliff faces. Many of the threatened rare plants that we are trying to save have specialized pollination mechanisms that cannot work with generalist pollinators. Even if these pollinators are still extant, one cannot preserve insects or hummingbirds in cold storage, as one can with seeds. It is often easy enough to keep a plant species alive but much harder to maintain the interactions it needs for pollination, seed dispersal, and, indeed, other mutualistic relationships with animals. Without these processes, evolution is halted. Perhaps a much closer collaboration is needed between workers in ex situ conservation of animals and of plants. All these examples from Hawaii emphasize the important role of botanic gardens in conservation. I have been on the staff of various botanic gardens for almost 40 years. During that time the environmental and political conditions for in situ conservation have deteriorated rapidly, and the number of species threatened with extinction has increased dramatically. Many botanic gardens have responded to this challenge and have established added conservation programs. We also have support from organizations such as the Center for Plant Conservation (CPC), the International Union for the Conservation of Nature (IUCN) Species Survival Commission (SSC), the International Plant Genetic Resources Institute (IPGRI), and Botanic Gardens Conservation International (BGCI). These entities are helping many botanic gardens and other ex situ practitioners improve their conservation programs, such as BGCI’s global agenda for botanic gardens (BGCI 2001). A botanic garden that does not emphasize plant conservation in its mission program, whether in education or in the ex situ conservation of species or habitats, is not adequately responding to the challenges of today’s world. As components of both agricultural and wild landscapes, plants are fundamental to human well-being. Ex situ conservation cannot afford to be only a process of collection and storage; the release of material for repatriation and reintroduction provides the ultimate service to the clients of ex situ conservation, be they protected area managers, private landowners, or rural communities (Maunder 1992; Sperling 2001). The science of ex situ conservation preserves not only wild species but also the huge number of
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varieties of domesticated species that humans have developed over the past 10,000 years, since the beginnings of agriculture. The importance of conserving landraces of crop species has been demonstrated recently in Afghanistan, where many of the locally adapted cereal varieties have been destroyed by drought and warfare. On top of the recent conflict, 2001 was the third year in a row in which the rains failed. This caused the loss of the majority of the seeds on which the farmers depended. Although the weather improved, the farmers still lacked a basic agricultural need: seeds of their traditional crops. Worse still, the National Gene Bank of Afghanistan was destroyed in 1992. Over many years Afghani farmers had selected varieties of their crops of wheat, chickpeas, barley, lentils, and fava beans that were appropriate to local conditions and taste; these included strains of crops that would grow in some of the most unfavorable places for agriculture. Fortunately, in the 1970s Geoffrey Hawtin, who is now director-general of the IPGRI in Rome, traveled throughout Afghanistan to collect seed for use by crop breeders around the world. Hawtin’s visit was just before the Soviet occupation that would have stopped such a venture. Some of Hawtin’s seeds were deposited at freezing temperatures in the seed bank of the International Center for Agricultural Research in the Dry Areas (ICARDA). Many of these seeds are being returned to their country of origin to help rebuild agriculture there. Although some of ICARDA’s improved varieties of wheat are also helping Afghanistan, there are many places where specially selected local landraces and varieties will do better. The small quantities of seeds of these varieties will be crucial to restoring agriculture in Afghanistan. Many of the most useful plants to humanity are the ones that are most threatened with extinction because of overuse. This is particularly true of medicinal plants. More than 80 percent of the developing world still relies on traditional medicines, mainly from plants, for their primary healthcare (Farnsworth and Soejarto 1991). Even in the developed world, the use of plant-based medicinal systems such as Chinese and Ayurvedic medicine is increasing. As a result, some of these important healing plant species are overcollected. The Royal Botanic Gardens, Kew, and Guy’s Hospital in London have set up an authentication center for Chinese medicines to combat the increasing trend of substitution of fake compounds because where the true medicinal plants are becoming scarce, other plants are often used. The ex situ cultivation of some of these plants can reduce the pressure on wild populations and improve the quality of life for many communities.
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A program that does this is Living Pharmacies (Farmácia Viva) of Brazil. It began in the city of Fortaleza, in northeastern Brazil, where a chemist, Professor J. Mattos, and a medical doctor, Dr. Adalberto, began growing plants and preparing medicines from them to treat people in one of the large slums of the city. They employ formerly abandoned street children to cultivate the plants in return for an education and food. The medicinal plant gardens that have been set up in several places in Brazil are reducing pressure on the species in the wild and performing a major social function by providing affordable medicines to local people. I have also seen medicinal plant gardens in India that are producing the ingredients of Ayurvedic medicines while reducing the need to collect wild-sourced material. We need to have a broad concept of ex situ conservation and to include projects such as these in our thinking about how to protect threatened species and resources, especially when we are dealing with species of economic use in poor areas. In recent years ex situ conservation has become a much more precise science with a wonderful array of tools. Foremost among these are molecular techniques that enable us to assess and monitor the genetic variation within populations. This is absolutely critical when we are dealing with small populations, whether in situ or ex situ. When only a few individuals exist, it is vital to make genetically appropriate crossings to obtain healthy progeny and to capture what little genetic diversity is left. Molecular methods are also useful in monitoring the purity of a species and ensuring that hybridization has not taken place. Hybridization is a risk when plants are grown in botanic gardens, and often not enough care is taken to avoid it. Tissue culture, which we have used for some years, is an invaluable tool for propagating rare species and obtaining disease-free lines. Seed storage methods have greatly improved in the last few decades, and there has been much research on dormancy breaking and recalcitrance. Wherever possible, ex situ conservation should be regarded as a temporary method, and practitioners should always be looking for ways to restore species to their natural habitats. In order for this to happen it is essential that the science of ex situ conservation not be isolated from that of in situ conservation and that an integrated approach be adopted (sensu Falk 1987). Many botanic gardens today have areas of natural vegetation within their boundaries or in adjunct campuses. This is ideal because it involves them in the practice of in situ conservation, and as a result their ex situ work usually benefits as well. The prime example of this is seen in
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the South African network of botanic gardens, which has a garden with large natural areas situated in each major ecosystem. I am pleased to be working with the Eden Project in Cornwall, England, because it is performing an important function for ex situ conservation: transmitting information to the public. Eden exists to promote the importance of plants to people and the sustainable use of all plants. This successful project, which received more than 2 million visitors in its first year of operation, is bringing a strong message of conservation to its visitors. To promote such a message, it is also necessary to practice conservation and sustainable management techniques. In the 5-acre rainforest biome you will find such threatened plants as Trochetiopsis ebenus (R. Brown ex Aiton f.) W. Marais (Sterculiaceae), the Saint Helena ebony, which was reduced to only two individuals in the wild, and Impatiens gordonii Horne (Balsaminaceae) from the Seychelles, of which only a few individuals remained. The Eden Project has developed partnership agreements with institutions in those island territories and in various other places from which it is exhibiting plants. In addition to multiplying material for reintroduction to the wild, we hope to bring the Impatiens species to the horticultural market to benefit conservation in the Seychelles. Visitors to Eden learn about the threats to these and other threatened rare plants and about what is being done to rescue them from extinction. Eden is both practicing and exhibiting ex situ conservation. There is still not enough of the latter in botanic gardens and reserves, and conservation would benefit greatly if more understanding could be instilled into the general public through the display and interpretation of the rare plants they grow. This book recognizes the limitations of ex situ conservation while urging us not to undervalue it. That ex situ conservation is vitally important and has prevented the extinction of many species of plants and animals is undeniable. Starting with Franklinia altamata Marshall, last seen in the wild in 1803, gardens have enabled the survival of many species that have become extinct in the wild. A recent example is the beautiful crocus-like Tecophilaea cyanocrocus Leyb. from Chile, a species that is quite common in horticulture but extinct in the wild (Maunder et al. 2001). There is an important niche for ex situ conservation, and I hope that this volume, the third CPC book on plant conservation, will promote it as a tool to support both species and habitat conservation.
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References BGCI (Botanic Gardens Conservation International). 2001. International Agenda for Botanic Gardens in Conservation. Kew, UK: Botanic Gardens Conservation International. Crumpacker, D. W., E. O. Box, and E. D. Hardin. 2001. Implications of climatic warming for conservation of native trees and shrubs in Florida. Conservation Biology 15(4):1008–1020. Falk, D. A. 1987. Integrated conservation strategies for endangered plants. Natural Areas Journal 7:118–123. Farnsworth, N. T., and D. D. Soejarto. 1991. Global importance of medicinal plants. Pages 25–51 in O. Akerele, V. H. Heywood, and H. Synge (eds.), The Conservation of Medicinal Plants. Cambridge, UK: Cambridge University Press. Glowka L., F. Burhenne-Guilman, H. Synge, J. A. McNeely, and L. Gündling. 1994. A Guide to the Convention on Biological Diversity. Environment Policy and Law Paper no. 30. Gland, Switzerland: IUCN. Gould, S. J. 2002. Punctuated equilibrium’s threefold history. Pages 1006–1021 in S. J. Gould, The Structure of Evolutionary Theory. Cambridge, MA: Harvard University Press. Maunder, M. 1992. Plant reintroduction: an overview. Biodiversity and Conservation 1:21–62. Maunder, M., R. S. Cowan, P. Stranc, and M. F. Fay. 2001. The genetic status and conservation management of two cultivated bulb species extinct in the wild: Tecophilaea cyanocrocus (Chile) and Tulipa sprengeri (Turkey). Conservation Genetics 2:193–201. Sperling, L. 2001. The effect of the civil war on Rwanda’s bean seed systems and the unusual bean diversity. Biodiversity and Conservation 10:989–1009. Tilman, D., and C. L. Lehman. 2001. Human caused environmental change: impacts on plant diversity and evolution. Proceedings of the National Academy of Sciences of the United States of America 98(10):5433–5440.
part one
The Scope and Potential of Ex Situ Plant Conservation Early perceptions of ex situ (off-site) plant conservation as a largely irrelevant novelty or possibly even a well-meaning but counterproductive distraction are giving way to a growing awareness that properly managed off-site collections can make the critical difference between extinction and survival. The diverse tools of ex situ plant conservation are a means to an end— survival in the wild—and a vital part of larger integrated conservation efforts. Part I reflects the remarkable advances that ex situ plant conservation has made. Beginning with application to a small number of unusually threatened species and practiced as a standalone approach, and serving by default as a management cul-de-sac (see Chapter 1, this volume), ex situ plant conservation is increasingly being used to support the integrated conservation of regional plant diversity (Chapters 3 and 4, this volume). Over the almost 500-year history of the modern botanic garden, the classic venue for ex situ plant conservation, curatorial principles and professional codes evolved slowly until the last 40 years, when we have seen a dramatic revolution in both professional ethics and the application of conservation science (Chapter 1). As outlined in Chapters 1 and 3–5, this revolution has been driven in part by an internal recognition that ex situ conservation is a duty for botanic gardens and an external expectation by both the public and conservation agencies. Set against the backdrop of an increasing appreciation of the sheer magnitude and increasing rate of species loss, the Endangered Species Act (ESA) in the United States and the ratification of the international Convention on Biological Diversity (CBD) can be viewed as two of the greatest external stimuli for ex situ plant conservation. The legal requirement of the ESA to recover threatened species in the United States provided impetus to try new approaches. More recently, and through the CBD’s national biodiversity strategies, has come explicit recognition that ex situ conservation is a legitimate and sometimes essential tool for species conservation and a
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valuable support to in situ conservation efforts. In turn, both of these extraordinary instruments are derived from decades of concern about the environment and from the intellectual and scientific frameworks developed by international agencies such as the World Conservation Union (IUCN), the International Plant Genetic Resources Institute (IPGRI), and Botanic Gardens Conservation International (BGCI). Whereas the international conferences of the 1970s by IUCN and others spurred the development of ex situ conservation in recent decades, the work of Vavilov and others in the early 1900s laid the strategic and scientific framework for ex situ plant conservation. It is on these foundations that national networks such as the Center for Plant Conservation (CPC; Chapter 4, this volume) and regional government agencies such as the State of Western Australia (Chapter 3, this volume) have advanced the field. The experience of both the CPC and Western Australia illustrates a growing mode of professional practice, an essentially collaborative ethic based on delivering services to conservation agencies with the objective of securing wild populations and holding viable insurance collections as a backup to wild stocks. The ethical context of ex situ conservation, reviewed by Rolston in Chapter 2, is a dynamic and sometimes troubling view, but one that provides the intellectual and moral framework for much of plant conservation. Rolston explores the philosophical underpinnings of the commonsense recognition that biodiversity loss is to be avoided and that wild populations are inherently more valuable, and informative, than cultivated representations. Based on lessons learned in the zoo community, in Chapter 5 Stanley Price and colleagues advance the ethical debate beyond the immediate concerns about species conservation and show how ex situ efforts and institutions can be used to leverage support for habitat and ecosystem conservation, both locally and globally. These and other advances help to move us beyond the “put a plant in a pot and the species is saved” stereotype toward a realization that ex situ activities are critical to integrated plant conservation. Success ultimately will be measured not just by the number of taxa stored safely but, more importantly, by how well ex situ efforts contribute to the overall effort to maintain biodiversity in the wild.
Chapter 1
Ex Situ Methods: A Vital but Underused Set of Conservation Resources Mike Maunder, Kayri Havens, Edward O. Guerrant Jr., and Donald A. Falk
Botanic gardens and other ex situ facilities such as seed banks are among the most extensive yet underused plant conservation resources in the world. For them to make a truly meaningful difference in how much plant diversity survives into the next century, ex situ plant conservation providers need to not only use the most effective and efficient means possible, but also increase their institutional capacity. In a sobering global review of the threats to biological diversity, Myers et al. (2000: 853) found that the “number of species threatened with extinction far outstrips available conservation resources, and the situation looks set to become rapidly worse.” This statement summarizes the challenge facing ex situ conservation at a time when the absolute need for in situ conservation has never been greater and the threats facing species diversity are increasing in type, severity, and scale. The world’s botanic gardens and other ex situ facilities, such as seed banks, are among the most concentrated sites of species richness on the planet, in effect artificial centers of species diversity. The world’s 1,800 botanic gardens hold an estimated 2.5 million accessions of growing plants representing about 80,000 species (Wyse Jackson 2001). These vast collections have been accumulated over many decades and represent a huge investment in human resources and infrastructure. This book reviews the effective role of ex situ collections and assesses the values and limitations of ex situ plant conservation techniques. The vast majority of ex situ samples, even those intended for conservation, have been collected on an ad hoc basis because they may be needed in the future for some unspecified purpose by an unspecified client. In addition, these collections are heavily skewed toward the cultivation of 3
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small and non–genetically representative samples of horticulturally amenable taxa and often suffer from poor documentation (Maunder et al. 2001b, 2001c). The majority of threatened species held in botanic garden collections are not specifically managed for conservation purposes and are characterized by a number of shared genetic and demographic characteristics (Box 1.1). However, the composition, status, and management of collections are rapidly improving as more facilities adopt conservation responsibilities. The recurring theme of this book is clearly outlined by Stanley Price et al. in Chapter 5: Where and how can ex situ investment make the most difference to in situ conservation? Plant conservation facilities operate under the premise that they contribute conservation services to a variety of different clients. The primary role, retaining samples of wild plant diversity under artificial and accessible conditions, has been ratified in a number of professional guidelines (BGCI 2001) and international policy documents (IUCN/UNEP/WWF 1980, 1991; Glowka et al. 1994). However, these services are provided by a range of institutions and facilities of diverse historical heritage that share few common standards or protocols for the management, documentation, and display of threatened plant material. The majority of ex situ facilities were developed and still serve as facilities for growing and displaying token or at least genetically nonrepresentative samples of taxonomic diversity. The challenge is to use the most effective tools and processes and to serve these agreed roles of maintaining offsite collections and making them available for restoration and other conservation purposes. The number of ex situ conservation facilities has increased dramatically in recent years (Wyse Jackson 2001), and they have become increasingly integrated under national and regional conservation initiatives. Nevertheless, many authorities hesitate to use them as a fundamental and effective component of plant conservation. This reluctance may originate from a number of perceptions. First, that ex situ conservation may undermine the integrity of, and need for, in situ conservation by devaluing wild populations and habitats. Second, it may reflect a lack of professional confidence in the technical ability of ex situ facilities to hold genetically diverse samples of threatened plant germplasm over extended periods of time. Much of the concern probably is based on a lack of understanding of what ex situ options exist and what their strengths and limitations are. For instance, storing seed is very different, in terms of both financial costs and biologi-
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box 1.1 Characteristics of Threatened Plant Populations in Botanic Gardens • Populations are small and often derived from a small number of closely
related founder individuals. • The cultivated stocks are subject to fluctuating population size as a result
of changing horticultural fashions and episodic mortality events. • Often little or no associated ecological or biological information is avail-
able to guide ex situ managers in cultivating and managing the stocks. • There is little information on the history of the taxa in cultivation and
often no satisfactory horticultural protocols. • Individuals are scattered through a number of collections with varying
horticultural and curatorial capacity and hence differing patterns of regeneration and mortality. • Individuals are susceptible to artificial selection, genetic drift, inbreeding, and hybridization with congenerics. • Persistence in collections is highest for horticulturally amenable taxa and particularly for taxa with display or commercial value. Based on Maunder and Culham (1997).
cal risks, from maintaining a cultivated collection. Third, ex situ collections can be viewed as potential conservation liabilities, a source of new invasives and pathogens that can affect wild populations and habitats (Reichard and White 2001). Ex situ conservation at its crudest may temporarily hold token samples of wild plant diversity. At best it can play a critical role as one component of an integrated conservation response supporting a primary objective: the retention and restoration of wild plant diversity. However, to achieve this objective, improved working practices and facilities are needed. We contend that it is because ex situ tools are not widely understood that they are undervalued and therefore underused. Understanding and effectively communicating the relative conservation roles, values, opportunities, and challenges faced by seed storage and growing collections may be among the biggest challenges practitioners face. A major purpose of this volume is to explore the value, limits, and range of available ex situ tools.
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The Plant Diversity Crisis The old World Conservation Union (IUCN) survey of plant species conservation status indicates that about 33,400 plant species are threatened with extinction (Walter and Gillett 1998), or about 10 percent of the world’s known 250,000–300,000 plant species. This IUCN survey records 380 plant extinctions (Walter and Gillett 1998), less than 1 percent of the recorded species of vascular plants. The plant extinctions recorded by the IUCN and World Conservation Monitoring Center (WCMC) reflect, in part, the geographic distribution of botanical knowledge and monitoring rather than actual rates of species loss. For the largest part of the planet there is no clear consensus on the rate of species and population loss, but this is improving as more IUCN Red Lists are undertaken. For example, in a review of recorded extinctions, rates of habitat conversion, and distribution of restricted endemic plant species for tropical Latin America, Koopowitz et al. (1994) produced estimates of recent extinctions that far exceed those the WCMC and IUCN record. This discrepancy is particularly notable for Brazil, where WCMC and IUCN recorded only five extinctions, whereas Koopowitz et al. estimate a loss since 1950 of about 2,200 species. There is an acute need to act decisively, creatively, and effectively. Extinction rates for both species and populations are increasing as a result of human changes to habitats (Hannah et al. 1994; Hughes et al. 1997). This trend is accelerating as surviving wild areas become increasingly degraded through isolation, fragmentation, competition from invasive species, and climate change (Sala et al. 2000). The expected result, particularly in the endemic-rich hotspots (sensu Myers et al. 2000), will be many more plant extinctions. The dearth of field survey work and the rapidity of habitat loss, particularly in the tropics, mean that many plant extinctions are likely to be identified only in retrospect, if at all. This raises the question, How should facilities, especially those in the high-diversity regions, most effectively allocate their limited resources? Should they focus a large proportion of their limited resources on managing a small number of threatened plant species, perhaps selected through an imperfect understanding of local conservation priorities? Or should they also use available resources for promoting and supporting the conservation of important habitat areas, such as recognized centers of plant diversity (Maunder et al. 2002; Chapter 5, this volume)? In addition to measuring yield from ex situ investment through
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increased taxonomic representation in cultivation or seed banks (e.g., scoring collections against national red lists), facilities could also score against quantitative assessments of genetic representation, contributions to implemented recovery plans, and the conservation of important plant habitats.
The Evolution of Ex Situ Plant Conservation Botanic gardens, as scientifically organized plant collections, were originally initiated as repositories serving academic study, predominantly medical and theological (Prest 1981). In a time still dominated by an essentialist worldview dating back to Plato, curation was driven by the desire to accumulate typological specimens. They subsequently developed as resources for both colonial agriculture and taxonomic science (Osborne 1995; McCracken 1997). Only in the last 50 years of a 500-year postRenaissance history have these collections and facilities been used to counter a human-mediated decline in species diversity. A specific ex situ conservation role could arguably develop only after the concept and reality of extinction, and in particular the role of humans in accelerating extinction rates, were first recognized and then accepted by the scientific community. Another foundational scientific advance that underlies current conservation thinking was the shift from an essentialistic, or typological, specimen-based approach to a populational view of the biological world (Mayr 1982). What it means to have a representative sample has profoundly changed. Rather than viewing individual differences as corrupt and imperfect manifestations of a Platonic ideal, biological variation has come to be appreciated as the raw material upon which natural selection acts and adaptive evolution depends. In other words, conservation of wild species as both a scientific and an ethical goal is a consequence of the revolution in late-eighteenth- and nineteenth-century scientific thought. In the late nineteenth and early twentieth centuries two tree species were assumed to have become extinct in the wild and to have survived only in cultivation, namely Ginkgo biloba (Ginkgoaceae) from China (Wilson 1919) and Amherstia nobilis (Fabaceae) from Myanmar (Blatter and Millard 1993). These species, along with the U.S. endemic tree Franklinia alatamaha (Theaceae), appear to have been treated as isolated novelties and did not prompt a broad conservation response from the botanic garden community. The development of ex situ conservation reflected a cultural and scientific transition for plant collections from a focus on accumulating
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the scope and potential of ex situ plant conservation
typological specimens as a curatorial goal to the adoption of population genetics as a working tool to conserve threatened species. Contemporary concerns about the loss of plant diversity and the need for effective ex situ storage can be traced, at least in part, to the groundbreaking work of Vavilov (1926, 1949–1950), who first recognized the value of crop landraces and wild relatives in supporting agriculture. The specific role of botanic gardens in supporting wild plant diversity was explored by Cugnac (1953), who outlined the need for specific ex situ conservation facilities working in close association with protected areas, the jardin conservatoire. The ark paradigm, the idea that ex situ facilities would hold stocks of threatened species during a period of habitat degradation (the “demographic winter” sensu Soulé et al. 1986), was established as a working objective by botanic gardens in the 1970s. For instance, Heslop-Harrison (1974: 31) saw cultivation as a necessary preliminary in which “the ultimate objective is to restore the devastation of former periods and rehabilitate an ecosystem.” This paradigm is manifest in the first IUCN Plant Red Data Book (Lucas and Synge 1978: 305). For instance, the entry for Dracaena ombet (Liliaceae sensu lato) stated, “It seems too late for such a proposal [in situ conservation] to be worthwhile. Great efforts must now be made to bring the ombet into cultivation and maintain it safely in the botanic gardens of the world.” An equally pessimistic view is expressed by Lavranos (1974: 23), who recognized “the utter futility of any thoughts on conservation in situ in such environments [NE Africa].” Lavranos proposed that “the only way to save threatened species is to get them into cultivation” out of the range country, with the hope that “we may see them reintroduced into the original biotope—if that still exists” (Lavranos 1974: 23). The ark paradigm, with ex situ conservation as an open-ended storage responsibility until a change in human demography and consciousness allowed species a wild future, is being replaced by the recognition that ex situ conservation can and should work in partnership with the management of extant wild populations. This reflects the perspective that ex situ conservation provides a service that answers the practical needs of the population manager and in situ agencies and is not a competing alternative to in situ conservation (Given 1987). The Center for Plant Conservation (CPC) in the United States is a pioneer of a client-based model for ex situ conservation (Thibodeau and Falk 1987; Kennedy 2002). This shift in emphasis toward integrated strategies
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sensu Falk (1987, 1990) has led to the recognition that ex situ management can and should play an important role in species conservation in the wild (Falk et al. 1996). The activities and scientific approach of the CPC have generated productive debate on the conservation role of botanic gardens (Feldman 1996; Robertson 1996; White 1996). This has resulted in the development of working collaborations with protected area authorities and government agencies (McMahan 1995; Cotterman and Jones-Roe 1996) and the adoption of population genetics as a guiding tool for botanic garden conservation activities (CPC 1991; McMahan and Guerrant 1991; Mistretta 1994). The establishment of regional plant conservation strategic alliances reflects this collaborative approach. Examples encompass a variety of scales, from continental or national, as in the Australian Network for Plant Conservation (Mill 2002) and the Plant Conservation Alliance of the United States (Olwell, pers. comm., 2003), to regional plant conservation plans, as in Andalucía, Spain (Hernández-Bermejo and Clemente-Muñoz 1994), and the New England Plant Conservation Program (New England Wildflower Society 1992) in the United States. Thus, even traditional habitat-based conservation strategies are moving from hands-off approaches to more active and interventionist methods. This trend toward recovery and reintroduction creates a strategic opportunity for ex situ institutions to serve as active partners in species-based research and recovery projects (Falk and Olwell 1992; Falk et al. 1996; Guerrant and Pavlik 1997). The value of ex situ conservation has been increasingly acknowledged in international treaties and legislation (Warren 1995). The Convention on Biological Diversity (CBD; Glowka et al. 1994) provides a major opportunity for ex situ facilities to establish themselves as valued resources serving stated national needs (BGCI 2001). However, it also brings ex situ activities into critical review, particularly with regard to the ownership and distribution of plant material. The CBD recognizes the value of ex situ conservation (Box 1.2), with an emphasis on undertaking these activities “preferably in the country of origin” and as a support to the “recovery and rehabilitation of threatened species and for their reintroduction into their natural habitats” (Glowka et al. 1994: 52). Practitioners are increasingly recognizing the need to be responsive to national priorities for biodiversity. This is encouraging practitioners to integrate plant conservation and biodiversity issues with broader agendas so that decision makers recognize the congruency of agendas, for instance in the areas of sustainable development, habitat restoration, healthcare, ecosystem services, and education.
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the scope and potential of ex situ plant conservation
box 1.2 Convention on Biological Diversity Specific Recommendations on Species-Level Conservation Article 8. In Situ Conservation . . . (d) Promote the protection of ecosystems, natural habitats and the maintenance of viable populations of species in natural surrounding. . . . (f) Rehabilitate and restore degraded ecosystems and promote the recovery of threatened species, inter alia, through the development and implementation of plans or other management strategies. . . . (h) Prevent the introduction of, control or eradicate those alien species which threaten ecosystems, habitats or species. . . . (k) Develop or maintain necessary legislation and/or other regulatory provisions for the protection of threatened species and populations. . . . Article 9. Ex Situ Conservation (a) Adopt measures for the ex situ conservation of components of biological diversity, preferably in the country of origin of such components; (b) Establish and maintain facilities for ex situ conservation of and research on plants, preferably in the country of origin of genetic resources; (c) Adopt measures for the recovery and rehabilitation of threatened species and for their reintroduction into their natural habitats under appropriate conditions; (d) Regulate and manage collection of biological resources from natural habitats for ex situ conservation purposes so as to not threaten ecosystems and in situ populations of species, except where special temporary ex situ measures are required under subparagraph (c) above. Glowka et al. (1994).
Tools and Facilities for Ex Situ Conservation The world’s ex situ plant resources encompass everything from the traditional garden plots of the tropics to intensive allotments, farms, and gardens to seed banks and botanic gardens. The world’s botanic gardens are estimated to cultivate some 4 million accessions, representing 80,000 taxa. The majority of these collections are maintained as small numbers of living plants in mixed collections serving a wide range of purposes. There are approximately 2,000 botanic gardens in 148 countries, but more than 40 percent of these botanic gardens are concentrated in Western Europe and
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North America. These collections cultivate a biased representation of the world’s botanical diversity, with more comprehensive representation at the generic level for attractive and horticulturally amenable plant groups such as the conifers, palms, cacti, bromeliads, and orchids. The levels of intraspecific genetic diversity are low because most species are represented by few individuals, often from a limited number of founders. These collections contain specimens of threatened species including some now extinct in the wild (sensu IUCN), such as Sophora toromiro (Fabaceae) (Maunder et al. 2000). Large collections of horticultural plants, mostly cultivars, are maintained by commercial nurseries and amateur horticulturists. Some countries have established national collections of garden plants; such networks exist in Australia, France, and the United Kingdom. For instance, the National Council for the Conservation of Plants and Gardens (NCCPG) in the United Kingdom coordinates more than 600 collections, maintained by professional and amateur horticulturists, containing 13,000 species and 39,000 cultivars. The evolution of the CPC exemplifies a broader shift in the role and value of the ex situ component of plant conservation strategies more generally. As initially conceived, the CPC envisioned growing collections of threatened plants. Over time, the emphasis has shifted dramatically toward the use of dormant seed collections because they are far more cost-effective and efficient means by which to store large genetically representative samples off site for long periods of time. Much of the criticism of and lack of enthusiasm for ex situ conservation is based more on the challenges associated with maintaining living collections rather than with banked seed. As we shall see, the challenges involved with actively growing collections often are much more formidable than those for seed collections (Chapters 12, 16, and 17 and Appendix 3, this volume). Ex situ tools are numerous and vary widely in their costs and benefits (both financial and biological), and in their spheres of application (Figure 1.1; Appendix 3, this volume). Ex situ methods encompass a wide variety of techniques of varying management intensity, capital and labor investment, and potential levels of genetic and demographic modification (Figure 1.1). At one end of the spectrum, propagules can be stored with minimal levels of artificial selection as banked seed or cryogenically stored tissue, with or without associated grow-outs for accession regeneration. Growing plants can be maintained as in vitro cultures, as living collections in pots, in gardens, in field gene banks, or in seminatural environments (i.e., inter-situ conservation). These growing collections can be maintained in specialist facilities with
Figure 1.1 A range of ex situ and in situ plant conservation methods and the relative ongoing effort or marginal resource needs. This figure assumes that facilities exist to perform the task, so the vertical axis does not reflect initial resource investment, which can be large (e.g., for cryopreservation, seed banking, in vitro, or other controlled-environment facilities). For additional information on their best applications, strengths, and limitations, see Appendix 3, Tables A3.1 and A3.2. Cryopreservation: Seeds, pollen, or tissue frozen in liquid nitrogen. Used for the long-term storage of agricultural and horticultural taxa; increasingly used for wild species. Seed banking: Seeds stored in conditions of low moisture and temperature. Routinely used for orthodox seeds of crops and wild species. Tissue culture storage: Somatic tissue stored in vitro under temperature and light conditions controlled for slow growth. Tissue culture propagation: Somatic tissue and seed propagated in vitro, used for the proliferation of clonal plants and controlled seedling production. Cultivation in dedicated conservation facility: Plants cultivated under taxon-specific horticultural regime with aim of cultivating and propagating the threatened species. Specialist cultivation in controlled environment: Plants cultivated under artificial environment, such as tropical species in heated glasshouses in temperate regions. High horticultural investment. Cultivation in mixed display or reference collections: Plants cultivated as part of reference collection under ambient environmental conditions. Majority of holdings in botanic gardens and arboreta held in large collections where the focus is on taxonomic representation or horticultural display.
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minimal levels of artificial sympatry or maintained in large collections with very high levels of artificial sympatry, such as traditional botanic garden display collections. DNA banks have been proposed as a response to biodiversity loss (Benford 1992). Although DNA banks play a vital role in genetic research and engineering, the use of DNA to resurrect extinct species is unlikely to become a realistic option.
Is Ex Situ Plant Conservation Delivering? It is difficult to fully assess the effectiveness of ex situ techniques as conservation tools. However, some components can be assessed against conservation objectives. The global portfolio of ex situ facilities is demonstrating a proven ability to store and cultivate a wide sample of the world’s botanical diversity. At one extreme, botanic gardens and seed banks are holding samples of plant diversity lost in the wild, thereby preventing, or at least delaying, some plant extinctions. For instance, botanic gardens are retaining diversity at three levels: retaining extirpated local provenances of taxa still surviving elsewhere in the national and global range, retaining extirpated national provenances of taxa still surviving elsewhere in the global range, and retaining taxa after extinction of all populations in the wild. Conti et al. (1992) record 15 European taxa extirpated from Italy, with 4 cultivated in Italian botanic gardens. Greuter (1994) records 37
Field gene bank: Open-air, extensive planting to maintain genetic diversity within a species, often used for woody commercial species. Commercial cultivation: Horticultural production of a selected taxon, with focus on production of a profit-generating crop based on biomass or numbers of individuals sold; little emphasis on genetic management apart from retention of selected strains or cultivars. Community garden: Production of plants by community group (village or family) as part of traditional agriculture to produce a used plant product, such as medicinals. Inter situ: Plants cultivated in horticulturally managed near-natural conditions, such as a managed population within restored seminatural vegetation. In situ, horticulturally managed wild populations: Wild plants subject to some degree of species-specific horticultural and demographic management, such as the hand pollination of wild orchid populations. In situ, managed wild populations: Wild plants growing in managed habitat and subject to community-level management, such as burning of grasslands. In situ: Wild plants subject to natural ecological processes.
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the scope and potential of ex situ plant conservation
extinctions from the Mediterranean Basin, at least 4 of which survive in cultivation (Maunder et al. 2001b). Botanic gardens can retain samples of species after extinction in the wild, such as Tulipa sprengeri (Liliaceae) (Maunder et al. 2001a). Globally, at least 100 taxa extinct in the wild survive in ex situ collections (Maunder, unpublished data, 1999). Although current stocks of threatened plants held in ex situ facilities may not always reflect stated national or regional conservation priorities, ex situ facilities have the demonstrated horticultural ability to maintain collections of a diverse taxonomic and ecological composition, albeit often accumulated through ad hoc collection and exchange. The conservation utility of existing botanic garden collections should be questioned; many accessions are held out of the range country in mixed collections with little genetic or demographic management and often inadequate accession data.
Genetic Diversity Traditionally botanic gardens and other ex situ facilities have focused on the accumulation of typological collections of alpha diversity, with cultivated stocks representing few individuals or genotypes. This situation is dramatically improving as botanic gardens place more emphasis on population management and increasingly use seed storage technology, but it is still difficult to maintain genetically representative collections off site even when there is a curatorial plan to do so. Ex situ collections, including commercial nurseries, maintain populations of some globally threatened plant species that far outnumber surviving wild populations; examples include Hyophorbe lagenicaulis (Arecaceae) from Mauritius, Ginkgo biloba from China (Ginkgoaceae), and Echinocactus grusonii (Cactaceae), the golden barrel cactus, from Mexico. It is suspected that some of these cultivated populations may hold important genetic variation that could support the recovery of wild populations, but these cultivated populations are rarely subject to any planned genetic or demographic management. Perhaps more importantly, these species, cultivated in many of the world’s public facilities, provide an untapped opportunity for public education, outreach, and direct support to field programs. For more than 10 years, the CPC network has had guidelines for the collection of “genetically adequate” samples for the nearly 600 taxa in their National Collection of Endangered Plants (Falk and Holsinger 1991). Although these guidelines recommend collecting propagules from many
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plants (10–50) in several populations (up to five), biological, financial, and logistical constraints have limited the level of collection for many taxa. In 1999, fewer than 20 percent of National Collection taxa had documented evidence of genetically adequate ex situ collections based on the 1991 collection guidelines (CPC database). In addition, all ex situ populations are vulnerable to the processes of random genetic drift, genetic erosion or change through drift, selection, and mutation accumulation. Living horticulturally maintained collections are particularly vulnerable to the distorting influences of artificial selection, hybridization, and infection by pathogens (see Chapter 16 and Appendix 3, this volume).
Conservation Activities Only a small proportion of ex situ collections are specifically managed to support national and regional priorities for conservation activities. For instance, in the United States, the CPC, a network of 33 botanic gardens and arboreta, maintains the National Collection of Endangered Plants, comprising 581 species in 2002. For those taxa, 47 reintroduction projects were under way in 2000 (CPC database). A survey of 119 European botanic gardens recorded a total of 345 European plant conservation projects from 49 institutions (Maunder et al. 2001b); however, only 51 projects at 25 institutions focused on 27 priority threatened plant species listed by the Bern Convention. The 51 projects were dominated by propagation, cultivation, and reintroduction projects. The results indicate that the majority of the cultivated Bern accessions in European botanic gardens are not linked to formal species recovery programs. However, about 40 percent of the gardens surveyed (49 of 119) are undertaking plant conservation projects reflecting local conservation priorities. These results indicate the potential for growth in conservation projects in American and European botanic gardens.
Distribution of Ex Situ Facilities In general, the areas of the world with the greatest need for ex situ plant conservation facilities are poorly resourced. Just 10 countries account for more than 70 percent of the world’s botanic gardens; these are largely developed countries (the United States, Germany, France, Australia, Russian Federation, the United Kingdom, Japan, and Italy), with only two developing
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nations (India and China). Europe has the highest concentration of botanic gardens in the world, with 21 Western European countries containing about 535 botanic gardens, including nine national networks covering 11 countries. These facilities are not evenly distributed; the five European nations of Germany, France, the United Kingdom, Italy, and the Netherlands contain a total of 433 botanic gardens (Wyse Jackson 2001), representing almost 81 percent of the Western European botanic gardens and just over 25 percent of the world’s total of 1,700 botanic gardens. In contrast, the Mediterranean Basin, a biodiversity hotspot and IUCN Center of Plant Diversity (Myers et al. 2000; Davis et al. 1994) containing the majority of Europe’s endemic and threatened plant species, is poorly resourced relative to the botanically less diverse northern areas. For instance, Greece and Turkey collectively hold 12 IUCN Centers of Plant Diversity (Davis et al. 1994) but have only 10 botanic gardens between them. In contrast, the United Kingdom has 80 botanic gardens but no Centers of Plant Diversity. Connecting conservation need to ex situ capacity is clearly a major challenge.
Conclusions Practitioners of ex situ plant conservation, contributing to the conservation of wild plant diversity in increasingly human-dominated landscapes, face a number of challenges. The future of these collections and, perhaps more importantly, the institutions themselves as effective agents of conservation depend on the identification and adoption of roles that are biologically, politically, and economically viable. In addition, practitioners need to clearly articulate the value of the core competencies of ex situ conservation (seed banking, horticulture, population and demographic management, public display, and education) to serve the needs of external stakeholders. For instance, although the genetic value of many existing botanic garden collections may be minimal, they represent a valuable opportunity to promote the need for conservation and to generate the political support and economic resources needed to support the retention of wild plant diversity (Czech et al. 1998; Le Maitre et al. 1997, Maunder et al. 2001b, 2001c). Similarly, a core skill resident in botanic gardens, the cultivation of wild plant species, is consistently undervalued by both botanic gardens and conservation agencies. We argue in this book that ex situ facilities are underusing some valuable core resources.
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There is still a wide gap between ecological, genetic, and physiological research and its routine application to ex situ conservation. Practitioners need to establish a dialog with relevant research groups to allow ex situ management to become biologically more effective and financially more efficient. Effective and easy-to-use (particularly by nonspecialists) protocols for planning and implementing ex situ plant conservation are urgently needed. These protocols can be built from the developing experience in recovery planning and captive breeding (Clark and Cragun 1994; Ellis and Seal 1995). Identifying and implementing contemporary roles for ex situ facilities that directly support the primary imperatives of species and habitat conservation will require an imaginative and opportunistic approach. The future challenge for ex situ conservation is to maintain plant populations as both evolutionary lineages and potential components of functioning wild habitats.
References Benford, G. 1992. Saving the “library of life.” Proceedings of the National Academy of Science USA 89:11098–11101. BGCI (Botanic Garden Conservation International). 2001. Botanic Garden Agenda for Conservation. London: Botanic Gardens Conservation International. Blatter, E., and W. S. Millard. 1993. Some Beautiful Indian Trees. Bombay: Bombay Natural History Society and Oxford University Press. Clark, T. W., and J. R. Cragun. 1994. Organizational and managerial guidelines for endangered species restoration programs and recovery teams. Pages 9–33 in M. L. Bowles and C. J. Whelan (eds.), Restoration of Endangered Species. Cambridge, UK: Cambridge University Press. Conti, F., A. Manzi, and F. Pedrotti. 1992. Libro Rosso delle Piante d’Italia. Rome: Associazione Italiana per il World Wildlife Fund. Cotterman, L., and C. Jones-Roe. 1996. Botanical gardens and arboreta: partners in conserving biological diversity. Natural Areas Journal 1(1):1–5. CPC (Center for Plant Conservation). 1991. Genetic sampling guidelines for conservation collections of endangered plants. Pages 225–238 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Cugnac, A. de. 1953. Le rôle des jardines botaniques pour la conservation des espèces menacées de disparition ou d’altération. Annales de Biologie 29:361–367. Czech, B., P. R. Krausman, and R. Borkhataria. 1998. Social construction, political power, and the allocation of benefits to endangered species. Conservation Biology 12:1103–1112. Davis, S. D., V. H. Heywood, and A. C. Hamilton (eds.). 1994. Centres of Plant Diversity: A Strategy for Their Conservation. Vol. 1. Europe, Africa, South West Asia and the Middle East. Gland, Switzerland: IUCN/WWF/ODA.
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Lucas, G. L., and H. Synge. 1978. The IUCN Plant Red Data Book. Morges, Switzerland: IUCN. Maunder, M., R. S. Cowan, P. Stranc, and M. F. Fay. 2001a. The genetic status and conservation management of two cultivated bulb species extinct in the wild: Tecophilaea cyanocrocus (Chile) and Tulipa sprengeri (Turkey). Conservation Genetics 2:193–201. Maunder, M., and A. Culham. 1997. Practical aspects of threatened species management in botanic garden collections. Pages 122–130 in T. E. Tew, T. J. Crawford, J. W. Spencer, D. P. Stevens, M. B. Usher, and J. Warren (eds.), The Role of Genetics in Conserving Small Populations. Petersborough: Joint Nature Conservation Committee and British Ecological Society. Maunder, M., A. Culham, B. Alden, G. Zizka, C. Orliac, W. Lobin, A. Bordeu, J. M. Ramirez, and S. Glissmann-Gough. 2000. Conservation of the toromiro tree: case study in the management of a plant extinct in the wild. Conservation Biology 14(5):1341–1350. Maunder M., S. Higgens, and A. Culham. 2001b. The effectiveness of botanic garden collections in supporting plant conservation: a European case study. Biodiversity and Conservation 10(3):383–401. Maunder M., B. Lyte, W. Baker, and J. Dransfield. 2001c. The conservation value of botanic garden palm collections. Biological Conservation 98:259–271. Maunder, M., M. Stanley Price, and P. S. Soorae. 2002. The role of tropical botanical gardens in supporting species and habitat recovery: East African opportunities. Pages 115–134 in M. Maunder, C. Hankamer, C. Clubbe, and M. Groves (eds.), Plant Conservation in the Tropics: Principles and Experiences. Kew, UK: Royal Botanic Gardens. Mayr, E. 1982. The Growth of Biological Thought: Diversity, Evolution and Inheritance. Cambridge, MA: Belknap Press of Harvard University Press. McCracken, D. P. 1997. Gardens of Empire: Botanical Institutions of the Victorian British Empire. London: Leicester University Press. McMahan, L. R. 1995. Working with the Feds. The Public Garden 10(2):16–19. McMahan, L. R., and E. O. Guerrant Jr. 1991. Practical pointers for conserving genetic diversity in botanic gardens. The Public Garden 6(3):20–25, 43. Mill, J. 2002. The Australian Network for Plant Conservation. Pages 91–113 in M. Maunder, C. Hankamer, C. Clubbe, and M. Groves (eds.), Plant Conservation in the Tropics: Principles and Experiences. Kew, UK: Royal Botanic Gardens. Mistretta, O. 1994. Genetics of species reintroductions: applications of genetic analysis. Biodiversity and Conservation 3(2):184–190. Myers, N., R. A. Mittermeier, C. G. Mittermeier, G. A. B. da Fonseca, and J. Kent. 2000. Biodiversity hotspots for conservation priorities. Nature 403:853–858. New England Wildflower Society. 1992. New England Plant Conservation Program. Wild Flower Notes 7(1):1–79. Osborne, M. A. 1995. Nature, the Exotic, and the Science of French Colonialism. Indianapolis: Indiana University Press. Prest, J. 1981. The Garden of Eden: The Botanic Garden and the Re-Creation of Paradise. London: Yale University Press. Reichard, S. H., and P. S. White. 2001. Horticulture as a pathway of invasive plant introductions in the United States. BioScience 51:103–113.
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Robertson, I. M. 1996. Botanical gardens in a contemporary world. Public Garden 11(1):16–21. Sala, O. E., F. S. Chapin III, J. J. Armesto, E. Berlow, J. Bloomfield, R. Dirzo, E. Huber-Sanwald, L. F. Huenneke, R. B. Jackson, A. Kinzig, R. Leemans, D. M. Lodge, H. A. Mooney, M. Oesterheld, N. L. Poff, M. T. Sykes, B. H. Walker, M. Walker, and D. H. Wall. 2000. Global biodiversity scenarios for the year 2100. Science 287:1770–1774. Soulé, M. E., M. Gilpin, W. Conway, and T. J. Foose. 1986. The millennium ark: how long a voyage, how many staterooms, how many passengers? Zoo Biology 5:101–113. Thibodeau, F. R., and D. A. Falk. 1987. Building a national ex situ conservation network: the U.S. Center for Plant Conservation. Pages 285–294 in D. Bramwell, O. Hamann, V. H. Heywood, and H. Synge (eds.), Botanic Gardens and the World Conservation Strategy. London: Academic Press. Vavilov, N. I. 1926. Studies on the origin of cultivated plants. Bulletin of Applied Botany, Genetics and Plant Breeding 16:1–248. Vavilov, N. I. 1949–1950. The origin, variation, immunity and breeding of cultivated plants. Chronica Botanica 13:1–366. Walter, K. S., and H. J. Gillett (eds.). 1998. 1997 IUCN Red List of Threatened Plants. Compiled by the World Conservation Monitoring Centre. Gland, Switzerland: IUCN, The World Conservation Union. Warren, L. M. 1995. The role of ex situ measures in the conservation of biodiversity. Pages 129–144 in C. Redgwell and M. Bowman (eds.), International Law and the Conservation of Biological Diversity. London: M. Kluwer Law International. White, P. S. 1996. In search of the conservation garden. The Public Garden 11(2):11–13, 40. Wilson, E. H. 1919. The romance of our trees: II, the ginkgo. Garden Magazine 30(4):144–148. Wyse Jackson, P. 2001. An international review of the ex situ plant collections of the botanic gardens of the world. Botanic Gardens Conservation News 3(6):22–33.
Chapter 2
In Situ and Ex Situ Conservation: Philosophical and Ethical Concerns Holmes Rolston III
The Natural and the Artificial In one sense, nature is quite a grand word, referring to everything generated or produced. Natura or physis is the source from which all springs. If one is a metaphysical naturalist, then nature is all that there is. Metaphysical naturalists may need the word in this sense for their cosmological purposes; the contrast class might be the supernatural, which, they may argue, is an empty set. Humans and all their cultural activities are included as natural; humans are generated within nature, and they break no natural laws. Under this definition, everything agricultural or technological is completely natural. So is everything industrial, political, economic, philosophical, or religious. So is anything that happens in a botanical garden. Such scope is problematic, however, because it prevents discriminating analysis of the differences between spontaneous nature and deliberated culture. A predicate, “natural,” that includes all actual and possible properties excludes nothing; denoting everything is like denoting nothing, at least nothing in particular. The most forceful objection to this sense of nature, in the context of environmental analysis, is that such a definition allows no useful contrast with culture, but we need to analyze that contrast carefully if we are going to relate our cultures to nature, asking about nature conservation goals. A straightforward contrast class to nature is culture. If I am hiking across wildlands, the rocks and wildflowers, the birds, and even their nests are natural, but if I come upon an abandoned boot or a candy wrapper, these are artifacts, unnatural. Expanding such examples into a metaphor, the whole of civilization is producing artifacts, in contrast to the products of wild 21
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spontaneous nature. Wild animals, much less plants, do not form cumulative transmissible cultures, elaborating these artifacts over generations. Humans evolved out of nature; our biochemistries are natural, and we draw our life support from the hydrological cycles and photosynthesis; we have genes and inborn traits; and we are subject to natural laws. But human life is radically different from that in wild nature. Unlike coyotes or bats, humans are not just what they are by nature; we come into the world by nature quite unfinished and become what we become by culture. Humans deliberately rebuild the wild environment and make rural and urban environments. Information in nature travels intergenerationally on genes; information in culture travels neurally as people are educated into transmissible cultures. They learn how to build fires, or make spears, or grow wheat, or make iron plows and grow more wheat, or make trains on which to ship their wheat to distant markets. They teach this know-how to ongoing generations. Humans argue about worldviews, about whether there should be natural prairies as well as wheat fields in Kansas. The determinants of animal and plant behavior are never anthropological, political, economic, technological, scientific, philosophical, ethical, or religious. The critical factor is the deliberated modification of nature that separates humans in their cultures from wild nature. Any transmissible culture, especially a high-technology culture, must be discriminated from nature. Boeing jets fly, as wild geese fly, using the laws of aerodynamics. The flight of wild geese is impressive; scientists can hardly be said to understand these “bird brains” and how they migrate. The information storage system in the goose genetics could, in its own way, be the equal of that by which Boeings fly. Some of the information in the geese is transmitted nongenetically, as when they learn migration routes by following other geese. Maybe we can even say that the geese deliberately build their nests or intend to fly south. But geese do not form cumulative transmissible cultures. It is only philosophical confusion to remark that both geese in flight and humans in flight are equally natural and let it go at that. No interesting philosophical analysis is being done until there is insight into the differences between the ways humans fly in their engineered, financed jets and the ways geese fly with their genetically constructed, metabolically powered wings. Geese fly naturally; humans fly in artifacts. Against this background, we can find some overlap and hybrids. The essential idea in calling nature a human resource is that some “source” in
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spontaneous nature is taken over by human deliberation and “re-sourced,” or redirected to human uses. Geese make resources of grasses and seeds; these plants make resources of sunlight and water. But the plants photosynthesize genetically; the geese build nests instinctively. Humans are differently resourceful because of what they learn about resources in their cumulative transmissible cultures. They know how to take a tomato from its natural source in an evolutionary ecosystem in Peru and redirect or “resource” it into channels of human interest and preference, as indigenous peoples did with the wild tomato, now grown around the world. We get better at it. Horticulturists in 1962 took a different wild tomato (Lycopersicon chmielewskii, Solanaceae) and bred it into the standard tomato (L. esculentum), enhancing it for the commercial tomato industry, resulting in $8 million a year profits (Rick 1974). In domesticated plants, nature is made over into an artifact that we can use. To use a more philosophical word, nature is transformed, its form transmuted into a more desirable humanized form. To use a scientific engineering word, human values carried by plants are synthetic. Hence we speak of agriculture, the deliberate, elaborated modification of fields and crops, or of horticulture, cultivating plants, with culture and nature in synthesis. Consider the growing of cotton. Human art has no independent powers of its own; it can only redirect natural processes. Cotton is a natural fiber, but cotton in fields is not spontaneously wild. The cotton fiber is produced genetically; the cotton is planted, fertilized, harvested, and spun by humans. In contrast to cotton, nylon is completely synthetic; no genetics produces the fibers. Of course, the chemists exploit natural properties, although these were never manifested in wild, spontaneous nature. It seems appropriate to say that cotton, though an artifact and hybrid, is more natural than nylon. We will need this relative sense of natural when we reach the distinction between in situ and ex situ conservation. Plants in botanical gardens become artifacts, but maybe they can be more natural than cultivars.
Plants and Intrinsic Values In wild, spontaneous nature, a plant is a living organism with a good of its own. Alternatively put, the plant defends its life as an intrinsic value, as it is doing when it photosynthesizes, making a resource of sunlight by capturing energy and redirecting it to plant metabolism. Like all other organisms,
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plants are self-actualizing. A plant is not a subject, nor is it an inanimate object, like a stone. Plants, quite alive, are unified entities of the botanical but not of the zoological kind. That is, they are not unitary organisms highly integrated with centered neural control, but they are modular organisms, with a meristem that can repeatedly and indefinitely produce new vegetative modules, additional stem nodes and leaves when there is available space and resources, and new reproductive modules, fruits and seeds. Plants make themselves. A plant grows, reproduces, and resists death, maintaining a botanical identity. Plants repair injuries and move water, nutrients, and photosynthate from cell to cell; they make tannin and other toxins and regulate their levels in defense against grazers; they make nectars and emit pheromones to influence the behavior of pollinating insects and the responses of other plants; they emit allelopathic agents to suppress invaders; they make thorns and trap insects. They can reject genetically incompatible grafts. From one perspective, all this is just biochemistry—the whir and buzz of organic molecules, enzymes, proteins—as humans are from one perspective. But from an equally valid and objective perspective, the morphology and metabolism that the organism projects are a valued state. Vital is a more ample word now than biological. We could even argue that the genetic set is a normative set; it distinguishes between what is and what ought to be, not in any moral or conscious sense, of course, but in the sense that the organism is an axiological system. The genome is a set of conservation molecules. A life is spontaneously defended for what it is itself. The plant, we can say, is valuable itself: “value-able,” able to protect this botanical form of life. That is, such life is intrinsically valuable. Philosophers and even zoologists may here protest: nothing “matters” to a plant; plants do not have the minimally sentient awareness necessary to be centers of felt experience. But, although things do not matter to plants, a great deal matters for them. Botanists ask of a failing plant, “What’s the matter with that plant?” If it is lacking sunshine and soil nutrients, and we arrange for these, we say the plant is benefiting from them, and everywhere else we encounter it, benefit is a value word. Objectively, biologists regularly speak of the selective value or adaptive value of genetic variations (Ayala 1982: 88; Tamarin 1996: 558). Plant activities have survival value, such as the seeds they disperse or the thorns they make. Plants are not valuers with preferences that can be satisfied or frustrated. We do not say that wildflowers have rights or need our sympathy or that we
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should consider their point of view. But we do claim that every organism has a good-of-its-kind; it defends its own kind as a good kind. An objector can say, “The plants don’t care, so why should I?” But plants do care— using botanical standards, the only form of caring available to them. Plant conservation does not begin when someone from the Botanic Garden goes into the field to see what is threatened and needs conservation. The plant is already, by itself and on its own, a project in conservation biology. The conservation of biological identity within organisms is the first law of life. Ethics and biology have had uncertain relations in recent centuries. An often-heard argument forbids moving from what is the case (a description of biological facts, as with these plants conserving their lives) to what ought to be (a prescription of duty, such as human caring for these plants). Any who do so commit the naturalistic fallacy. On the other hand, if spontaneous natural lives are of value in themselves, and if humans encounter and jeopardize such value, it seems that humans ought not to destroy values in nature, not at least without overriding justification producing greater value. Perhaps some of these plant kinds are bad kinds (such as poison oak), but because in their place they are adapted fits, they are presumptively well suited for life in their niches. Perhaps many of them are of no particular value to us, but it seems both unscientific and arrogant to conclude that there is nothing of value there at all. Indeed, the presumption can be the other way around. If there is already conservation biology in the wild, if a plant is already engaged in the biological conservation of its identity and kind, long before conservation biologists come on the scene, then what conservation biologists ought to do is respect plants for what they are in themselves: projects in conservation biology. That aligns human ethics with objective biology. We want these plants for the uses we might make of them. Given the multiple ways in which humans use plants—agriculturally, industrially, medically, recreationally, aesthetically, scientifically, as cultural symbols, as environmental indicators, and as part of the human life support system—humans are going to be helped or hurt by their flora, of which even rare plants may form a critical part. Biodiversity means opportunities of many kinds, so we save them for the benefits they may bring. But we also may be wishing to protect something of this integrity, this value, in plants in the wild. For most people active in conservation biology this is a genuine concern. That puts ex situ conservation botanists in something of a bind, however, because in the form of caring they take for their plants, human
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botanists have removed the wild plants from their autonomous independence in the world and brought them under their care. This careful management may, willy-nilly, have prejudiced that wild value. The plants in botanical gardens may be hybrids, relatively, and to the same degree something of this value that we want to preserve may have been lost.
Plants in Ecosystems A plant is what it is where it is, that is, in situ. In the wild, both the individual plants and the species lines in which they stand are embedded in ecosystems. The situation of surviving plants even in their native locations may already be skewed by cultural disturbances, so that they are in marginal, not typical, habitat. Plants, autotrophs, have a certain independence that animals, heterotrophs, do not have. Plants need only water, sunshine, soil, nutrients, and local conditions of growth; animals, often mobile and higher up the trophic pyramid, may range more widely but in this alternative form of independence depend on the primary production of plants. Every form of life is what it is in a niche, shaped as an adaptive fit. The product, an individual organism, is process in a historical lineage, populations in their species lines. Such a lineage is the outcome of entwined genetic and ecological processes; the generative impulse springs from the genes, defended by information coded there, but the whole organism survives when selected by the environment in a niche occupied by the species. At this level, conservation concerns the processes as much as the products. On evolutionary scales, these processes have involved regular species turnover when a species becomes unfit in its habitat, goes extinct, or tracks a changing environment until transformed into something else. On these timescales, species too are ephemeral. But the speciating process is not. Persisting through vicissitudes for two and a half billion years, speciation is about as long-continuing as anything on Earth can be. In that sense, evolutionary ecosystems have been the fundamental unit of survival, dynamically vital in elaborating the biota from zero to several million species. Evolutionary ecosystems conserve life, as much as do individuals in their species lines. The biodiversity on hand is a legacy of remarkable fertility and exuberance: several billion years of creative struggle. We do not yet have a complete theoretical account of this richness of life, but bioscience gives us this certainty: the evolutionary odyssey is prolific, pro-life.
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Organisms defend only their own selves or kinds, but the system spins a bigger story. Organisms defend their continuing survival; ecosystems have stimulated new arrivals. Species increase their kinds, but ecosystems have increased kinds and increased the integration of kinds. The system is a kind of field with characteristics as vital for life as any property contained in particular organisms. The individuals located in species lines in ecosystemic niches are so placed with random and contingent elements, but plants in their niches are not simply in accidental aggregations. The ecosystemic matrix is the depth source and support of individual and species alike. We argued earlier that plants are valuing organisms. Can we not now view ecosystems or, more broadly, the planetary biospheric system as valuegenerating systems and, in a real sense, value-able, able to generate value? Over the millennia, there is natural selection for adapted fit; there appear the myriad species filling up their habitats. There are extinction and respeciation. Forests repeatedly evolve; so do grasslands. This self-organizing has been called autopoiesis. This generativity is the most fundamental meaning of the term nature, “to give birth.” Ecosystems are the womb of life. But are they the kind of womb that plants and animals can ever leave? Ecosystems are both womb and matrix of life. Plants and animals live in biotic communities, and an ethic of respect for life must embrace these communities. “A thing is right,” concluded Aldo Leopold, “when it tends to preserve the integrity, stability, and beauty of the biotic community. It is wrong when it tends otherwise” (1968: 224–225). Leopold wanted a land ethic, one that included concern for individual plants, animals, and people but also and fundamentally loved and respected biotic communities. Now we reach the conclusion that the appropriate unit for concern is the fundamental unit of development and survival. But zoos and botanical gardens are not ecosystems. And what if the preservation of individuals is impossible without the preservation of ecosystems? In wild nature, there are no organisms in isolation; there are only organisms in ecosystems. Perhaps we were too hasty in locating those intrinsic values in plants, forgetting that a plant is not self-contained, despite its being an autotroph, but situated in an ecosystem. So when we, in culture, move the plant to a botanical garden, ex situ, it may first seem that we have transplanted the whole plant and that the plant is flourishing in its new home. We first think we have the whole plant, but then we realize that we do not have the whole in which it was planted. We forget that the plant is at home only in its ecology; that is the root meaning of ecology, the logic of a home. In that sense,
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the animal in the zoo and the plant in the garden no longer have any ecology; they are not at home. Botanists are concerned to save genetic resources; they may believe they have done this when they have seeds in their seed banks, when they have plants in their botanical gardens. But plants and their seeds evolved in ecosystems, and they can continue only in ongoing ecosystems. In the current debate among biologists about the levels at which selection takes place—individual organisms, populations, species, genes—the recent tendency to move selective pressures down to the genetic level forgets that a gene is always emplaced in an organism that is emplaced in an ecosystem. The molecular configurations of DNA are what they are because they record the story of a particular form of life in the macroscopic, historical ecosystem. What is generated arises from molecular mutations, but what survives is selected for adaptive fit in an ecosystem. We cannot make sense of molecular life without understanding ecosystemic life. One level is as vital as the other.
Captive Plants Nevertheless, halfway between the molecular genetic and the evolutionary ecosystemic levels, we have these plants, the phenotype organisms, in botanical gardens, ex situ. The plants are also halfway between nature and culture, we could say. We have what I will call—analogously to animals in zoos, if also a little provocatively—captive plants. You might wish to say “managed” plants instead, objecting that whereas animals can be held in captivity, plants cannot. The idea of captivity deprives an organism of its locomotive freedom, which animals have. They want to move beyond the bars but cannot. Plants, by contrast, have no locomotive freedom; therefore, they cannot be held captive. “Captive plants” is a category mistake. Ask the question another way. Have you deprived these plants of their autonomy? Are they still defending goods of their own, on their own? Alternatively put, Are they still wild? Or are they hybrids of nature and culture? You probably do not think of what goes on in a cotton field as being “wild.” “Natural,” yes, when the rains come and the plants take up water, when the sun comes out and the plants photosynthesize. But the plowing and the fertilizer, the managed care, mean that the cotton plants are no longer wild. Hybrids have not lost their naturalness entirely, but they have lost their wildness. They are domestics. Some of these domestics, such as
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maize, have been so modified genetically by long selective breeding that they can no longer survive in the wild. Most of them, such as tomatoes, have altered genomes that are hybrids of nature and culture. Yes, you may say, that is true of cultivars that have been bred for human purposes. But it need not be true of plants that, though taken under human care in botanical gardens, are not selectively bred but left to themselves. But are they left to themselves? A new, controlled environment has been selected for them; water and nutrients are supplied, and they are watched over while left to themselves in this environment believed to be artifacted to mimic the wild. One needs only to move a few feet or a few inches away from the plant, and it is evident that this is no wild ecosystem. The plants no longer have to, or can, disperse their seeds, for example, but this is a first priority for any wild plant. This is a managed botanical garden. Maybe the problem is that the plants are left all by themselves, isolated from the webworks of ecosystems. They have been removed from their coadapted gene complexes, perhaps those of the insects that pollinate them, or the fungi in the soil in which they root. They have no niche. Once a dynamic organism is severed from its functional context, it ceases to be that thing. Aristotle (1961) has a memorable analysis in which he comments of a hand severed from a body that it used to be a hand, but is no more because of the disconnection (De partibus animalium, 640–641). We see that with organs in organisms, but it is equally true of organisms in ecosystems. By this analysis, a wolf in a zoo used to be a wolf. The brains of lions in zoos rapidly disintegrate. A bear without a forest is a compromised bear. It has lost what Aristotle calls—rather provocatively for us—its “soul,” its psyche or anima. Very few of the animals in zoos could be returned to the wild, even if we wanted to do so. They have become dependent on humans; they never developed the needed survival skills. They may no longer have the genetic competence for such skills, as with all the Siberian tigers in zoos today. Returning a zoo monkey to the wild is like turning a cocker spaniel loose in the wilderness. A half dozen species endemic to San Clemente Island, off the coast of California, were threatened with extinction because they were being eaten by feral goats. The goats were introduced by sailors as a fresh meat supply a century and a half ago. New concern for the conservation of these plants arose after the Endangered Species Act was passed. Some were transplanted to the Santa Barbara Botanic Garden. By this account, we would have to say of a San Clemente Island bushmallow (Malacothamnus
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clementinus, Malvaceae) translocated to Santa Barbara that it used to be a Malacothamnus clementinus but is no more. It is an amputated species. U.S. Fish and Wildlife authorities and the U.S. Navy, which controls the island, have eliminated the goats by translocating and killing them. About 14,000 live goats were moved to the mainland; about 15,000 goats were shot. This was to protect the endangered plants and to restore ecosystem health more generally to the island. They did this despite protests, including lawsuits, from animal rights advocates that this was cruel and that the welfare of the goats ought to take priority over that of a few endemic plants. Several dozen goats were shot for each known surviving plant. The rationale is that this restores an ecosystem in which Malacothamnus clementinus can continue to be Malacothamnus clementinus because the plant is at home in its ecology (Ottie 1982; Mohlenbrook 1983: 183–184; C. Winchell, pers. comm., 1991). Eliminating natural selection at once begins to alter a species, even if no artificial selection is intended to replace it. Species formerly under selection pressures may undergo random drift. More likely, there will be unintended artificial selection pressures. Geneticists have found that endangered fish, kept in hatcheries to be bred for reintroduction programs, are genetically different in two or three generations (Meffe 1986). In an effort to eradicate a cattle pest, the screwworm fly was mass reared to produce males that could be sterilized by gamma rays. Researchers found that the lack of natural competition inadvertently selected genome changes in a few generations such that males had reduced competitive ability (Bush et al. 1976). Similarly in fruit flies: “even ‘properly managed’ populations of captive Drosophila lost 74 per cent of their reproductive fitness after 11 generations and had lower genetic diversity than large wild populations. Captive animals rapidly adapt genetically to captivity. Animals adapted to captivity are likely to reproduce more poorly in the wild” (Ralls and Meadows 1993: 690; Frankham and Loebel 1992). Even if the lack of competitive natural selection pressures in the zoo or botanic garden is insignificant, there seems to be a strong possibility of putting captured animals and plants, even when soon returned to the wild, through a genetic bottleneck, or inbreeding depression. Another problem is that if plants are needed for reintroduction to the wild, the seeds that are grown will be selected for maximum reproductivity in gardens, but selecting the most fecund plants in the garden may not be the same as selecting maximum reproductivity in the wild, which is tested in a different environment and over the entire lifespan of the plant.
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Or, realizing the dangers of genetic bottlenecks and small population samples, conservation biologists may go to extra effort to preserve and to mix plants from multiple locations, thereby increasing genetic diversity, as they intend, but inadvertently reducing a plant’s fitness in the particular locales in which it is subsequently reintroduced. “Active management to maintain maximum genetic diversity may ironically be at odds with rapid adaptation to local conditions” (Guerrant 1996: 186). Plants are subject to outbreeding depression, and mixing stocks (perhaps in the well-intended interests of increasing diversity) can be counterproductive, disrupting evolutionary and ecological adaptation to local conditions. Plants seem to be adaptively fine-tuned to their particular localities, and translocating them disrupts this idiographic adaptation (Guerrant 1996: 199). One cannot manage without a strategy, and whatever the strategy, it is likely to relax natural selection pressures. The idea of managed, wild plants is a contradiction in terms. For example, plant succession never takes place in botanic gardens or seed banks, yet every species is more or less affected by whatever tendencies toward succession are present in its natural habitat. Or if one prefers the more chaotic accounts of recent ecology, chaos is not present in botanic gardens or seed banks. Processes of dynamic change, omnipresent in ecosystems, are absent in botanic gardens. Or perhaps we should say that the processes of dynamic change are cultural rather than natural. To this extent, the captive plants become artifacts.
Wild, Compromised, and Faked Nature The question we finally reach is whether ex situ conservation will complement or undercut in situ conservation. An answer likely to be given is that in situ conservation is best, but where it is not possible, ex situ conservation is third best. Second best is interim ex situ conservation prospective to in situ restoration. In a general way, one can hardly disagree with such pragmatism, but there are pitfalls to such compromise that must be analyzed. Set ideals aside and get real, one might object. When one is faced with win-or-lose decisions, especially in political democracies and capitalist economies, win all you can and be realistic about what you must lose. In actual decision-making contexts, the best rule is compromise. But this is not necessarily true; this depends on the contexts of opportunity and jeop-
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ardy. Compromising politically and compromising botanically are two different activities, and they relate with uncertainty. Compromises often put at risk what they propose to save. An old saying is that “Half a loaf is better than none,” and this is true for those who are arguing over bread. But what if they are arguing over a horse? Some values do not compromise without being devalued, even destroyed. A dozen rare plants in a botanical garden, rescued and transplanted there prospective to relocation, though seemingly flourishing in their new location, may be doomed the moment their site is bulldozed over for a new dam. Compromises must be set in the perspective of what has already been lost. Perhaps the rare plants are taken from an area that might have been set aside as sanctuary. Whereas some plants are naturally rare, many have been made artificially rare through habitat loss. Possibly a sanctuary of a few thousand acres would have been only 1 or 2 percent of the regional habitat the species once occupied. In dispute, we might be tempted to compromise, save the plants in a garden, and plan to relocate them elsewhere. But such a compromise only further skews the imbalance of nature and culture, and if no viable population is saved, compromise loses both sanctuary and species. “In politics compromise is the name of the game.” So? Where there is choice against choice, one can expect that positive values will be at stake on both sides, and in a pluralist democracy we can often expect that compromise will optimize such values. Compromises can be fair and equitable. We incline to compromise when issues are complex, when there is evidently some value on both sides, when a decision is needed that is impossible to postpone or when postponing will result in value loss for both sides. Often, too, the facts and projections are uncertain, which makes us less sure of our position. Compromise can win something, and uncompromising purity is a sure route to defeat. Better to have some of the plants in gardens, with a chance of reintroduction, than to have no plants at all. But those alert to the logic of compromise also know that compromises can mean destruction. Compromise is likely to cast the solution in terms of who has interests to adjudicate and is noisy about them. But the better question is, What is of value in the world, and how ought we behave so as to optimize those values? Compromise is likely to mean that decisions are made in courts (or outside courts lest courts be invoked), but this means that those who have power to do adversarial work succeed; this may not always be the best way to reach decisions. Adversaries are not always the
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best optimizers; there is no invisible hand that guides adversarial relations into optimal solutions. There are values here to be discovered, not just interests to be defended. If the rationale for conservation is largely humans and their plant resources, compromising conservationists will argue that we have what we want in the botanical gardens and the seed banks: the genetic resources of interest to humans. Any ongoing dynamism of wild plants in their habitats is on evolutionary scales, too iffy and indiscernible to make any difference to what we want, 99 percent of which we can get with the present genetic diversity on hand in ex situ storage. We can compromise, save the genetic diversity, and build the new ski area. Such a reply is likely to prove in error on several counts. Seed banks and botanic gardens can preserve only an infinitesimal fraction of the allelic diversity and evolutionary potential in wild nature—the genetic bottleneck problem. Rich potential resources in wild nature will lapse into extinction while we labor under the illusion that we have what we want in the gardens and banks. Furthermore, evidence already cited shows that genomes can change more rapidly than often supposed. If fishes, fruit flies, and tomato plants can respond to altered selective pressures as quickly as they have, this suggests that the dynamism of natural selection in the wild is significant on similar timescales. True, the altered selective pressures of culture may be more dramatic than changes to be expected on evolutionary timescales. But the domesticated plants may soon have less genetic diversity than we thought. There is high probability that the results of our compromise, taking the plants into managed care, even intending their restoration, will be different from what we thought. Much environmental law allows for mitigation. Developers who encounter endangered plants in their way will consider mitigation. That seems common sense. Move the plants. Create a new wetland, or riparian zone, or translocate to a similar habitat in the next county, where a nature reserve is possible at a third of the cost. Botanic gardens and seed banks will be seen as sources of mitigation. But as everyone familiar with mitigation efforts knows, these sources have been notoriously unsuccessful. Therefore, ex situ conservation, even when it is claimed to be prospective to restoration, will be used to justify the increased invasion of areas that, without ex situ conservation, would not have been so readily invaded. Zoos and botanical gardens will undermine the imperative to conserve existing sites.
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Although I am a philosopher, not a biologist, if I read the literature correctly, reintroductions, though sometimes successful, have often proved more complex than anticipated. Most reintroductions over the last century have failed; most of the time the causes of failure (or success) are unknown. With better conservation biology, reintroduction success is improving, still with formidable unknowns (Guerrant and Pavlik 1997: 92–93, 104). That does not bode well for compromises that settle for ex situ conservation interim to recreated in situ conservation. One will always be trading the absolute certainty of existing populations against the high uncertainty of replacement populations. At a minimum that places a high burden of proof on those who propose mitigation. If and when the compromised plants are successfully restored to the wild, we might have what some have called “faked nature” (Elliot 1998; Katz 1992). The restored plants have suffered the loss of temporal continuity by having been removed from the wild to the managed garden, even if they are later restored to the wild. They do not cease to be artifacts when they are cleverly put back in place by restoration biologists. When you take visitors to reintroduction sites, you probably do not say, “Here, let me show you some wild plants.” You say, “Here, let me show you our successful reintroduction.” But by that you reveal that these plants are different from wild ones, different because of the human intervention. Standing before a Torrey pine (Pinus torreyana, Pinaceae) along the coast of southern California, the proper response is not, “Wow, there is a rare species, surviving across millennia!” but “Hurrah for the U.S. Forest Service!” Their biologists in 1986 collected 30,000 seeds from 150 trees for ex situ storage and propagation and reintroduced the pine, producing nearly 6,000 trees. Besides this, they had to control an outbreak of the ips beetle. One admires not so much the trees as the skills of the restoration biologists who put them there. The pines are not really wild. Once upon a time, they were, but now, though apparently in situ, the truth is that they exist thanks to biologists and their ex situ facilities. This objection can be met, though perhaps only partly. One has to recognize that nature returns. Nature is still in situ, and if we situate the plants there, they grow wild again. The compromise is not forever. Notice that there are all kinds and degrees of restoration. At the one extreme, if a forest has been clearcut or stripmined, there is nothing there; the landscape is blitzed, so any new forest is a replacement, a replica. This would be like replicating the Nina, one of Christopher Columbus’s ships. The replica is
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made from scratch and has no historical continuity with the original. This is not really restoration; it is replication. On the other end of a spectrum, if a few of the trees in the forest have been cut by selection and new trees replanted to substitute for them, there is restoration. If some of the Torrey pine is removed and the Forest Service puts others back, this is restoration, the original, once damaged and now restored. A replica is a new creation, without continuity to the old one. Replicas can exist simultaneously with originals. Restorations cannot. Nature restored need not be nature faked. Restoration in nature, unlike restoration in art, is really rehabilitation. A restored painting, which is an artifact, does not heal itself when restored; it is a passive object. One does not rehabilitate paintings. But once we put the parts back in place, nature may heal itself. One can revegetate after a strip mine, but one cannot rehabilitate it because there is nothing to rehabilitate. One can rehabilitate a prairie that has been not too badly overgrazed. Overgrazing allows many introduced weeds to outcompete the natives; perhaps all you have to do is pull the weeds and let nature do the rest. That is undoing as much as doing. Overgrazing allows some native plants to outcompete other natives, those that once reproduced in the shade of the taller grasses. So perhaps, after the taller grasses return, you will have to dig some holes, put in some seeds that you have gathered from the missing plants, held ex situ in the botanic garden, cover them up, go home, and let nature do the rest. Perhaps you can just put the seeds in the weed holes. The naturalness returns. The restoration ceases to be an artifact. In the days before high-tech medicine, physicians who were congratulated on their cures used to say, modestly, “Really, I just treated you, and nature healed you.” A physician who sets a broken arm just holds the pieces in place with a splint, and nature does the rest. The doctor is not really to be congratulated for his or her skills at creating arms. The doctor arranges for the cure to happen naturally. One does not complain, thereafter, that one has an artificial limb. Likewise with restoration. It is more like being a midwife than being an artist or engineer. You arrange to get the raw materials back on site and place them where they can do their thing. The point is that restorations of this kind do not fake so much as facilitate nature, help it along, mostly by undoing the damage humans have introduced and then letting nature do for itself. As the restoration is completed, the wild processes take over. The sun shines, the rains fall, the for-
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est grows. Birds arrive on their own and build their nests in the restored pines. Natural selection takes over. The adapted fits survive in their niches. Succession resumes. In due course, lightning will strike and wildfire burn the forest again, after which it will regenerate itself. Even a new species could evolve. If such things happened decades, centuries, millennia after some thoughtful humans had once facilitated a restoration, it would seem odd to label all these events as artifacts, lies, fakes. Perhaps the best way to think of it is that the naturalness of a restored area is time bound. Any restoration is an artifact at the moment that it is deliberately arranged, but it gradually ceases to be so, and spontaneous nature returns— as long as humans back off and let nature take its course. Nevertheless, the unbroken historical continuity in natural systems is important. That we, after restoration, back off to let nature take its course proves that we could wish that the course of nature had never been broken on the landscape we now conserve. We are glad to have a broken arm healed; we would just as soon never have broken it. Although the spontaneity of natural systems might all return, the historical discontinuity can never be repaired. In that respect, the restored area does suffer permanent loss of natural value. Natural systems, like human beings, are not replaceable in their historical identity and particularity. They are characteristically idiographic and deliver their values in historical process, diminished in value if interrupted. Restoring does not restore this interruption. If one is appreciating the present spontaneity of wild nature—the plant or animal in its ecology—it can be returned, and after complete restoration it will be present undiminished. But if one is appreciating the evolutionary history—the plant or animal in its historical lineage—even though the genetics may be back in place, there has been interrupted wildness. The forest is not virgin, not pristine. It is less real. The danger is that ex situ conservation, in its admirable zeal for restoration and refining its skills at this, will discover that it has made more attractive this second-best solution. This would be something like a physician discovering that he was so skilled at resetting broken arms that his patients were more careless and that he was resetting twice as many as before. If nature means absolutely pristine nature, totally unaffected by human activities, past or present, there is little remaining on Earth if our detection instruments are keen enough. One can undoubtedly detect various human-introduced pollutants in the plants in Yellowstone or note that the vegetation is different because of fire suppression. Invasive, exotic plants are
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a problem, and they threaten to make the flora unnatural. Global warming threatens to shift natural patterns. Everything in nature has been compromised, they say; it is useless to seek real nature. Taken to its logical conclusion, this argument holds that the slightest human intervention has a totalizing effect and brings straightaway the end of nature. This is like saying that the whole moon is pristine no more because the astronauts took a few steps on it or that the sky is not natural because some jet planes have flown through it. It is true that certain human actions have unintended consequences that spread everywhere; there are contagious effects that eventually interrupt everything, that seep into the nooks and crannies of all nature. However, most human activities do not have such far-reaching effects. The world is too pluralist for that. Not everything is that tightly bound up to everything else. For instance, is it the case that, because humans first removed and then restored the bison and the wolves in the Yellowstone Park ecosystem, we have lost any possibility of letting the park be natural? In an absolute sense this is true because there is no square foot of the park in which humans, disturbing the predation pressures, have not increased and not shifted the patterns of ungulate grazing. That affects the grasses and the forbs, the willows and the beavers. But it does not follow that nature has absolutely ended, everywhere compromised because it is not absolutely present. It does not follow that there is no native vegetation at all, because all of it has detectable human effects. Answers come in degrees. Events in Yellowstone can remain 99.44 percent natural on many a square foot, indeed on hundreds of square miles, in the sense (recalling the language of the Wilderness Act) that they are substantially “untrammeled by man.” We can put the wolves back and clean up the air, and we have recently done both, and both will have effects on the flora. Where the system was once disturbed by humans and subsequently restored or left to recover on its own, wildness can return. Mutatis mutandis, this applies to restored plants. After a generation or so, the plants do not know their interrupted history, even if we recall it in our history books. Ex situ conservation is always a means, never an end. Sometimes the conserved plants are human resources for profit or pleasure; some of the resulting human experiences in botanical gardens, such as enjoying the orchids there or conducting scientific studies of their genetics, could be considered ends in themselves. But such pleasures and studies ought
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also motivate us to lament this compromising captivity unless this is also a means to their reintroduction into the wild. By contrast, in situ conservation can be an end in itself, in more ways than one. Humans on wild sites conserve these plants for their intrinsic value, not instrumentally to some human uses of them. They respect the integrity these plants have on their own. Conservationists can and ought to take pleasure in this, too, but this human conservation biology complements the autonomous plant conservation biology. These goods-of-their-kind, restored, continue dynamically to defend their forms of life as good kinds, and they do so as good adapted fits in their ecosystems. The biodiversity of life on Earth is remarkable indeed, but this is still more remarkable: biologists committed to the care and conservation, the restoration and rehabilitation of such value in nature. Defending the goods of one’s kind, a fact of nature, passes over to defending the goods of others, not one’s kind, an environmental ethic unique to the human. References Aristotle. 1961. Parts of Animals. Cambridge, MA: Harvard University Press (Loeb Classical Library). Ayala, F. J. 1982. Population and Evolutionary Genetics: A Primer. Menlo Park, CA: Benjamin/Cummings. Bush, G. L., R. W. Neck, and G. B. Kitto. 1976. Screwworm eradication: inadvertent selection for noncompetitive ecotypes during mass rearing. Science 193:491–493. Elliot, R. 1998. Faking Nature: The Ethics of Environmental Restoration. London: Routledge. Frankham, R., and D. A. Loebel. 1992. Modeling problems in conservation genetics using captive Drosophila populations: rapid genetic adaptation to captivity. Zoo Biology 11:333–342. Guerrant, E. O. Jr. 1996. Designing populations: demographic, genetic, and horticultural dimensions. Pages 171–207 in D. A. Falk, C. I. Miller, and M. Olwell (eds.), Restoring Diversity. Washington, DC: Island Press. Guerrant, E. O. Jr., and B. M. Pavlik. 1997. Reintroduction of rare plants: genetics, demography, and the role of ex situ conservation methods. Pages 80–108 in P. L. Fiedler and P. M. Kareiva (eds.), Conservation Biology for the Coming Decade. 2nd ed. London: Chapman & Hall. Katz, E. 1992. The big lie: human restoration of nature. Research in Philosophy and Technology 12:231–241. Leopold, A. 1968. A Sand County Almanac. New York: Oxford University Press. Meffe, G. K. 1986. Conservation genetics and the management of endangered fishes. Fisheries 11(1):14–23. Mohlenbrook, R. H. 1983. Where Have All the Wildflowers Gone? New York: Macmillan.
2. In Situ and Ex Situ Conservation: Philosophical and Ethical Concerns Ottie, F. N. 1982. Feral animal removal program, San Clemente Island, California: decision. Federal Register 47(23), February 3, p. 5033. Ralls, K., and R. Meadows. 1993. Breeding like flies. Nature 361:689–690. Rick, C. M. 1974. High soluble-solids content in large-fruited tomato lines derived from a wild green-fruited species. Hilgardia 42:492–510. Tamarin, R. H. 1996. Principles of Genetics. 5th ed. Dubuque, IA: W.C. Brown.
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Chapter 3
Western Australia’s Ex Situ Program for Threatened Species: A Model Integrated Strategy for Conservation Anne Cochrane
The state of Western Australia represents some 33 percent of the landmass of the island continent of Australia (Figure 3.1). More than 12,000 vascular plant taxa are estimated to grow in the state (Western Australian Herbarium 1998), equating to nearly half of the total taxa in Australia. The most floristically diverse region of Western Australia is the South West Botanical Province, covering some 12 percent of the total land area of the state. An estimated 8,000 taxa are recorded from this province. The climate is Mediterranean, with wet winters and hot, dry summers that support extensive agriculture, forests, woodlands, and heath. Woody and herbaceous perennials dominate, and the major plant families include the Proteaceae, Fabaceae, Mimosaceae, Epacridaceae, Restionaceae, Orchidaceae, and Myrtaceae. The majority of the vegetation has coevolved with fire, and a predominant feature of these plants is their dependence on fire disturbance for successful recruitment (Hopper et al. 1990). The flora of the South West Botanical Province has the highest concentration of rare and threatened endemics in Australia and is exceptional from a global perspective (Hopper et al. 1990; Boden and Given 1995; Brown et al. 1998). Almost 75 percent of the taxa found in the South West Botanical Province are endemic to the region (Hopper et al. 1990), and more than 80 percent of the state’s threatened, rare, or poorly known taxa are found here. This diversity of species and high rate of endemism result from the major climatic changes the region experienced during the late Tertiary and Quaternary and from the isolation of the province for some 30 million years from its floral origins (Hopper et al. 1996). The southwest 40
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Figure 3.1 Western Australia and the species-rich South West Botanical Province, with Department of Conservation and Land Management Regional boundaries delineated. Inset: Australia.
flora has survived for a long time in the absence of any major extinction events and appears to be both ancient and remnant. Most threatened plants are both naturally localized and numerically rare (Hopper et al. 1990), rendering them more prone to extinction from local disturbances such as disease, drought, weed invasion, and accidental destruction. The majority have not been investigated for their economic or other uses, and the extinction of any one species would represent an irreplaceable lost opportunity for plant use, study, and appreciation (Hopper and Coates 1990; Armstrong and Abbott 1996).
Threatened Species Listing and Legislative Protection Priority setting for listing threatened species in Western Australia considers not only the implications of the biodiversity loss but also the poten-
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tial limitations in resources and cultural, scientific, and commercial values and legislative requirements (Coates and Atkins 2001). In 1999, 2,318 taxa were listed as rare, threatened, or poorly known and in need of further survey (Atkins 1999). Collectively these taxa are called conservation taxa. Only 334 of these taxa have legislative protection under the Western Australia Wildlife Conservation Act 1950 as Declared Rare Flora (threatened flora) (Wildlife Conservation Rare Flora Notice 1999); 102 of these threatened taxa are ranked under the International Union for Conservation of Nature (IUCN) Red List categories as Critically Endangered (IUCN 1994; Atkins 1999). The poorly known taxa are called Priority taxa for which further survey is needed to accurately assess their conservation status. These taxa are divided into four categories (1–4) depending on the number of known populations and the number of populations considered under immediate threat. This list provides a means of setting conservation and research priorities in Western Australia (Burgman et al. 2000). More than 50 threatened ecological communities have also been identified in the southwest of the state, with 21 of these considered highly threatened (English and Blyth 1999). The Western Australian State Department of Conservation and Land Management (CALM) is the statutory authority administering the Wildlife Conservation Act 1950 through the provisions of the Conservation and Land Management Act 1984. It plays a major role in managing large areas of land and water with nature conservation as a primary management objective. National parks, nature reserves, conservation parks, state forests, and timber reserves make up some 7.6 percent of the land area of Western Australia managed by the department. It is also responsible for recommending new areas to be placed in the conservation estate, preparing recovery plans for threatened taxa and ecological communities, curating the state’s flora collections, and conducting scientific research relating to the conservation and management of the state’s biodiversity.
The Impact of Invasion, Fragmentation, and Conversion on Biodiversity Many of the major threats to the Western Australia flora are a direct result of the extensive clearing and associated degradation of vegetation that have occurred over the past 200 years since European settlement. Some 72 per-
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cent of threatened flora populations and 56 percent of vegetation types identified in Western Australia are found outside conservation reserves (Coates and Atkins 1997). Much of the remaining vegetation occurs to a large extent along road verges and on private property. In the wheatbelt region of Western Australia alone, some 93 percent of the native vegetation has been cleared (Hobbs and Saunders 1994). As a consequence of land clearing and the resultant habitat fragmentation, large areas of Western Australia have been subjected to rising groundwater levels and salinization. The removal of deep-rooted native perennial vegetation and its replacement with shallow-rooted annual pasture has caused a fundamental change in the groundwater hydrology of the southwest (McFarlane et al. 1993; George et al. 1995). It is estimated that waterlogging and salinity may lead to the extinction of 450 taxa in the next one or two decades (G. Keighery, pers. comm., 2000). The arrival and spread of the root rot fungus disease Phytophthora cinnamomi, commonly called dieback disease, is now considered a biological disaster of global significance given the richness and high degree of endemism of the flora in southwestern Australia (Government of Western Australia 1998). The disease can significantly alter the floristic composition and modify the structural complexity of vegetation communities (Shearer and Dillon 1996). It is possible that some 20 percent or more of the native flora is susceptible to the disease, with genera in the Proteaceae, Epacridaceae, Fabaceae, and Myrtaceae considered the most susceptible (Malajczuk and Glenn 1981; Wills 1992). These families account for up to 50 percent of the species in many ecosystems of southwestern Australia (Burbidge 1993). Significant declines in plant populations have already occurred (Withers et al. 1994; Wills 1992) despite mitigation procedures involving the application of the fungicide phosphite, which stimulates the defense mechanisms in treated plants. There is no known method of eradicating the disease (Podger 1972; Shearer et al. 1991; Burbidge 1993), and the application of phosphite every few years is unlikely to constitute a viable long-term control measure. A number of ecological communities and plant species face rapid extinction unless phosphite treatment and ex situ conservation strategies are effectively sustained, along with hygiene and quarantine measures. Competition with invasive weeds has reduced the survival chances of a range of rare and threatened plant populations in the region, as has grazing by feral animals such as rabbits and introduced stock (Coates and
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Atkins 1997). The rapid decline in population size caused by habitat destruction and the isolation of many rare plants may render them susceptible to extinction when sudden, unexpected changes occur in the landscape. Loss of genetic diversity and inbreeding in these small populations are significant threats to both short- and long-term survival (Barrett and Kohn 1991; Falk 1992). In addition to habitat clearing, salinity, disease, and weed invasion, mining, quarrying, and accidental destruction by road maintenance activities can impact heavily on the natural vegetation. The long-term impact of climate change may affect species with narrow environmental tolerances (Government of Western Australia 1998). Changes in geographic distribution, reproduction, physiological response, and persistence of some species can be expected when changes in temperature and humidity occur as a result of climate change.
The Path to Recovery: Integrated Conservation Western Australia’s integrated strategy for the conservation of threatened flora involves a wide range of individuals, groups, and organizations, with CALM as the key player. The department’s framework for action focuses on a series of wildlife management and recovery plans that progress from regional and district threatened flora management programs to recovery plans for specific taxa and threatened ecological communities. Recovery plans seek to involve the community and other land managers in the conservation of threatened flora populations, with members of the recovery teams made up of local government, landowners, and community groups such as local wildflower societies, agency staff, and scientists with knowledge of flora conservation. The team works cooperatively to develop and implement the recovery actions needed to ensure species survival. Regional or district plans are written on a geographic area basis that are defined by the CALM district and regional boundaries. These area-based management plans represent a cost-effective use of limited resources (Brown et al. 1996). The area-based management plans provide the focus for further intensified survey, confirming conservation status and paving the way for single-species recovery plans. In many cases basic survey and inventory work is still needed to determine the presence and extent of natural populations and the habitats they occupy.
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The Western Australian Threatened Species and Communities Unit is responsible for developing and implementing recovery plans for threatened taxa and communities. Major management or research actions identified as highest priority to assist in the conservation and management of individual species and communities include germplasm collection and storage, liaison and education, monitoring, weed research and control, fire ecology, land acquisition, disturbance ecology, fencing, disease research and control, habitat restoration, translocation and recovery, and hydrology (Coates and Atkins 1997). CALM plays a major role in implementing recovery actions such as fencing, weed and disease control measures, and placement of roadside markers to mark rare plant populations and monitoring the health and demography of plant populations. Protocols for recovery are based on an understanding of the target taxa’s population dynamics, reproductive biology, and population genetics. Phylogenetic and molecular systematic data are also used, when appropriate, in the description, classification, and conservation of flora. The threats are identified and strategies developed for their control. Current information on the conservation status of rare and threatened flora is provided. The department is also responsible for a large part of the ex situ work being conducted for rare and threatened flora. In addition to the department’s involvement in Western Australia’s integrated conservation strategy, other agencies such as Perth’s Botanical Gardens and Parks Authority (formerly Kings Park and Botanic Gardens), environmental consultants, mining companies, and university groups also provide support for survey, research, and ex situ collections.
The Threatened Species Ex Situ Program Western Australia faces an enormous and urgent task of conserving some 2,000 rare, threatened, and poorly known taxa, with more than 100 considered on the brink of extinction in the next two decades. Some 16 percent of threatened species in Western Australia are known from only one population. In total, 64 percent are known from 1,000 or fewer individuals; 10 percent are known from fewer than 50 individuals (Atkins, pers. comm., 1999). Unfortunately, land acquisition, fencing, invasive weed control, and restoration projects are unlikely to overcome the insurmountable problems of disease and salinity faced by some of Western Australia’s most
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threatened flora. Ameliorating these threats will be an ongoing process, requiring millions of dollars and a highly motivated public effort. It is unlikely that this task will be achieved in the next 50 years. In all probability it may be impossible to prevent the loss of some species in the wild. Ex situ strategies may be the last hope for biodiversity conservation and for future restoration of many of these critically endangered taxa. At the national level there has been widespread recognition that unpredictable events can threaten rare species and that ex situ conservation can provide insurance under these circumstances (Armstrong 1991; Armstrong et al. 1993; Anonymous 1996). Ex situ conservation can provide critical support to in situ measures by improving the understanding of regeneration techniques, the safekeeping of genetic material, and the provision of that material for reintroduction programs. The ex situ program therefore is intended to complement in situ conservation rather than replace it, as a means to an end rather than an end in itself (BGCI and IUCN 1989). In late 1992 the state government of Western Australia established an ex situ storage facility for genetic material from rare and threatened native taxa. This facility focused initially on plant taxa threatened by Phytophthora cinnamomi but subsequently expanded to include taxa at risk from other threats. The ex situ facility was designed to use cryopreservation for the long-term storage of somatic tissue, but the estimated cost of accessioning such a large number of taxa into tissue culture for cryostorage was considered to be prohibitive and was judged not achievable in the short term. On advice from staff of the Seed Bank, Royal Botanic Gardens Kew, at Wakehurst Place, it was considered more cost-effective to store seedbased material under conventional gene bank conditions (low temperature and low moisture). The majority of species from southwestern Australia produce orthodox rather than recalcitrant seeds, capable of being desiccated and frozen, without loss of viability. The technology for long-term storage of orthodox seed is well advanced (Roberts 1991) and is simple, inexpensive, and applicable over a wide range of species. In addition, whole seed conservation has many advantages over other means of gene storage (e.g., pollen and tissue). Seeds may have wider genetic representation than vegetative material, seeds are immediately available for seedling production, and seed is less expensive to store than pollen or tissue samples. Prioritization of species for ex situ conservation is based primarily on the level of threat. Where species are known from many populations, ex situ conservation methods can effec-
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tively sample flora populations over the full range of their distribution. Numerous small populations may be challenging to preserve on site, whereas the sampling and storage of propagules over many populations of a species are reasonably easily undertaken. The state facility, the Threatened Flora Seed Centre (TFSC), had two primary objectives: to develop a comprehensive seed-based germplasm collection for rare and threatened plant taxa in Western Australia with the initial aim of capturing 75–80 percent of all genetic variation in each taxon and to use appropriate protocols for the medium- and long-term storage of seed from rare and threatened plant taxa in Western Australia and maintain an integrated database on seed provenance and seed biology for each taxon (CALM 1995). This strategy was to result in the storage of sufficient genetic resources from each threatened taxon to ensure its successful reintroduction to the wild after extinction of wild populations. Material would be available for biochemical, physiological, and molecular research, and material could be provided for ex situ propagation as needed for recovery programs and educational purposes. A series of guidelines and standards for collection, storage, monitoring, and documentation of germplasm have been developed in Western Australia to ensure that the highest-quality genetic material is acquired and maintained. These guidelines are based on those adopted in Australia by the Australian Network for Plant Conservation (Touchell et al. 1997). These are derived from published international standards (Hanson 1985; Ellis et al. 1985; Brown and Briggs 1991; Wieland 1995; Smith and Linington 1997). Guidelines for quality assessment, quantification, and germination testing have been developed by the TFSC on a species-by-species basis following Hanson (1985) and Touchell et al. (1997). There is no definitive method for sampling and handling of all species; for instance, defining an adequate sample will vary between populations and species and will depend on the extent and distribution of genetic variation within a species as well as the biology, ecology, and longevity of the species. In Western Australia many species have seeds that need multiple cues to stimulate germination and overcome dormancy. These include the application of smoke (Dixon et al. 1995; Roche et al. 1997b, 1998; Tieu et al. 1990), alternating temperatures (Bellairs and Bell 1990; Bell and Bellairs 1992; Bell et al. 1993; Bell 1999), heat stratification (Bell and Williams 1998), seed aging methods such as seed coat removal, leaching, and scarification (Schatral 1996; Schatral et al. 1997; Roche et al. 1997a; Cochrane
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the scope and potential of ex situ plant conservation
et al. 1999), and the use of growth hormones (Bell et al. 1995). For many species these dormancy mechanisms are still poorly understood. The capacity to maintain genetic variation in seed storage facilities for future reintroductions depends on the effective maintenance of germplasm viability in storage. Assessing the response of seed to storage conditions over the long term is hampered by a poor understanding of the complex dormancy mechanisms that prevent seed from germinating under standard laboratory conditions of temperature, light, and moisture. In addition, differential germination and the survival of individual propagules or genotypes may inadvertently reduce the genetic variability in an accession and select for genotypes that are well adapted to seed storage conditions but poorly adapted to survival in wild conditions. Where dormancy mechanisms are poorly understood, there is the likelihood that there will be selection pressures on seed with low levels of dormancy. Assessment of genetic variation using seedling material from a collection, in which only proportions of propagules germinate, will not demonstrate the true representativeness of that collection. This problem is inevitable until a clear understanding of the cues needed to germinate all seed in a collection is available. The problem pertains not only to seed germination but also to the continued survival of that seed to maturity (Figure 3.2). A recent assessment of the germinability of more than 200 taxa held in ex situ storage at the TFSC demonstrated that the majority of seed from rare and threatened taxa collected in southwestern Australia needed multiple cues to stimulate germination (Cochrane et al. 2002). In addition to complex dormancy mechanisms, seed abortion and parthenocarpy are widespread in many genera in southwestern Australia. Despite the difficulties of poorly understood germination needs, high levels of seed abortion, and problems with assessing numbers of propagules held, the TFSC has demonstrated that seed from a wide range of families in Western Australia have the ability to retain viability under gene bank conditions (Cochrane and Kelly 1996). More than 630 collections from more than 215 rare and threatened taxa have been accessioned since the inception of the gene bank facility. Eighteen families and 49 genera are represented in these collections. More than 65 percent of these taxa are threatened, representing some 45 percent of the total number of threatened taxa known from Western Australia. Not only is the facility providing an ex situ service, but its work is adding to a broader understanding of the biology of a range of threatened native species.
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Figure 3.2 Mortality rate over growth stages for Lambertia echinata ssp. occidentalis during experimental translocation.
The Role of Collaborating Agencies, Groups, and Individuals In addition to displays of threatened flora, the Botanic Gardens and Park Authority (BGPA) in Perth is highly involved in the conservation of rare and threatened Western Australian plants. Over the past few years the TFSC and BGPA have coordinated seed collections to improve the effectiveness of the ex situ program. Duplicate collections of seed of many of the most highly threatened species are held in both centers for safe storage. In the event of a disaster at one of the facilities, germplasm for those replicated taxa will not be lost (Given 1994). In addition to seed collections, material for tissue culture and nursery propagation is also made. The majority of material for reintroduction in Western Australia is propagated in the BGPA nursery from cuttings or in vitro cultured material or from germinated seed provided by the TFSC facility (see “Case Studies” later in this chapter). In addition, the Rare and Threatened Garden at BGPA provides a good opportunity for public education about the state’s rare and threatened flora. A number of local government authorities throughout Western Australia have established small rare flora gardens in their country towns to
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promote awareness of the state’s unique flora. Local school groups often maintain these gardens, and in some areas the local government shires have adopted a special threatened species as their local emblem. This form of community involvement plays a vital role in public education on conservation issues. Nongovernment organizations are becoming increasingly active in the conservation of threatened species in Australia. For example, the World Wide Fund for Nature Australia (WWF) is making a sizable contribution to funding conservation programs nationally. In Western Australia the ex situ program has been assisted financially with research on the reproductive capability of plants threatened by herbivore grazing and the establishment of traps for seed collection. Funding has also been provided to assist with recovery work on highly threatened taxa. In all cases, the Australian Trust for Conservation Volunteers, a nonprofit conservation organization, have been engaged to carry out the work. Projects such as these ensure that the community becomes aware of the need for ex situ programs and therefore provides the additional support needed for agency conservation work. A number of projects administered through the nongovernment organization Greening Australia (Western Australia) are contributing to ex situ biodiversity conservation through community participation in the collection and storage of native seed for broad-scale revegetation projects. The nursery industry in Western Australia provides limited input into the conservation of rare and threatened plants. Only a few nurseries specialize in native plants, and even fewer give priority to endemic species from the region. Seeds of many native species are widely available, but many species are hard to germinate. Some seed merchants have seeds from a number of the state’s rare and threatened species for sale, although many seed merchants do not specify provenance details. In most cases these seeds are unlikely to be useful for conservation and species recovery because of their indeterminate origins.
Vital Links: The Ex Situ Program and Conservation Biology Research For conservation measures to be effective, it is imperative to conduct research that is aimed at understanding the nature and dynamics of the biological systems under threat. Because of the increasing speed at which species are being lost, it is essential that research has an immediate rele-
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vance to the management of conservation taxa. In Western Australia this research is being undertaken by government agencies and in some cases by universities through scholarship programs and collaborative projects as well as by private environmental consultants. Industry, particularly mining companies that have a vested interest in the distribution, management, and protection of threatened flora communities on their leases, is also funding research into threatened taxa in Western Australia. Establishing effective ex situ collections requires an understanding of the patterns of genetic variation in the target species to guide the strategy for sampling that variation. In most cases there are insufficient resources to sample all known populations, and genetic data can assist in the development of appropriate sampling protocols. Ignoring population-based variation, and thus the evolutionary and ecological processes associated with the generation and maintenance of this variation, will lead to loss of evolutionary lineages that may be as unique as taxonomic entities (Coates 2000). Such studies in conservation genetics have provided the basis for sound sampling strategies to be formulated for a number of threatened taxa, thereby providing enormous benefits to the ex situ program. Research into population genetic structure in Western Australia over the past decade has made important contributions to the conservation and management of rare and threatened taxa (Hopper et al. 1990; Hopper 1998). Limited genetic analyses of native species have found that there is greater variability between populations than within populations for many rare and threatened species (for example, see Moran and Hopper 1987; Sampson et al. 1988; Coates and Hamley 1999; Coates 2000). These data indicate that a suitable strategy to capture the greatest range of genetic variation for a taxon would be to collect few individuals from many populations rather than many individuals from few populations. Using ex situ material for research can make important contributions to the understanding of the reproductive biology of threatened taxa. For instance, research into the germination needs of native taxa has broadened the knowledge of the seed biology of Western Australia’s native flora (see Bell 1999 for a review of recent literature). In particular, studies of smokestimulated germination have revolutionized the propagation (Dixon et al. 1995; Roche et al. 1997b, 1998) and recovery of threatened species (Rossetto and Dixon 1998). Reproductive biology and ecological studies for a number of highly threatened native taxa have been undertaken in the past few years (e.g., Cochrane et al. 2000). Analysis of soil seed banks, pollina-
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tion mechanisms, reproductive strategies, and demography play an important role in understanding vegetation dynamics and can provide information vital for the recovery process. An understanding of the impact of disease, most notably Phytophthora cinnamomi, on threatened plant species is also a vital part of Western Australia’s conservation effort. The reintroduction of species to the wild depends on knowledge of the susceptibility of those species to the disease and the level of disease infection at the site. For the past 4 years the ex situ program has provided material for laboratory investigation into the susceptibility of rare and threatened flora to the pathogen Phytophthora cinnamomi (Shearer, pers. comm., 1999).
The Challenge: Species Reintroduction For a successful reintroduction it is necessary to use the best possible tools to improve the success of a reintroduction within the limitations of resources and time (Pavlik 1996). Unfortunately, a lack of funding for reintroduction projects often precludes extensive scientific research, and the lack of scientific input into reintroduction programs has been criticized (Sarrazin and Barbault 1996). In Western Australia, species reintroductions have been undertaken that take advantage of research into rare and threatened native taxa (Monks and Coates 1999). The following case studies illustrate the interplay of ex situ collections, research, and management in a bid to accomplish an effective recovery strategy for the survival of threatened species. This involves identifying and addressing threats, determining systematic relationships and population genetic structure, and understanding demographics, biology, ecology, and reproductive and pollination mechanisms (Guerrant 1996; Australian Network for Plant Conservation 1997; IUCN 2000). These case studies demonstrate the importance of a working gene bank for collecting and maintaining genetically representative seed material held for species reintroduction. They confirm the significance of genetic and demographic characteristics of the founding populations. This is particularly important in Western Australia, where there is a high level of intraspecific genetic differentiation in many geographically limited and threatened species. These case studies also show how the ex situ facilities form an integral part of the recovery process by providing expertise for further testing of the reproductive biology and recruitment potential of the newly established populations. Recovery
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actions such as species reintroductions entail long-term commitment, particularly for longer-lived woody perennial species that may not reach reproductive maturity before 10 years. Recovery planning and demographic monitoring of these reintroductions will be an ongoing responsibility.
Case Studies Lambertia orbifolia (Proteaceae) L. orbifolia is a threatened attractive, erect shrub or small tree up to 4 m tall with four to six orange-red tubular flowers forming in the upper leaf axils. The species is known from several populations in two disjunct locations some 200 km apart (see Figure 3.1). The coastal and semicoastal populations on the Scott River Plains number in the thousands, whereas the inland populations at Narrikup consist of 169 plants. Plants are killed outright by fire, regenerate from seed, and are highly susceptible to Phytophthora cinnamomi. The Narrikup populations occur on degraded road verges and are in poor condition, affected by disease (aerial canker and P. cinnamomi) and exotic weed invasion. The coastal and near-coastal populations occur on private lands in large tracts of mainly healthy remnant vegetation. Substantial management-related research has been conducted on L. orbifolia. Data have been gathered on the reproductive biology of the species (Sage 1994), pollination biology studies were conducted as part of a 1-year student program (Whittaker and Collins 1997), and genetic studies have been undertaken (Cochrane, pers. comm., 1999). The use of molecular markers indicated a high level of genetic diversity, with marked differences between the two disjunct population groups (Coates and Hamley 1999; Byrne et al. 1999); these results warranted the management of the distinct Narrikup form of the species as a separate conservation unit and the writing of a recovery plan for this form. The results also indicated that there was a suitably broad range of genetic diversity in the small Narrikup populations to support successful translocation and population enhancement programs (Coates and Hamley 1999). Because of high levels of genetic diversity between the two population groups and little diversity within the population groups, seeds from several populations in both the Narrikup and Scott River Plains groups were sampled for ex situ storage at the TFSC. These collections are designed to hold a large proportion of the genetic variability of the species. Initial viability
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of seed was high (up to 100 percent), with only one early collection showing a decrease in viability in storage over 5 years. In 1998 plants of L. orbifolia were translocated to a conservation reserve less than 4 km from the known Narrikup populations (Coates et al. 1998). Three main criteria were considered in site selection: proximity to the known populations, security of tenure, and similarity of biotic and abiotic features to the known populations. Seed collected from the Narrikup populations between 1993 and 1997 and stored ex situ at the TFSC was germinated in 1997–1998. Seeds from both Narrikup populations were mixed because of low diversity between populations, guaranteeing a greater number of source plants for the translocated population. Seedlings were transferred to BGPA for cultivation, and 216 six-month-old seedlings were planted in five grids in winter 1998. Seedlings were mulched, shaded, or protected by “gro-cones” (plastic enclosures used for wind and predator protection) or left untreated. Plants were fenced to prevent herbivore predation. After 12 months in the ground some 98 percent of the original plants have survived (Monks, pers. comm., 1999). Bimonthly monitoring of the original populations and the new population includes monitoring of flower and fruit production, number of surviving seedlings, growth measurements, reproductive status, regeneration, and plant health. Additional seed collections from the new populations will be made over the next few years, and research on their viability and reproductive output will be included in the monitoring program (Monks and Cochrane, pers. comm., 1999). In this case study adequate amounts of viable seeds were available from a large number of individuals from the Narrikup populations for ex situ storage and recovery. There was high survivorship during seedling growth and subsequent transplanting, ensuring that the translocated population is as genetically representative of the Narrikup area as possible. Preparatory research into the demographics, ecology, and biology of the species ensured that recovery work was based on sound scientific principles. The genetic differences between the Narrikup and Scott River Plains population groups stressed the need to keep germplasm from the different groups separate and to use seeds only from the Narrikup form for that reintroduction. Hopes for the survival of this translocation into a selfsustaining population are high. Demographic monitoring and careful evaluation over the next few years will be needed to determine the success of the project. High levels of hygiene around the translocation site must be maintained to ensure that disease does not affect the new population.
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Rulingia sp. Trigwell Bridge (Sterculiaceae) Rulingia sp. Trigwell Bridge is a small, hairy-leaved perennial shrub up to 1.5 m tall and 1 m wide bearing terminal inflorescences of creamy-white flowers. It was first discovered in 1970 and is known from only one natural population of three plants (one adult and two juvenile plants) on private property. The plants grow in rock fissures on a lateritic ridge in association with open, low jarrah (Eucalyptus marginata) and marri (Corymbia calophylla) some 200 km southeast of Perth (see Figure 3.1). Because of its restricted nature and the small numbers of plants, Rulingia sp. Trigwell Bridge was listed by the Department of Conservation and Land Management as threatened and ranked by the department’s Western Australian Threatened Species and Communities Unit as Critically Endangered (sensu IUCN). This species is threatened by weed invasion, grazing, inappropriate fire regime, and lack of natural recruitment (Stack et al. 1999). One of the first recovery actions to be completed for this species was the fencing of the remaining few surviving plants in 1994. In the same year, seeds were sent to CALM-TFSC and to Perth’s Botanic Gardens and Parks Authority for propagation trials and ex situ conservation. Germination trials established that this species responded well to heat treatment and scarification of the hard seedcoat, with high seed viability for both freshly collected and stored seed (Cochrane, pers. comm., 1994). In 1995 experimental smoke trials were conducted in situ to stimulate the germination of seed in the soil seed bank, but with no success (Williams and Fitzgerald 1998). In the same year, seed and vegetative material were again collected by BGPA to ensure that sufficient material was available for the recovery of the species. The absence of Phytophthora cinnamomi at the site was confirmed by soil analyses, and plant inoculation trials on propagated seedlings established that Rulingia sp. Trigwell Bridge was not susceptible to the dieback disease (Shearer, pers. comm., 1998). In 1995 a recovery team was appointed to assist with the implementation, development, and coordination of recovery activities. In 1997 BGPA undertook research into a range of propagation techniques including micropropagation, in vitro physiology, slow growth, and cryostorage (Williams and Fitzgerald 1998). Reintroduction of Rulingia sp. Trigwell Bridge began with the development of a translocation proposal in 1997 (Fitzgerald and Williams 1997). Demographic monitoring of translocated seedlings demonstrated
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high survivorship, vigorous growth, and the production of large quantities of viable seeds (Fitzgerald, pers. comm., 1998). Plants were noted to flower and fruit within their second year of growth after propagation from seed. The TFSC has conducted germination trials on second-generation seeds and has more than 10,000 seeds stored under low-temperature and lowmoisture conditions (Cochrane, pers. comm., 1999). Despite large quantities of seeds being produced from the transplanted seedlings, there has been no demonstration of recruitment from these new populations, and direct seeding trials under experimental conditions using stored seeds have commenced (Fitzgerald, pers. comm., 1999). Additional work on seed dispersal and predation, fire response, and recruitment will be needed. Continued monitoring of health and dynamics of these populations will determine their ability to persist over the long term. This information will help the recovery team implement further actions to rescue the species. Unlike many critically endangered southwestern taxa, Rulingia sp. Trigwell Bridge flowers and sets seed at an early stage in its life cycle. The high fecundity of the species enables large quantities of seeds to be held ex situ and an assessment of seed viability from cultivated plants to be made. Large quantities of seeds can also be used for direct seeding trials more closely emulating natural recruitment events. On-site research into the effects of a variety of different disturbance events may also be conducted using the large numbers of available seeds. Soil seed bank and postdispersal predation studies will be possible and should provide important data to the recovery team for ongoing species management. This case study is an example of a species that is highly fecund and grows readily in cultivation and when reintroduced into the wild but demonstrates minimal recruitment in its natural environment despite the production of large quantities of highly viable seed.
Verticordia fimbrilepis subsp. fimbrilepis (Myrtaceae) V. fimbrilepis subsp. fimbrilepis is an erect shrub up to 60 cm high with slender branches and clusters of purplish-pink flowers. Possibly because of extensive habitat clearing for agricultural purposes, this taxon is limited to several small, disjunct road verge populations near the town of Narrogin in the southern wheatbelt of Western Australia (see Figure 3.1). Additional populations are located some 180 km northeast of these near the town of Aldersyde. These northern populations may be genetically distinct, but
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until recently, insufficient material has been available from the small southern populations for ex situ conservation or genetic analysis. Potential threats to these populations include road maintenance, weed invasion, rising salinity, disease, inappropriate fire regimes, and feral animal pests (Mitchell et al. 1997). The CALM Interim Recovery Plan for V. fimbrilepis subsp. fimbrilepis recommended translocation to a secure site to overcome the threats associated with small population size (Mitchell et al. 1997). To support the translocation it was planned to establish populations in controlled cultivation and to test establishment techniques in order to improve the success of the translocation (Graham and Mitchell 1998). Plants were propagated vegetatively at BGPA using wild material from the five remaining individuals in this population. One hundred fifty plants were established in late July 1998 in trial plots under three treatment conditions (control, mulched, and watered). Plants were monitored bimonthly. By May 1999, 37 percent of plants had died (Graham, pers. comm., 1999). In February 1999 the TFSC collected fruits from all living plants at the new translocated population. Reproductive success in terms of fruit-to-seed (seed set) and seed-to-seedling (germination) ratios was highly variable between clones and between treatments. Previous investigation by the TFSC into the seed set of this taxon indicated that these results were higher than expected in the wild. The exposed block nature of planting at the translocation site may have contributed to the high fruit-to-seed ratio through more effective pollination. The translocation and the associated research into reproductive biology were designed to assist management in planning translocation projects. The initial investigations into seed set and germination highlighted the dangers of using small amounts of vegetatively propagated material representing single clones. The regeneration potential of the population depends not only on the seed set and germinability of propagules but also on the proportions of germinable seed successfully attaining adulthood and reproductive status. Small samples are rarely representative of the population from which they are drawn (Barrett and Kohn 1991), and if sampling effects become cumulative this could lead to the loss of alleles, resulting in genetic drift. The deleterious effects of small sample sizes may be manifested in successive generations. This population of V. fimbrilepis subsp. fimbrilepis does not appear to be suffering reproductive difficulties, but the interclonal differences in seed set and germination at this translo-
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cation site may indicate the need for genetic augmentation. It will be necessary to closely examine the fitness, growth rates, and survivorship of offspring for indications of inbreeding depression in future generations.
Grevillea scapigera (Proteaceae) G. scapigera is a prostrate woody perennial up to 40 cm tall and up to 2 m wide with slightly hairy stems, glabrous, slightly pungent leaves, and scented, creamy white flowers. Genetic studies suggest that this species is a relatively restricted endemic and not closely related to any other member of the genus (Rossetto and Dixon 1998). It is an outcrossing, short-lived (ca. 9 years) disturbance opportunist recruiting from seed (Rossetto et al. 1998). Populations are confined within a 40-km radius of the wheatbelt town of Corrigin, some 250 km southeast of Perth (see Figure 3.1). The absence of large or contiguous reserves in the current distribution range of the species is considered a major limiting factor for long-term survival of the species (Rossetto and Dixon 1998). G. scapigera was first discovered in 1954, but by 1980 the species was considered to be in a state of rapid decline with no indication of recruitment (Rossetto et al. 1998). It was thought to be extinct when the last known wild plant died in 1986. The species was afforded legal protection in 1987. In 1989 material from a grafted plant located at the Royal Botanic Gardens in Sydney, Australia, was obtained and tissue cultured at BGPA. In the same year another 13 plants were found in three different locations. By 1991 one of these populations was dead, and by 1992 there was only one remaining plant in each of the other two populations. Increased survey efforts located a total of 28 additional plants in four populations by 1993, although all were on highly vulnerable road verges. The number of known plants has never exceeded 50. Most plants are now mature and reaching senescence. Natural recruitment appears to be occurring only sporadically. In addition, seeds are thought to be subject to high levels of predation by parrots and seed-eating insects (Rossetto et al. 1998). Weed invasion is also a problem, and herbicide control to suppress weed growth and encourage regrowth of native species has been conducted. The effects of salinization and disease on the species are unknown. A recovery team was established to oversee the recovery efforts, which aimed to manage the existing populations, propagate plants through in vitro propagation, store germplasm in cryopreservation, and conduct research into genetics, seed biology, and recruitment. Ex situ procedures were developed
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by BGPA for propagation and somatic tissue storage of the species including in vitro micropropagation, cell suspension and callus culture, somatic embryogenesis, and cryopreservation of shoots and seeds (Rossetto and Dixon 1998). These procedures provided a rapid means for large-scale production and protection of germplasm until on-site recovery efforts could be initiated. In 1996 a translocation project was planned based on an understanding of the genetics and biology of the species, with three trial plantings in situ (1993 to 1995) (Rossetto et al. 1998). Because most existing plants were located on narrow road verges, alternative sites for recovery were needed. Material was translocated into secure reserves under CALM authority or in the Corrigin Land Conservation District. Five different genotypes were planted in Corrigin Reserve in 1996, and another five were planted in the following year (259 plants) to maximize the chances for outcrossing between genotypes. In 1998 an additional 282 plants representing seven different clones were planted at the same site. Survival rates have ranged from 3 to 80 percent (Rossetto and Dixon 1998). In vitro stock was acclimatized in the nursery before planting, and plants flowered in the first spring after planting. G. scapigera provides an excellent example of the importance of holding ex situ collections when the survival of naturally occurring populations is at risk. Disastrous events such as locust plagues, roadside maintenance, and natural senescence combine to threaten the small populations. Ex situ collections have given this species a chance for survival and provided the means to conduct research into the genetic variability of the species using random amplified polymorphic DNA markers (RAPDs) (Rossetto et al. 1995). With potentially low survival levels from reintroduction programs, it is vital that ex situ collections be held in case of recurring failure. The key clones used in the recovery operations are being maintained as active tissue cultures or held in cool storage, and the BGPA nursery holds healthy stock plants of some clones for propagation and research. Ten clones representing 85 percent of the measured genetic diversity for the species have been successfully cryostored, and in 1998 and 1999 seeds from transplanted material were collected for ex situ conservation under standard gene bank conditions at the CALM TFSC and at BGPA.
Conclusions These case studies have demonstrated the means by which Western Australia’s ex situ program provides vital links between on-site management
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and off-site research. The program contributes material for education and for research into molecular systematics, population genetics, and the impact of disease. Information on population genetics assists in the formulation of sampling strategies undertaken in the ex situ program and facilitates recovery. Knowledge of disease susceptibility helps in prioritizing endangerment, which in turn determines which species are targeted for collection. The ex situ program provides a means for investigating the reproductive biology of threatened plant species. Research into reproductive biology provides a vital understanding of the nature of growth and recruitment in threatened populations, helping to ensure a successful recovery. Long-term research into the physiological changes in seed viability during storage and germination and monitoring of that seed over time is a prerequisite for the ex situ program. And for some threatened plants, an ex situ program is the key to survival. In Western Australia cooperative management programs involve a variety of government institutions, including state government agencies such as the Department of Conservation and Land Management, Botanic Gardens and Parks Authority, Main Roads and Westrail, local government authorities, community groups, and individual landowners (Blyth et al. 1995). Liaison between these groups and close collaboration occurs through recovery teams who administer the recovery plans. This cooperation is central to the management of threatened flora and to its conservation. Funding for projects comes from federal, state, and nongovernment organizations, but it is essential that agencies with continuity of existence and funding address long-term conservation issues (Burbidge 1993; Morse 1993). In Western Australia a rapidly developing knowledge of the biota and an understanding of the threats to its survival promise more effective conservation. Despite a lack of resources and a limited funding base, ex situ conservation has played and will continue to play a major role in the conservation of threatened flora in the state. Careful allocation of limited resources is vital. An ex situ program offers flexibility, holding genetic material in long-term storage until a decision can be made as to the optimum allocation of resources. Decisions can be made on whether to direct scarce resources to the most highly threatened species or to those most likely to benefit from intervention. Some species are so close to extinction in the wild that the enormous investment of resources needed for their recovery may be better spent on other taxa less severely reduced and with some hope of protection.
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Acknowledgments I am deeply indebted to many people in the Department of Conservation and Land Management for their contributions to this chapter, particularly Dr. David Coates, team leader of the Flora Recovery Project. Between 1992 and 1998 the Threatened Flora Seed Centre was funded by the federal government’s Threatened Species and Communities Section, Biodiversity Group, Environment Australia. Environment Australia has also funded much of the research on conservation biology described in this chapter. I would also like to thank the staff at Perth’s Botanic Gardens and Parks Authority and the many individuals, community groups, and local agencies that have contributed to the ex situ program in Western Australia. References Anonymous. 1996. The National Strategy for the Conservation of Australia’s Biological Diversity. Canberra: Commonwealth Department of the Environment, Sport and Territories. Armstrong, J. A. 1991. Genebanks or genemorgues? The need for a national plant germplasm program. Pages 207–210 in G. Butler, L. Meredith, and M. Richardson (eds.), Conservation of Rare and Threatened Plants in Australasia. Canberra: Australian National Botanic Gardens/Australian National Parks and Wildlife Service. Armstrong, J. A., and I. Abbott. 1996. Sustainable conservation: a practical approach to conserving biodiversity in Western Australia. Pages 21–29 in G. C. Grigg, P. T. Hale, and D. Lunney (eds.), Conservation Through Sustainable Uses of Wildlife. Brisbane: Centre for Conservation Biology, University of Queensland. Armstrong, J. A., N. Gibson, F. Howe, and B. Porter. 1993. The role of ex situ conservation. Pages 353–357 in C. Moritz and J. Kikkawa (eds.), Conservation Biology in Australia and Oceania. Chipping Norton, Australia: Surrey Beatty & Sons. Atkins, K. J. 1999. Declared Rare and Priority Flora List for Western Australia. Perth: Department of Conservation and Land Management. Australian Network for Plant Conservation. 1997. Guidelines for the Translocation of Threatened Plants in Australia. Canberra: Australian Network for Plant Conservation. Barrett, S. C. H., and J. R. Kohn. 1991. Genetic and evolutionary consequences of small population size in plants: implications for conservation. Pages 3–30 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Bell, D. T. 1999. The process of germination in Australian native species. Australian Journal of Botany 47(4):475–517. Bell, D. T., and S. M. Bellairs. 1992. Effects of temperature on the germination of selected Australian native species used in the rehabilitation of bauxite mining disturbance in Western Australia. Seed Science and Technology 20:47–55. Bell, D. T., J. A. Plummer, and S. K. Taylor. 1993. Seed Germination Ecology in Southwestern Western Australia. The Botanical Review 59(1):25–73. Bell, D. T., D. P. Rokich, C. J. McChesney, and J. A. Plummer. 1995. Effects of tem-
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perature, light and gibberellic acid on the germination of seeds of 43 species native to Western Australia. Journal of Vegetation Science 6:797–806. Bell, T. B., and D. S. Williams. 1998. Tolerance of thermal shock in seeds. Australian Journal of Botany 46:221–233. Bellairs, S. M., and D. T. Bell. 1990. Temperature effects on the seed germination of ten kwongan species from Eneabba, Western Australia. Australian Journal of Botany 38:451–458. BGCI and IUCN (Botanic Gardens Conservation International and International Union for the Conservation of Nature). 1989. The Botanic Gardens Conservation Strategy. Gland, Switzerland: IUCN. Blyth, J. D., A. A. Burbidge, and A. P. Brown. 1995. Achieving cooperation between government agencies and the community for nature conservation, with examples from the recovery of threatened species and ecological communities. Pages 343–357 in D. A. Saunders, J. L. Craig, and E. M. Matiske (eds.), The Role of Networks. Chipping Norton, Australia: Surrey Beatty & Sons. Boden, R., and D. Given. 1995. South-West Botanical Province Western Australia, Australia. Pages 484–489 in S. D. Davis, V. H. Heywood, and A. C. Hamilton (eds.), Centres of Plant Diversity. A Guide and Strategy for Their Conservation. Vol. 2. Asia, Australasia and the Pacific. Cambridge, UK: WWF and IUCN. Brown, A. H. D., and J. D. Briggs. 1991. Sampling strategies for genetic variation in exsitu collections of endangered plant species. Pages 99–119 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Brown, A., D. Coates, M. Fitzgerald, and C. Welbon. 1996. An area-based multiple species approach to threatened flora conservation and management in the Merredin area of Western Australia. Pages 130–137 in S. Stephens and S. Maxwell (eds.), Back from the Brink. Refining the Threatened Species Recovery Process. Sydney: Surrey Beatty & Sons. Brown, A., C. Thomson-Dans, and N. Marchant (eds.). 1998. Western Australia’s Threatened Flora. Perth: Department of Conservation and Land Management. Burbidge, A. A. 1993. Conservation biology in Australia: where should it be heading, will it be applied? Pages 27–37 in C. Moritz and J. Kikkawa (eds.), Conservation Biology in Australia and Oceania. Chipping Norton, Australia: Surrey Beatty & Sons. Burgman, M., B. R. Maslin, D. Andrewartha, M. R. Keatley, C. Boak, and M. McCarthy. 2000. Inferring threat from scientific collections: power tests and an application to Western Australian Acacia species. Pages 7–26 in S. Ferson and M. Burgman (eds.), Quantitative Methods for Conservation Biology. Berlin: SpringerVerlag. Byrne, M., B. Macdonald, and D. J. Coates. 1999. Divergence in the chloroplast genome and nuclear rDNA of the rare Western Australian plant, Lambertia orbifolia Gardner (Proteaceae). Molecular Ecology 8:1789–1796. CALM (Department of Conservation and Land Management). 1995. Science and Information Division Strategic Plan 1995–1999. Perth: Department of Conservation and Land Management. Coates, D. J. 2000. Defining conservation units in a rich and fragmented flora: implications for the management of genetic resources and evolutionary processes in south-west Australian plants. Australian Journal of Botany 48(3):329–339.
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Coates, D. J., and K. A. Atkins. 1997. Threatened flora of Western Australia: a focus for conservation outside reserves. Pages 432–441 in P. Hale and D. Lamb (eds.), Conservation Outside Nature Reserves. Brisbane: Centre for Conservation Biology, The University of Queensland. Coates, D. J., and K. A. Atkins. 2001. Priority setting and the conservation of Western Australia’s diverse and highly endemic flora. Biological Conservation 97(2):251–263. Coates, D. J., and V. L. Hamley. 1999. Genetic divergence and the mating system in the endangered and geographically restricted species, Lambertia orbifolia Gardner (Proteaceae). Heredity 83:418–427. Coates, D. J., L. Monks, and E. Hickman. 1998. Translocation Proposal: Round Leaf Honeysuckle, Lambertia orbifolia C. A. Gardner (Proteaceae). Perth: Department of Conservation and Land Management. Cochrane, A., K. Brown, N. Meeson, and C. Harding. 1999. The germination requirements of Hemigenia exilis (S. Moore) (Lamiaceae): seed plug removal and gibberellic acid as a successful technique to break dormancy in an arid zone shrub from Western Australia. Calmscience 3(1):21–30. Cochrane, A., S. Cunneen, and C. Yates. 2000. Population Structure, Soil Seed Bank Dynamics, Germination Requirements and Fire Response of the Critically Endangered Cyphanthera odgersii (F. Muell.) Haegi Subspecies occidentalis Haegi (Solanaceae). Unpublished report to the Department of Conservation and Land Management, Perth. Cochrane, A., and A. Kelly. 1996. Germination and storage of seed from rare and threatened plants of the south-west of Western Australia. Native seed biology for revegetation. Pages 101–106 in S. M. Bellairs and J. M. Osborne (eds.), Proceedings of the Second Australian Workshop on Native Seed Biology. Brisbane: Australian Centre for Minesite Rehabilitation Research. Cochrane, A., A. Kelly, K. Brown, and S. Cunneen. 2002. Relationships between seed germination requirements and ecophysiological characteristics aid the recovery of threatened native plant species in Western Australia. Ecological Management and Restoration 3(1):47–60. Dixon, K. W., S. Roche, and J. S. Pate. 1995. The promotive effect of smoke derived from burnt native vegetation on seed germination of Western Australian plants. Oecologia 101:185–192. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1985. Handbook of Seed Technology for Genebanks. Principles and Methodology. Rome: International Board for Plant Genetic Resources. Ellis, R. H., T. D. 1991. Seed moisture content, storage, viability and vigour. Seed Science Research 1:275–279. English, V., and J. Blyth. 1999. Development and application of procedures to identify and conserve threatened ecological communities in the South-West Botanical Province of Western Australia. Pacific Conservation Biology 5:124–138. Falk, D. A. 1992. From conservation biology to conservation practice: strategies for protecting plant diversity. Pages 397–431 in P. L. Fiedler and S. K. Jain (eds.), Conservation Biology. New York: Chapman & Hall. Fitzgerald, R., and K. Williams. 1997. Rulingia sp. (Trigwell Bridge) Translocation Proposal: Supplement Existing Population with Introduction of Additional Plants to Site and Establish New Population to Trigwell Nature Reserve. Perth: Department of Conservation and Land Management.
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George, R. J., D. J. McFarlane, and R. J. Speed. 1995. The consequences of a changing hydrologic environment for native vegetation in southwestern Australia. Pages 9–22 in D. A. Saunders, J. L. Craig, and E. M. Mattiske (eds.), Nature Conservation 4: The Role of Networks Volume 2. Chipping Norton, Australia: Surrey Beatty & Sons. Given, D. R. 1994. Principles and Practice of Plant Conservation. London: Chapman & Hall. Government of Western Australia. 1998. Environment Western Australia 1998: State of the Environment Report. Perth: Department of Environmental Protection. Graham, M. S., and M. D. Mitchell. 1998. Translocation Proposal: Shy Feather Flower, Verticordia fimbrilepis subspecies fimbrilepis. Perth: Department of Conservation and Land Management. Guerrant, E. O. Jr. 1996. Designing populations for reintroduction: demographic opportunities, horticultural options and the maintenance of genetic diversity. Pages 171–208 in D. A. Falk, C. I. Millar, and M. Olwell (eds.), Restoring Diversity: Strategies for Reintroduction of Endangered Plants. Washington, DC: Island Press. Hanson, J. 1985. Procedures for Handling Seeds in Genebanks. Rome: International Board for Plant Genetic Resources. Hobbs, R. J., and D. A. Saunders. 1994. Effects of landscape fragmentation in agricultural areas. Pages 77–95 in C. Moritz and J. Kikkawa (eds.), Conservation Biology in Australia and Oceania. Chipping Norton, Australia: Surrey Beatty & Sons. Hopper, S. D. 1998. An Australian perspective on plant conservation biology in practice. Pages 255–278 in P. L. Fiedler and P. M. Kareiva (eds.), Conservation Biology for the Coming Decade. New York: Chapman & Hall. Hopper, S. D., and D. J. Coates. 1990. Conservation of genetic resources in Australia’s flora and fauna. Proceedings of the Ecological Society of Australia 16:567–577. Hopper, S. D., M. S. Harvey, J. A. Chappill, A. R. Main, and B. Y. Main. 1996. The Western Australian biota as Gondwanan heritage: a review. Pages 1–46 in S. D. Hopper, J. A. Chappill, M. S. Harvey, and A. S. George (eds.), Gondwanan Heritage: Past, Present and Future of the Western Australian Biota. Chipping Norton, Australia: Surrey Beatty & Sons. Hopper, S. D., S. van Leeuwin, A. P. Brown, and S. P. Patrick. 1990. Western Australia’s Endangered Flora and Other Plants under Consideration for Declaration. Perth: Australian Heritage Commission and CALM. IUCN (International Union for the Conservation of Nature). 1994. IUCN Red List Categories. Gland, Switzerland: IUCN Species Survival Commission. IUCN (International Union for the Conservation of Nature). 2000. IUCN/SSC Guidelines for Re-Introductions. Accessed January 2000 at http://www.rbgkew.org.uk/ conservation/RSGguidelines.html. McFarlane, D. J., R. J. George, and P. Farrington. 1993. Changes in the hydrologic cycle. Pages 147–186 in R. J. Hobbs and D. A. Saunders (eds.), Reintegrating Fragmented Landscapes. New York: Springer-Verlag. Malajczuk, N., and A. R. Glenn. 1981. Phytophthora cinnamomi: a threat to the heathlands. Pages 241–247 in R. L. Specht (eds.), Ecosystems of the World 9B: Heathlands and Related Shrublands: Analytical Studies. Amsterdam: Elsevier. Mitchell, M., K. Kershaw, E. Holland, G. Stack, and A. Brown. 1997. Shy Feather Flower (Verticordia fimbrilepis subsp. fimbrilepis), Interim Recovery Plan 1997–1999. Perth: Department of Conservation and Land Management.
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Monks, L., and D. Coates. 1999. Restoring diversity. Restoring hope. Landscope 15(1):17–21. Moran, G. F., and S. D. Hopper. 1987. Conservation of the genetic resources of rare and widespread eucalypts in remnant vegetation. Pages 151–162 in D. A. Saunders, G. W. Arnold, A. A. Burbidge, and A. J. M. Hopkins (eds.), Nature Conservation: The Role of Remnants of Native Vegetation. Chipping Norton, Australia: Surrey Beatty & Sons. Morse, J. 1993. Genebanking Australia’s threatened flora. Pages 69–75 in L. D. Meredith (ed.), Cultivating Conservation: Integrated Plant Conservation for Australia. Hobart: Australian Network for Conservation. Pavlik, B. M. 1996. Defining and measuring success. Pages 127–155 in D. A. Falk, C. I. Millar, and M. Olwell (eds.), Restoring Diversity: Strategies for Reintroduction of Endangered Plants. Washington, DC: Island Press. Podger, F. D. 1972. Phytophthora cinnamomi, a cause of lethal disease in indigenous plant communities in Western Australia. Phytopathology 62:972–981. Roberts, E. H. 1991. Genetic conservation in seed banks. Biological Journal of the Linnean Society 43:23–29. Roche, S., K. W. Dixon, and J. S. Pate. 1997a. Seed ageing and smoke: partner cues in the amelioration of seed dormancy in selected Australian native species. Australian Journal of Botany 45:783–815. Roche, S., K. W. 1998. For everything a season: smoke-induced seed germination and seedling recruitment in a Western Australian Banksia woodland. Australian Journal of Ecology 23:111–120. Roche, S., J. M. Koch, and K. W. Dixon. 1997b. Smoke enhanced seed germination for mine rehabilitation in the southwest of Western Australia. Restoration Ecology 5(3):191–203. Rossetto, M., and K. W. Dixon. 1998. Corrigin Grevillea (Grevillea scapigera) Recovery (ESP 495). Major Review 1998. Unpublished report to Environment Australia, Canberra. Rossetto, M., K. W. Dixon, K. Atkins, and D. J. Coates. 1998. Corrigin Grevillea Recovery Plan. Perth: Wildlife Management Program, Kings Park and Botanic Gardens, Department of Conservation and Land Management and Australian Nature Conservation Agency. Rossetto, M., P. K. Weaver, and K. W. Dixon. 1995. Use of RAPD analysis in devising conservation strategies for the rare and endangered Grevillea scapigera (Proteaceae). Molecular Ecology 4:321–329. Sage, L., and B. B. Lamont. 1994. Conservation Biology of the Rare and Endangered Species Lambertia orbifolia. Final report to Science and Information Branch, Department of Conservation and Land Management, Western Australia. Sampson, J. F., S. D. Hopper, and S. H. James. 1988. Genetic diversity and the conservation of Eucalyptus crucis Maiden. Australian Journal of Botany 36:447–460. Sarrazin, F., and R. Barbault. 1996. Reintroduction: challenges and lessons for basic ecology. Trends in Ecology and Evolution 11:474–478. Schatral, A. 1996. Dormancy in seeds of Hibbertia hypericoides (Dilleniaceae). Australian Journal of Botany 44(2):213–222. Schatral, A., J. M. Osborne, and J. E. D. Fox. 1997. Dormancy in seeds of Hibbertia cuneiformis and H. huegelii (Dilleniaceae). Australian Journal of Botany 45(6):1045–1053.
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Shearer, B. L., and M. Dillon. 1996. Susceptibility of plant species in Banksia woodlands on the Swan Coastal plain, Western Australia, to infection by Phytophthora cinnamomi. Australian Journal of Botany 44:433–445. Shearer, B. L., R. Wills, and M. Stukely. 1991. Wildflower killers. Landscope 7(1):29–34. Smith, R. D., and S. Linington. 1997. The management of the Kew Seed Bank for the conservation of arid land and U.K. wild species. Bocconea 7:273–280. Stack, G., R. Evans, and V. English. 1999. Trigwell’s rulingia (Rulingia sp. Trigwell Bridge) Interim Recovery Plan 1999–2002. Perth: Department of Conservation and Land Management. Tieu, A., K. W. Dixon, K. Sivasithamparam, J. A. Plummer, and I. M. Sieler. 1990. Germination of four species of native Western Australian plants using plant-derived smoke. Australian Journal of Botany 47:207–219. Touchell, D. H., M. Richardson, and K. W. Dixon (eds.). 1997. Germplasm Conservation Guidelines for Australia. An Introduction to the Principles and Practices of Seed and Germplasm Banking for Australian Species. Canberra: Australian Network for Plant Conservation. Western Australian Herbarium. 1998. FloraBase: Information on the Western Australian flora, Department of Conservation and Land Management: http://www.calm.wa.gov.au/science/florabase.html. Whittaker, P. K., and B. G. Collins. 1997. Pollen vectors for the rare plant species Lambertia orbifolia. Unpublished report to the Department of CALM, School of Environmental Biology, Curtin University of Technology, Perth. Wieland, G. D. 1995. Guidelines for the Management of Orthodox Seeds. St. Louis, MO: Center for Plant Conservation. Wildlife Conservation Rare Flora Notice. 1999. Government Gazette, Western Australia (December 17):6194–6199. Williams, K., and R. M. Fitzgerald. 1998. Major project review for Rulingia sp. Trigwell Bridge. Endangered Species Project #493. Unpublished Report, Department of Conservation and Land Management, Perth. Wills, R. T. 1992. The ecological impact of Phytophthora cinnamomi in the Stirling Range National Park, Western Australia. Australian Journal of Ecology 17:145–159. Withers, P. C., W. A. Cowling, and R. Wills. 1994. Plant diseases in ecosystems: threats and impacts in south-western Australia (proceedings of the Symposium of the Royal Society of Western Australia and the Ecological Society of Australia.) Journal of the Royal Society of Western Australia 77(4):97–185.
Chapter 4
The Role of Federal Guidance and State and Federal Partnerships in Ex Situ Plant Conservation in the United States Kathryn L. Kennedy
The best place to conserve plant biodiversity is in the wild, where a large number of species present in robust populations can persist in their natural habitats with their associated ecological links (McNaughton 1989). The Convention on Biological Diversity recognizes this as the preferred and most cost-effective way to conserve the maximum amount of biodiversity (Glowka et al. 1994). For single species it is also the most effective way to maintain diverse populations and to regenerate populations through natural reproduction and recruitment under conditions of natural selection (Woodruff 1989). Similarly, maintaining viable populations of species in the wild ensures that these species do not become static specimens but maintain their ecological functions and evolutionary potential. Viable wild populations of species will continue to contribute to species interactions and ecological processes that determine both the local plant community and environmental character (e.g., soils, fauna, hydrology). They remain dynamic, responsive to change in the environment, and maximize their chances of adapting to environmental change. Preserving all levels of the biodiversity hierarchy, namely species diversity, community structure and dynamics, and potential for evolutionary change, is the main challenge facing those who seek to conserve and manage wildlands today (Soulé 1991). For many years conservationists’ and biologists’ primary approach to conservation was governed by the philosophy that habitat should be secured and threats relieved through protection and better management and that accordingly communities and species would recover on their own. When adequate areas of good-quality habitat 67
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are available, with viable populations of constituent species, these areas unquestionably represent a valid focus for conservation investments, and protection and suitable in situ management are likely to be adequate measures for conservation (McNaughton 1989). However, the in situ management of wild populations will become more difficult as habitats are further affected by fragmentation, overuse, invasion by exotic species, changes in ecological processes (e.g., burning cycles or eutrophication), and climate change (Tilman and Lehman 2001).
The Place for Ex Situ Conservation Why are we involved in ex situ work? For some federally listed and other threatened species, in situ approaches probably will not be sufficient (Vrijenhoek 1989). Many listed species have declined to the point that they face a real likelihood of extinction (Bruegmann et al. 2002). If habitat fragmentation has resulted in isolated population segments and extirpation of habitat or populations from intervening areas, restoration of populations and habitat may be necessary to retain species viability (McNaughton 1989; Conway 1989). If populations have become very small, intervention may be needed to prevent an otherwise irreversible decline. The longer a species remains at low population levels, the more likely it is that deleterious population effects such as stochastic catastrophic events (some of which will be natural events), inbreeding depression, or genetic drift will cause extinction (Gilpin and Soulé 1986; Fenster and Dudash 1994). Population-level intervention may be needed, and it is in this arena that ex situ techniques are appropriately brought to bear as a tool to assist in restoring wild diversity (Given 1987). It is clear that restoration ecology, at the level of both species reintroduction and habitat restoration, is a necessary tool for the conservation biologist and natural resource manager (Maunder 1992; Sinclair et al. 1995). In 1984, when the Center for Plant Conservation (CPC) was first founded, the concept of ex situ conservation made some people uneasy, and some conservationists were opposed to its use. There were three major concerns. First was the fear that protecting plant materials off site and removing plant material from the wild would confuse the public, who might perceive ex situ work alone as adequate “conservation.” Conservationists were worried that if plant material was successfully “preserved” in collections and gene banks, this would undermine support for conservation
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of those species’ native habitats and ecosystems (Allen 1994). Second, with conservation resources scarce, there was skepticism about the expenditure of precious funds for expensive ex situ work. It was often regarded as a lower priority than work in the wild, and some feared that seeking funding to establish ex situ collections would take public and private dollars away from needed conservation actions in the wild. Finally, many felt that the premise that plant materials could be restored to the wild was unproven and risky. Even though ex situ work was proposed as a safety net for wild species in decline, and safeguarding plant material ex situ was regarded as a potentially valuable tool for keeping recovery options open by keeping plant material available, there were concerns about proceeding. Some felt that such programs could be more damaging than helpful, given the impacts to wild populations and our lack of understanding of population dynamics and genetics. Others were concerned about the possibility of introducing pests or pathogens into the wild via plant materials used for reintroduction. Fortunately, the response of agencies and the public to the potential benefits of the work has helped allay the first of these concerns, a perception that ex situ work would undermine in situ work. As the science of genetics and demographic analysis of populations has advanced, it has become clearer to professionals that many species in significant decline may be saved only through ex situ methods. Clear guidance, interpretation, and policies from practitioners and conservation agencies help ensure that ex situ work is undertaken not in place of in situ efforts but in concert with them and have been vital to this understanding. Putting ex situ conservation in proper context emphasizes that this work enables restoration action in addition to other in situ conservation action for recovery planning and implementation. This is not just the view of the 32 participating institutions of the CPC. Conway (1989: 208, 209) stated that although gene banks cannot achieve conservation by themselves, “they are essential insurance for the protection of specialized species . . . and crucial in habitat restoration” and that “intensive biotechnology programs are weak and limited tools. But they are as significant to the preservation of biological diversity as the fragment of nature they can save and restore that other kinds of conservation efforts cannot.” It is my view that ex situ work conducted in public institutions such as botanic gardens, museum research programs, and zoos probably has helped make the public and donors more aware of the need for both in
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situ and ex situ conservation action. It certainly has not reduced support for conservation in the wild. The interpretation of institutional work in conservation and the educational value of being able to show and explain the plight of imperiled species to visitors are invaluable (Chapter 5, this volume). In the long term, public understanding and support of conservation of biodiversity are critical components of successful conservation and stewardship. We must continually work to refine and improve this message and make it more available to communities across the nation. The CPC’s experiences in building public support for programs in existing institutions have validated Conway’s (1989) view that zoos and botanic gardens are uniquely positioned to become centers for ex situ conservation and research and to make the link between human communities and natural communities. Conway also addressed the concern for funding competition. He noted that sources available to ex situ or biotechnology programs generally come from separate funding sources interested in exploring these technical tools in addition to habitat-based work. He postulated that institutions doing ex situ work may actually have access to donors, institutional funding, and even local and municipal dedicated funds that otherwise would not be brought to bear for conservation action. As for the demonstrated ability of ex situ work to support reintroduction and restoration as conservation tools, there have been both successes and failures. The technology is still very new, and refinements and long-term data for evaluation are needed (Maunder 1992; Falk et al. 1996). Certainly there are risks associated with understanding population structure, disease, and the impact of intervention strategies on the existing populations and communities. Good science is helping us evaluate the relative risks, particularly advances in population genetics and improvements in genetic analysis, which give us the tools needed to make better decisions and plans. Good policies, protocols, and precautions combined with careful preparation and implementation can help minimize risks. Perhaps the greatest argument in favor of using these techniques is that without using them we are de facto choosing to let threatened species and their habitats decline until lost forever. Ex situ techniques give us the capacity for active restoration. Along with identified risks, it should not be forgotten that ex situ conservation also offers benefits beyond preserving restoration options. Genetically representative ex situ collections provide material for research that minimizes impacts to wild populations, offer potential adaptive management options for in situ work (such as providing uninfected or adaptively resistant material to counter the threat of disease), maintain stock to pro-
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duce material for education and appreciation, and provide materials for nondestructively evaluating species for potential human use. Finally, ex situ conservation can safeguard genetic lineages from total loss to science and society if the species goes extinct in the wild (Maunder et al. 2001). Today, CPC participating institutions provide these benefits through ex situ work on almost 600 species (Kennedy 2002). There is no doubt that the CPC’s ex situ work has already helped stave off extinction for many species, such as Florida ziziphus (Ziziphus celata, Rhamnaceae). The species was known from only a handful of wild populations, which were no longer reproducing well in the wild. Careful ex situ work at Bok Tower Gardens, Florida, rejuvenated reproductive potential and produced vigorous plant material for pilot reintroduction work in its Lake Wales Ridge habitat in 2001. Ex situ work by other CPC institutions has allowed the production of appropriate material for restoration work for species such as Pyne’s ground plum (Astragalus bibullatus, Fabaceae) in cedar glade habitat in the central basin of Tennessee, Pitcher’s thistle (Cirsium pitcheri, Asteraceae) in beach and dune habitat in Illinois (McEachern et al. 1994), Texas trailing-phlox (Phlox nivalis ssp. texensis, Polemoniaceae) in deep sand areas of the southern pineywoods area of East Texas, Ventura marsh milk-vetch (Astragalus pyncnostachyus, Fabaceae) to the sandy coastal areas of southern California, the yellow-flowered golden paintbrush (Castilleja levisecta, Scropulariaceae) to the gravel prairies in Washington State, the western lily (Lilium occidentale, Liliaceae) to the Pacific coastal wetlands in Oregon, and Munroidendron racemosum (Araliaceae) in a midelevation native forest restoration project in Kauai, Hawaii. CPC institutional work alone has supported preliminary restoration work on over 80 species. Much of this work is preliminary; it will take many years to evaluate success, and many may not succeed, at least initially. But it is ex situ work that has made it possible, and the ability to undertake this work undoubtedly will lead to greater understanding and eventually to successful restoration projects that provide greater security in situ for threatened species.
The Urgent Need for Ex Situ Support in the Recovery of Threatened U.S. Plants In 1997 I undertook a study of final recovery plans for federally listed plant species to examine the status and recovery needs of U.S. listed plants. From plans for 256 plant species, some sobering statistics emerged. According
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to recovery plans, fewer than 10 sites remained in the wild for 65 percent of listed species, and 49 percent had fewer than 5 sites remaining. In addition, for 74 percent of the listed species, the majority of known sites contained fewer than 100 individuals. Reintroduction or augmentation of populations to achieve recovery was recommended for 87 percent of the species. General estimates for viable populations of plants range from 103 to 106 for populations subject to environmental stochasticity and natural catastrophes (Shaffer 1987; Menges 1991). For some with life history strategies better able to escape these impacts, fewer reproductive adults may suffice (Pavlik 1996), though rarely less than 100. Consequently, for 74 percent of the species surveyed, it appears basic population viability is at risk. Furthermore, populations of only a few hundred individuals are considered likely to be at risk for deleterious population-level genetic processes such as inbreeding depression (Barrett and Kohn 1991). Populations of 100 or less not only are likely to have reduced viability but also are considered at risk of losing alleles through fluctuations in gene frequencies (Barrett and Kohn 1991) and at heightened risk of extinction from random or cyclical environmental phenomena. The longer these diminished populations go without being recovered, the more likely it is that additional individuals will be lost and that additional genetic erosion will occur. Given this profile of federally listed plant species and their recovery plans, the real need for an active ex situ component for most species is clear. Not only will plant material be needed for restoration, but immediate attention to securing ex situ material is also needed to prevent the continued loss of genetic diversity that will make recovery more difficult and expensive. As of 2003, there are 746 listed plant species and 139 candidates for listing (see http://ecos.fws.gov/tess_public/html/boxscore.html). If these same general conditions hold true for them, then about 5 percent of the flora of the United States probably is in need of some degree of restoration and ex situ work.
Guiding Recovery of Threatened Species: The Endangered Species Act and the U.S. Fish and Wildlife Service The federal Endangered Species Act of 1973 (ESA), as amended, is the most powerful federal legislation for the protection and recovery of threatened species. Regulated species are designated as Endangered, Threat-
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ened, or Proposed. For plants this federal regulatory protection pertains primarily to federal lands and, in a few situations, to private lands as well. The U.S. Fish and Wildlife Service (USFWS) has the administrative and regulatory responsibility for implementing the ESA for most plants. The USFWS is responsible for evaluating the status of species over their range, listing those that qualify, planning comprehensive recovery for species as a whole, and administering permit or consultation requests for necessary and reasonable activities that otherwise might be prohibited. The ESA mandates other USFWS activities related to listed species as well, such as enforcement and international treaties. In addition, the USFWS manages the natural resources in the National Wildlife Refuge System and administers the Partners for Wildlife Program for private landowners. Under the ESA, cooperative consultations are mandated between the USFWS and any federal agency funding, permitting, or implementing actions that may affect listed species on federal or private land. The USFWS is required to produce a document (called a biological opinion) that reviews the potential impacts of any such action (positive or negative) and steps to be taken to minimize damage and maximize recovery. Actual responsibility for avoiding damage and destruction, as well as for any recovery actions undertaken for listed species on federal lands, rests with the federal agency managing those lands. Because of the USFWS’s regulatory and coordination role, it provides leadership nationally in threatened species management practices. USFWS policies and guidance are updated periodically to reflect current conservation science. The USFWS seeks out the best available scientific information and advice in formulating revisions to policies and guidance and in drafting recovery plans, and invites peer review. Although draft policies and plans often generate lively discussion and dissent in the academic community and from public stakeholders subject to regulation, there is generally good compliance in the federal and private sector with final USFWS policies. The CPC signed a memorandum of understanding with the USFWS in June 2000 formalizing their valuable partnership. The memorandum pledges cooperation between the USFWS and the CPC at the national and local level in implementing recovery for listed species through identification of tasks that may be undertaken cooperatively. The CPC has agreed to provide expertise in recovery planning and review. In addition the memorandum provides for cooperation in developing and sharing out-
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reach materials and opportunities and in seeking ways to work cooperatively to implement both ex situ and in situ tasks that will lead to recovery. The mandate of the USFWS as stated in the ESA is to “provide a means whereby the ecosystems upon which endangered species and threatened species depend may be conserved” (ESA as amended, 1988, Sec. 2(b)). Consequently, the policies and guidance for ESA administration rightly take a habitat-based approach. Other mandates for the USFWS are contained in the ESA, the implementing regulations, and USFWS policies and guidance. Understanding USFWS policies regarding ex situ work, restoration, and reintroduction is important.
USFWS Policy Regarding Controlled Propagation One policy ex situ workers must be familiar with is the recently revised USFWS Policy for Controlled Propagation of Species Listed Under the Endangered Species Act (USFWS 2000). Although the policy governs USFWS activities, their partnership in most conservation activities means that the policy applies to USFWS cooperators as well. The 14-point policy sets rigorous standards and goals but also is structured to build on the work of experienced ex situ providers by encouraging the use of the protocols already developed by groups such as the American Zoo and Aquarium Association (AZA) and the CPC for responsible management and restoration. Flexibility is provided through exemptions and avenues for special permission in unique cases. The first requirement states that captive propagation should be used only when other efforts to maintain or improve the species’ status in the wild have failed, are likely to fail, or are shown to be ineffective or insufficient to achieve full recovery. Given the high cost of intensive ex situ methods, in cases where habitat management is underway and appears likely to succeed in restoring target populations, this approach is supportable. If timely intervention is anticipated, however, ex situ work may be appropriate. Cooperation in priority setting is important to effective resource use. The second and third points address the need for captive propagation plans to be integrated with other recovery actions working both in situ and ex situ, oriented toward significant progress in restoration to the wild and consistent with recovery plans. This is obviously essential to good teamwork and ultimate success. Such integration will entail cooperation and
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communication among all restoration partners, and ex situ providers have an ethical responsibility to facilitate this interaction. Another group of requirements in the propagation policy relates to planning for captive propagation through recovery plans, genetic management plans for ex situ collections, recordkeeping, and written reintroduction plans. Written plans are recommended by the CPC as well. I recommend that practitioners have plans peer reviewed by experienced ex situ practitioners. The requirements for written plans drew negative comments from botanists in academia and gardens who view it as needlessly bureaucratic, but I disagree with these objections. Planning is the best way to maximize the chances of success, formalize risk assessments for donor and receiving populations and communities, specify protocols and goals, and avoid potential errors resulting from impulsive decisions. Written plans take additional time, and we all want to fight “paper paralysis,” but without them it is difficult to document best conservation practices and demonstrate success. Where initial species information is scarce these plans may be brief, but the operating assumptions and decisions still must be stated. Good plans lay valuable groundwork for those following in your footsteps. If these plans are not written, misunderstandings are more likely, and the heuristic value of examining modifications and evaluating success is lost. We advocate such plans in the CPC, and although compliance is not perfect, we recognize its importance and strive to meet this goal. In coordinating captive propagation to provide high-quality programs, the USFWS has increased its own requirements for planning, coordination, annual reporting, and compliance reviews. Overworked biologists in bureaucracies are very reluctant to do this needlessly. If they are willing to subject themselves to this workload to improve quality, tracking, and scientific understanding, we can do no less. Other points address good conservation practices for captive propagation to secure adequate funding, maintain species integrity and genetic variation, avoid disease and pest introduction, avoid escape from the known range of the species, disperse (and where possible duplicate) collections at more than one location for better security, and comply with existing laws and permit requirements. Intercrossing is not permitted unless specifically noted as needed in the recovery plan to compensate for loss of genetic viability, and even in these cases intercrossing between populations must be approved by the director of the USFWS (Chapter 16, this volume).
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The USFWS goal was to have controlled propagation programs in compliance with this policy within 1 year of the policy’s promulgation, which would have been in September 2001. The USFWS may continue involvement in programs not in conformity only at the recommendation of regional directors and the approval of the director.
USFWS Recovery Implementation Guidance The USFWS Policy and Guidelines for Planning and Coordinating Recovery of Endangered and Threatened Species (USFWS 1990) is under revision. Section IV of the current report, “Service Policies Relating to Recovery,” includes guidance for captive propagation or cultivation that has been superseded by the new policy. Additional guidance in Section IV of interest to ex situ workers addresses relocation of listed species and notes that listed species may not be relocated or transplanted outside their historic range without specific case-bycase approval of the director. The section also discusses the USFWS working definition for reintroductions and introductions at that time, and includes a discussion of experimental populations that may be established outside the current range of the species but within the historic range. It is essential for ex situ practitioners to be aware of the revision process under way for recovery guidance and to participate in the planning and review processes of this and other guidance and future revisions. This is necessary to ensure that new guidance is effective, is based on the best available scientific information, and provides best conservation practices for the foreseeable future.
Government Agencies as Leaders and Partners in Conservation Federal agencies work within the mandates of the ESA and other natural resource laws that require their compliance, give them administrative responsibility, or mandate program objectives. These other resource laws also provide support for recovery of imperiled plants. Examples of other federal legislation include the U.S. National Forest Management Act of 1976 (currently under revision), which requires that the U.S. Forest Service (USFS) maintain viable populations for all the species in their care; the Defense Appropriations Act of 1991 (instituting the Legacy Resource Man-
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agement Program) and the Sikes Act of 1960, guiding management of natural resources on Department of Defense lands; the National Park Service Organic Act of 1916, governing conservation of National Park Service resources; the Federal Land Management Act, governing Bureau of Land Management practices; the Food, Agriculture, Conservation, and Trade Act of 1990, authorizing the U.S. Department of Agriculture (USDA) to establish the National Genetic Resources Program; and other more familiar acts such as the National Fish and Wildlife Coordination Act, the National Environmental Policy Act, and the Clean Water Act. Many state agencies also have specific regulatory legislation. The shared responsibility between the USFWS and other federal agencies for populations on federal lands (or affected by federal actions) and the need for consultation in most cases lead to a strong partnership between the USFWS and other federal agencies in conserving and managing listed species. In 1994, 10 federal agencies signed a memorandum of understanding (known as the Native Plant Conservation Initiative) pledging interagency cooperation. This partnership has strengthened interagency coordination and cooperation in native plant conservation. Working with federal and state agency partners is integral to the ultimate success of any ex situ conservation effort (McMahan 1995). In the United States, federal lands held by the Department of Interior’s Bureau of Land Management, the USFWS, and the National Park Service, along with the lands of the USFS and the Department of Defense, make up 30 percent of the land area of the United States. These lands and those of other federal agencies contain a significant percentage of known listed species sites and potential habitat for others. It is these lands that have the highest level of protection for federally listed plants. Federal agencies are prohibited from intentional damage and destruction of listed plant species on their lands, are responsible for protection and management, and generally embrace the security and restoration options that ex situ programs offer. With today’s federal and state budget priorities and lack of recognition of the urgency for plant conservation in public policy, agencies are seldom in a position to completely fund ex situ work, but they can often find some funding assistance for priority actions. In addition, federal and state agencies often have programs to assist private landowners in habitat restoration and endangered species recovery, such as the USFWS Partners for Wildlife Program and the USDA-funded programs in farm bill legislation. Congress also provides funding to the National Fish and Wildlife Foundation
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and the National Park Foundation. These foundations raise additional funds and provide matching grants to private partners for projects that further agency objectives on federal or private land. Some states have similar park foundations. Agency resource management staff and scientists outside traditional landholding agencies, such as those in the USDA and the U.S. Geological Services Biological Resources Division, are valuable resources. They have extensive species knowledge and ecological, genetic, and modeling expertise and assistance to offer ex situ workers. For example, the USDA has significant expertise in ex situ work with plants and seeds and shares a concern for biodiversity conservation to support sustainable agriculture. Their programs include the USDA National Genetic Resources Program and its associated National Plant Germplasm System with the premier National Center for Genetic Resources Preservation (formerly known as the National Seed Storage Laboratory). The CPC has a long-standing cooperative agreement with the USDA National Center for Genetic Resources Conservation, benefiting the mission of both parties. Similarly, the USDA’s Natural Resource Conservation Service (formerly the Soil Conservation Service) has Plant Materials Centers, where much horticultural and genetic expertise resides. Although they may not specialize in native species, they may be very helpful in troubleshooting and partnering for analysis and storage. As we acknowledge the value of federal and state expertise and guidance, we must also observe that agencies generally have too few botanists in their field offices to address the workload. Most agency offices need and welcome the involvement of other botanists as partners for planning and implementation. With so many plant species in need of urgent attention, neither the public nor the private sector can handle the job alone. Federal–private partnerships are critical to achieving recovery. The CPC’s approach to plant conservation work is organized around the concept that local institutions can provide professional assistance in recovery and stewardship to state and federal agencies and that doing so will help move urgent recovery work forward. Federal agencies are reaching out for partners, and their interagency Native Plant Conservation Initiative memorandum of understanding was expanded to allow private partners to sign on. This effort has evolved into the Plant Conservation Alliance, with 15 federal agencies and more than 200 organizations pledged to work cooperatively for native plant conservation.
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Improving State and Federal Conservation Partnerships Federal and state partnerships are crucial to success. State agencies are crucial where they are actively involved in plant conservation or management of state or federally listed species. Lands managed by state and federal agencies are critical to recovery for many threatened plants. Federal agencies have legislative mandates for coordination, regulation, and permitting of recovery action for listed species on federal and in some cases private lands. Some state agencies also have a mandated or regulatory role. Federal agencies generally embrace ex situ work as a valuable tool for recovery and offer important guidance and support resources. Clearly plant conservationists must support and nurture federal–private partnerships. There are a number of simple but important steps that partners can follow to optimize benefits and progress for plant conservation. All of these steps relate to good communication at as many levels as possible and support for partnerships and agency efforts (Clark and Cragun 1994). • Find state and federal partners who have expertise and experience
that may be helpful. Get to know their resource people. As discussed earlier, expertise does not reside only in the land management agencies. Explore the work of other federal agencies and cooperative federal research programs with universities. • Share species information regularly, including problems, achievements, and needs. Be generous with your data, literature review time, and copy machine funds. Who would benefit from the information you have or have seen? Provide copies of publications, unpublished reports, or summaries of interest from your own work or those of others. If you are a research scientist, publish your results in a timely and effective manner and make detailed data available as soon as possible. Archive copies of your work and notes with state and federal resource managers when you finish a project or retire. • Cooperatively assess information gaps and needs for good ex situ management. Summarize these needs, and be sure the entire conservation community is aware of them. • Know the federal and state agencies responsible for regulatory or management responsibilities for imperiled plants in your area. Establish a face-to-face relationship with those responsible for management or recovery of species you work with. Schedule meetings for progress reports and information exchange. Discuss
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•
•
•
•
•
•
issues of conservation theory and practice and how they relate to species of mutual concern. Become familiar with statutes, policies, and guidance regarding plant resources for all regulatory and management agencies in your area. Understand the mandates, powers, programs, and constraints of federal and state agencies and respect the framework in which agency biologists operate. Doing so helps avoid noncompliance, frustration, duplication, and wasted time and resources. It also will illuminate how agencies make decisions and coordinate with their partners. Know, abide by, and assist agencies in their compliance with statutory requirements and policies. Get required permits, submit reports on time, and provide information needed for agencies to fulfill their reporting requirements. Coordinate with state and federal partners in setting priorities for ex situ conservation. Review the status of species you believe are of high concern and know where and why your priorities are in agreement and where they are not. Participate in planning, review, and comment processes with state and federal agencies for species plans and ex situ management plans. Serve on recovery teams, attend important public meetings, and provide official comments to agencies so that they have the benefit of your voice and recommendations in planning and implementation. Ask to be added to their mailing lists for information and reports. Know the timing and process for developing agency budget requests and for allocating and spending current year funds. Ask that local and regional budget requests address plant recovery needs, including ex situ work where necessary. Provide information that may be helpful for setting priorities and writing proposals for integrated initiatives and general justifications. Working together, convert conservation planning into specific project proposal form. Nothing makes assessments and planning documents more effective than putting down in written form exactly who can or should do it, when they can do it, how long it will take, and what it will cost. Stating conservation needs in this form makes it easy to incorporate them in budget requests, funding proposals, grant proposals, and legislative packages.
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• Support agency budget requests for resources for plant conservation.
Use your voice to support local, regional, and national budget requests both within and outside the agency. • Help find supplemental funding to implement ex situ and in situ conservation projects. It is not appropriate to regard conservation implementation as the sole responsibility of state and federal agencies. Local communities and private partners must be willing to participate in the process and funding of resource conservation. Prepare grant proposals and apply for funding from nongovernment sources that may be able to provide or supplement funding for critical action. Ask agency partners to help make the case for needed funding.
Acknowledgments I drew on many resources and experiences in writing this chapter. I am grateful for my experiences with the Texas Parks and Wildlife Department and the USFWS. My association with the CPC dates back to my advisory days in the mid-1990s. These organizations have taught me a lot and were invaluable partners for me during my years in the field and as an agency bureaucrat. The scientific advisors to CPC and the hard-working professionals in the CPC institutions (directors, horticulturists, population geneticists, taxonomists, ecologists, and many other students of plants in situ and ex situ) have done much of the work referred to here. I am particularly indebted to Kay Havens, Ed Guerrant, Mike Maunder, Peggy Olwell, Carol Spurrier, Chris Walters, Loyal Mehrhoff, and the regional botanists of the USFS for examples, reference materials, thoughtful discussions, and ideas about partnerships and opportunities to achieve more for our imperiled flora. References Allen, W. H. 1994. Reintroduction of endangered plants. BioScience (44)2:65–68. Barrett, S. C. H., and J. R. Kohn. 1991. Genetic and evolutionary consequences of small population size in plants. Pages 3–30 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Bruegmann, M. M., V. Caraway, and M. Maunder. 2002. A safety net for Hawaii’s rarest plants. Endangered Species Bulletin 27(3):8–11. Clark, T. W., and J. R. Cragun. 1994. Organizational and managerial guidelines for endangered species restoration programs and recovery teams. Pages 9–33 in M. L.
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Bowles and C. J. Whelan (eds.), Restoration of Endangered Species. Cambridge, UK: Cambridge University Press. Conway, W. G. 1989. The prospects for sustaining species and their evolution. Pages 199–209 in D. Western and M. Pearl (eds.), Conservation for the Twenty-first Century. New York: Oxford University Press. Falk, D. A., C. I. Millar, and M. Olwell (eds.). 1996. Restoring Diversity: Strategies for Reintroduction of Endangered Plants. Washington, DC: Island Press. Fenster, C. B., and M. R. Dudash. 1994. Genetic considerations for plant population restoration and conservation. Pages 34–62 in M. L. Bowles and C. J. Whelan (eds.), Restoration of Endangered Species. Cambridge, UK: Cambridge University Press. Gilpin, M. E., and M. E. Soulé. 1986. Minimum viable populations: processes of species extinction. Pages 19–34 in M. E. Soulé (ed.), Conservation Biology: The Science of Scarcity and Diversity. Sunderland, MA: Sinauer Associates. Given, D. R. 1987. What the conservationist requires of ex situ collections. Pages 103–117 in D. Bramwell, O. Hamann, V. H. Heywood, and H. Synge (eds.), Botanic Gardens and the World Conservation Strategy. London: Academic Press. Glowka, L., F. Burhenne-Guilman, H. Synge, J. A. McNeely, and L. Gündling. 1994. A Guide to the Convention on Biological Diversity. Environment Policy and Law paper no. 30. Gland, Switzerland: IUCN. Kennedy, K. 2002. The Center for Plant Conservation. Endangered Species Bulletin 27(3):5–7. Maunder, M. 1992. Plant reintroduction: an overview. Biodiversity and Conservation 1:21–62. Maunder, M., R. S. Cowan, P. Stranc, and M. F. Fay. 2001. The genetic status and conservation management of two cultivated bulb species extinct in the wild: Tecophilaea cyanocrocus (Chile) and Tulipa sprengeri (Turkey). Conservation Genetics 2:193–201. McEachern, A. K., M. L. Bowles, and N. B. Pavlovic. 1994. A metapopulation approach to Pitcher’s thistle (Cirsium pitcheri) recovery in southern Lake Michigan dunes. Pages 194–218 in M. L. Bowles and C. J. Whelan (eds.), Restoration of Endangered Species. Cambridge, UK: Cambridge University Press. McMahan, L. R. 1995. Working with the Feds. The Public Garden 10(2):16–19. McNaughton, S. J. 1989. Ecosystems and conservation in the twenty-first century. Pages 109–120 in D. Western and M. Pearl (eds.), Conservation for the Twenty-first Century. New York: Oxford University Press. Menges, E. S. 1991. The application of minimum viable population theory to plants. Pages 45–61 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Pavlik, B. M. 1996. Defining and measuring success. Pages 127–155 in D. A. Falk, C. I. Millar, and M. Olwell (eds.), Restoring Diversity: Strategies for Reintroduction of Endangered Plants. Washington, DC: Island Press. Shaffer, M. L. 1987. Minimum viable populations: coping with uncertainty. Pages 69–86 in M. E. Soulé (ed.), Viable Populations for Conservation. Cambridge, UK: Cambridge University Press. Sinclair, A. R. E., D. S. Hik, O. J. Schmitz, G. G. E. Scudder, D. H. Turpin, and N. C. Larter. 1995. Biodiversity and the need for habitat renewal. Ecological Applications 5(3):579–587. Soulé, M. E. 1991. Conservation: tactics for a constant crisis. Science 253:744–750.
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Tilman, D., and C. L. Lehman. 2001. Human caused environmental change: impacts on plant diversity and evolution. Proceedings of the National Academy of Sciences of the United States of America 98(10):5433–5440. USFWS (U.S. Fish and Wildlife Service). 1990. Policy and Guidelines for Planning and Coordinating Recovery of Endangered and Threatened Species. Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. USFWS (U.S. Fish and Wildlife Service). 2000. Policy regarding controlled propagation of species listed under the Endangered Species Act. Federal Register 183(65):56916–56922. Vrijenhoek, R. C. 1989. Population genetics and conservation. Pages 89–98 in D. Western and M. Pearl (eds.), Conservation for the Twenty-first Century. New York: Oxford University Press. Woodruff, D. S. 1989. The problems of conserving genes and species. Pages 76–88 in D. Western and M. Pearl (eds.), Conservation for the Twenty-first Century. New York: Oxford University Press.
Chapter 5
Ex Situ Support to the Conservation of Wild Populations and Habitats: Lessons from Zoos and Opportunities for Botanic Gardens Mark R. Stanley Price, Mike Maunder, and Pritpal S. Soorae
The motives for keeping collections of wild animals and plants have evolved over the centuries in response to changing scientific, social, and political environments. This chapter looks at the evolution of zoo conservation activities and assesses the contribution of zoos to the conservation of tropical habitats and ecosystems. We draw lessons from the evolution of the modern zoo that have a direct relevance to ex situ plant conservation activities. The last 50 years have seen three remarkable changes in the world’s zoos. The first is a general improvement in the conditions under which animals are both kept and displayed, the second is that zoos and similar institutions have actively adopted the agenda of ex situ conservation, and the third is the adoption of in situ conservation as a responsibility. The role of the zoo as ark has been further modified with increasing emphasis on the zoo as a multidisciplinary facility supporting broad environmental conservation and education issues, as outlined in the concept of the biopark (sensu Robinson 1992). Although the value of ex situ conservation management has been broadly accepted, it has been subject to debate regarding its strategic application (Ebenhard 1995; Snyder et al. 1996). We derive lessons from this evolution and apply them to plant conservation, and botanic gardens in particular. Although zoos have a long historical tradition of collecting and studying animal diversity, their conservation role has been explicitly described as a broad goal only since the mid-twentieth century. However, the need to bring 84
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animals into captive management for conservation was recognized by a number of early zoo pioneers. For instance, as early as the 1880s, the planned U.S. National Zoo, Washington, D.C., was proposed as a “home and a city of refuge for the vanishing races of the continent” (Lefkowitz-Horowitz 1996: 128). Late-nineteenth- and early-twentieth-century conservation initiatives were responsible for retaining a small number of large mammal species that would otherwise have become extinct, most notably the American bison (Bison bison) and Père David’s deer (Elaphurus davidianus). In the twentieth century, several conflicting trends have influenced the management of captive stocks of wild animals. Human society has systematically studied the earth and gained an increasingly sophisticated understanding of its biological diversity and ecosystem processes (a learning process that is clearly not finished yet); conversely, the impacts of human society on wild habitats and wild species have become increasingly dire. At the same time, the animal holdings of ex situ collections have increased in terms of both the numbers of individuals held and the diversity of species, so that zoos are holding increasingly rare and valuable animals. Consequently, they have turned their attentions to contributing to the conservation of these same species rather than using them only for public recreation, profit, or academic study. The concerns and roles of ex situ institutions, with respect to conservation of animal species, have expanded dramatically over the last few decades. Since the 1970s zoos have promoted the ark paradigm, the idea that ex situ facilities would hold stocks of threatened species during a period of imminent and intense pressure on wild populations, the “demographic winter” sensu Soulé et al. (1986). This role was derived in part from an ethical decision to reduce collecting from the wild and to manage captive stocks self-sufficiently from wild founders. The next stage, the management of species for conservation and potential reintroduction, led to collaboration with in-country conservation projects and a realization that zoos could play a role in combating the global biodiversity crisis (Robinson 1992; Rabb 1994; Conway 1995, 1996). In the late twentieth century there have been some spectacularly successful examples in which ex situ conservation has made a major contribution, or in some cases been the critical factor, in the effective preservation of animal species and their subsequent reintroduction to the wild; examples include the Arabian oryx (Oryx leucoryx), black-footed ferret (Mustela nigripes), California condor (Gymnogyps californianus), and golden lion tamarin (Leontopithecus rosalia rosalia).
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Although out-of-habitat (ex situ) conservation is an accepted tool for conservation, most ex situ investment takes place out of the range country of the target species, and it is expensive in terms of both capital and personnel (Balmford et al. 1995). The emphasis by zoos on conservation biology is in dramatic contrast to that of botanic gardens, where until recently taxonomy was the dominant “mother science.” Although both botanic gardens and zoos had early links with colonial development and agriculture, the institutional background for the majority of botanic garden research has been systematic and economic botany, with a strong commitment to maintaining diverse collections serving the needs of researchers and public horticultural display. In contrast, the strategic programs of many zoos have developed in direct response to the expectations of at least three important, often vocal stakeholders: • A critical and demanding public (as an essential source of income
and political support) • National and international conservation agencies • Pressure from fellow institutions and colleagues to participate in regional collaborative breeding programs for threatened species
The fundamentally different natures of plant and animal collections have also influenced institutional ethics and practices. For example, botanic gardens have not been forced to adopt the genetic management of display species because of the flexibility of vegetative propagation and the ease of propagule distribution between collections. The botanic garden community has focused on regional and local roles, with no established infrastructure for the international management of threatened species. However, the development of dedicated conservation research facilities at a number of botanic gardens (e.g., Chicago Botanic Garden; Millennium Seed Bank at Royal Botanic Gardens, Kew; and Kings Park and Botanic Garden, Australia) shows the developing commitment to both conservation biology and international conservation responsibilities. The modern zoo is no longer isolated from the business and public domain. Indeed, zoos as successful public attractions are becoming increasingly commercial in their approach to advertising and retail sales (Davis 1997). It is notable that corporate businesses run spectacular animal displays with high levels of capital investment, such as the Disney Animal Kingdom, Florida (Hancocks 2001). The need for any zoo to be financially and politically viable has led to the establishment of very effective promo-
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tional mechanisms, with projects and displays that match public expectation, funded by a mixture of grant and donor support, commercial activities, and visitor revenues. In addition to providing satisfying displays and services, the public now expects zoos to conduct research and support conservation activities. A survey by the American Zoo and Aquarium Association (AZA 1999) illustrates the impact of zoos in these areas. In 1997 alone, AZA-listed zoos and aquaria in the United States supported nearly 700 field conservation and research projects in 80 countries. The same group invests $51 million in scientific research each year, and between 1990 and 1999, zoo and aquarium scientists and their university collaborators have published more than 4,000 articles in scientific journals, books, and conference proceedings. Since 1960, the experiences of the zoo community have been disseminated through the forum of the International Zoo Year Book, which contains peer-reviewed scientific papers on husbandry, conservation, and zoo development issues. Furthermore, a singular new discipline, zoo biology, the science of managing ex situ animal populations, has its own dedicated journal, Zoo Biology. But it is notable that the botanic garden community has no international refereed publication dedicated to the scientific management of threatened plant stocks. This scientific investment by zoos has established a set of scientific resources and tools that are being widely used beyond the zoo community (Ryder and Feistner 1995), for instance, through • Conservation genetics (Templeton and Read 1984; Wayne et al. 1986) • Small population and metapopulation management (Lacy 1987;
Foose and Ballou 1988) • Veterinary and assisted reproduction techniques (Spencer 1993) • Conservation and collection planning (Hutchins et al. 1995;
Balmford et al. 1996) • Information management and planning (Ellis and Seal 1995; Westley and Vredenburg 1997) • Environmental and social education (Esson and Tomlinson 1997) • Biodiversity display and interpretation (Coe 1985; Reade and Waran 1996)
The International Agenda for Botanic Gardens in Conservation (BGCI 2001) promotes a valid integrated and scientific approach to plant conservation, but the world network of botanic gardens has yet to establish such
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a systematic, collaborative, in-depth contribution to plant conservation as has been achieved by the zoos. The broader question we examine in this chapter is, Where and how can ex situ investment make the most difference to in situ conservation, and to what extent do ex situ institutions achieve this? We explore this question by examining the role of zoos and looking at the potential conservation impact of botanic garden plant collections and displays.
Definitions and Usage The terms in situ and ex situ are used widely in conservation literature, often with confusion over their precise meanings. Given the ambiguities in the mosaic of present approaches, we propose some redefinitions and distinctions that, if generally adopted, would improve clarity for conservation: • “In situ” refers to an organism or population living within its natural
range or habitat in its own range country; alternatively, this can be described as “in-country and in-range.” • “Ex situ but in-country” refers to the situation in which the organism or population remains in its country of origin but is located out of its natural range; it might be in a city zoo in its own country. • “Out-of-country” means that the organism or population is anywhere outside its native country (and hence is also outside its range); this is the classic zoo ex situ situation, but “out-of-country” is proposed as the most apt descriptor.
Zoo Policy for Conservation The conservation role of zoos has been explicitly referred to in a number of important strategy documents. For instance, Caring for the Earth (IUCN/UNEP/WWF 1991: 40) states, “Zoological gardens have a key role in maintaining ex situ populations of animals” and calls for combining in situ and ex situ techniques. Similarly, the Global Biodiversity Strategy (WRI/IUCN/UNEP 1992) calls for conservationists to “strengthen the conservation role of zoological parks” and “strengthen the collaboration among off-site and on-site conservation institutions, partly to enlarge the role of off-site facilities in species reintroduction, habitat restoration and rehabilitation.” The World Zoo Conservation Strategy (WZCS) explicitly links zoo management with habitat conservation issues (IUDZG/CBSG/IUCN/
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SSC 1993: vii): “It cannot be stressed enough that, where there is still hope, the conservation potential of the zoo community will be aimed primarily at supporting the conservation of natural habitats and ecosystems. Where such conservation is no longer possible the Strategy underlines the importance of within zoo species conservation until such times as suitable habitats can be restored or created and maintained.” The WZCS states that all zoos should support the objectives of international conservation policy documents by • Actively supporting, through co-coordinated programs, the
conservation of populations of endangered species in situ and ex situ and, through these, the conservation of natural habitats, biotopes, and ecosystems • Offering support and facilities to increase scientific knowledge that will benefit conservation and lending support to the conservation community by making available relevant knowledge and experience • Promoting an increase of public and political awareness of the necessity for conservation, natural resource sustainability, and the creation of a new equilibrium between people and nature
The modern conservation role of the zoo has evolved through strong collaboration with the International Union for the Conservation of Nature (IUCN; Holdgate 1999) and the specialist groups of its Species Survival Commission (Rabb and Sullivan 1995), particularly the Conservation Breeding Specialist Group (Westley and Vredenburg 1997). This has resulted in major changes in the strategic vision for many zoos (Durrell and Mallinson 1998). The Convention on Biological Diversity (CBD) recognizes the value of ex situ conservation (Article 9) and places an emphasis on undertaking these activities “preferably in the country of origin” and as a support to the “recovery and rehabilitation of threatened species . . . for their reintroduction into their natural habitats” (Glowka et al. 1994: 6). In addition, the CBD calls on nation states to regulate and control the collection of material for ex situ conservation “so as not to threaten ecosystems and in situ populations of species, except where special temporary ex situ measures are required” (Glowka et al. 1994: 7). The CBD assigns use and control over biodiversity to individual range countries. Indeed, Article 15 (“Access to Genetic Resources”) originated from a widely held perception that northern countries were deriving benefit from the genetic resources of the southern nations without appropriate
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box 5.1 Article 9 of the Convention on Biological Diversity: Ex Situ Conservation Under Article 9 each Contracting Party shall, as far as possible and as appropriate, and predominantly for the purpose of complementing in situ measures: (a) adopt measures for the ex situ conservation of components of biological diversity, preferably in the country of origin of such components; (b) establish and maintain facilities for ex situ conservation of and research of plants, animals and micro-organisms, preferably in the country of origin of genetic resources; (c) adopt measures for the recovery and rehabilitation of threatened species and for their reintroduction into their natural habitats under appropriate conditions; (d) regulate and manage collection of biological resources from natural habitats for ex situ conservation purposes so as not to threaten ecosystems and in situ populations of species, except where special temporary ex situ measures are required under subparagraph (c) above; and (e) cooperate in providing financial and other support for ex situ conservation outlined in subparagraphs (a) to (d) above and in the establishment and maintenance of ex situ conservation facilities in developing countries.
compensation (Mugabe et al. 1996). Because this politically charged view applies to all components of biodiversity, an increasing number of countries are developing regulations to control access to genetic resources. These trends are having a profound impact on the relationship between northern biodiversity facilities (e.g., research and display facilities including zoos, botanic gardens, and natural history museums) and the source countries (Grajal 1999). The CBD stipulates that nation states have the sovereign right to exploit their own biological resources and the authority to subject access to national legislation. Box 5.1 summarizes the main points in the CBD text that relate to ex situ conservation. The CBD has been adopted by Botanic Gardens Conservation International (BGCI) as the overarching policy document for guiding botanic garden conservation activities (BGCI 2001); in this respect botanic gardens have adopted a different emphasis than zoos, with a strong focus on international policy adherence.
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What and Where Are the Main Facilities for Ex Situ Conservation? The main facilities that contribute to the ex situ conservation of animals are zoos and aquaria. They are estimated to total about 2,107 facilities in 125 countries (World Resources Institute 1998). There are many other facilities that differ in their size and scope (often focusing on selected taxa) and in what they are aiming to do, usually with some contribution to the conservation of the animals they hold. The latter diverse group includes orphanages, rescue and rehabilitation centers, sanctuaries, wildlife parks and centers, crocodile ranches, and butterfly farms. These numerous facilities usually are privately owned, and their conservation efforts are more often directed toward the care, welfare, and possible release back into the wild (usually locally and very rarely into other countries) of individual animals. A number of such specialized bodies are making significant scientific contributions to reintroduction projects, including the International Crane Foundation and the Peregrine Fund. In general, however, the international body of zoos, which tend to be the larger and more complex institutions, represents the biggest network of conservation-focused institutions for which the in situ and ex situ conservation of populations and species are very significant objectives. These facilities are not evenly distributed and are concentrated into a small number of zoo-rich nations. For instance, six nations (the United States, Germany, Japan, France, China, and the United Kingdom) collectively hold 1,202 zoos, representing 57 percent of the world total. These six nations include only two with biodiversity hotspots (Myers et al. 2000): the United States and China. On a regional scale Europe has 699 zoos, Africa 66, North America 522, Central America 120, South America 130, Asia 497, and Oceania 73. Within these regions, ex situ facilities are not evenly distributed; for instance, 22 (33 percent) of Africa’s 66 zoos are in the Republic of South Africa. The association is not complete, but the level of per capita gross national product is one determinant of the level of investment in zoos: relative affluence encourages both government and public investment in zoos. Botanic gardens show a similar pattern of distribution at a regional and global level. There are 98 botanic gardens in sub-Saharan Africa, but 19 (22 percent) of these are in the Republic of South Africa. Ten countries rich in botanic gardens (the United States, Germany, China, India, France, Australia, the Russian Federation, the
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United Kingdom, Japan, and Italy) collectively hold 1,202 botanic gardens (Wyse Jackson 2001), accounting for about 50 percent of the world total. In conclusion, although the global portfolio of ex situ facilities represents an enormous body of capital and human resources, their distribution is inevitably skewed toward rich nations. The greatest concentrations of wild species are in the tropical developing countries, yet these are the areas with the fewest financial resources for ex situ conservation. Thus, where animal species need conservation management out of their natural habitat or range, this is more likely to be in an institution outside their own country. These conclusions raise two questions: are these out-of-country facilities keeping the species of highest priority for conservation, and what activities and resources do these out-of-country facilities offer conservation?
Are Out-of-Country Zoo Facilities Managing the Species That Are Most in Need? The world is estimated to hold a total of 27,300 species of birds, mammals, reptiles, and amphibians (Mittermeier and Konstant 1999). This statistic alone confirms the fact that zoos cannot house more than a fraction of all animal species (Sheppard 1995; Balmford et al. 1996). On the other hand, there may be no need to maintain captive populations of a significant number of species, and zoo strategies may be better focused on selected species (and habitats) rather than on maintaining nonviable numbers of more species. The IUCN Reintroduction Specialist Group (RSG) recently surveyed the representation in zoos of African vertebrate taxa (excluding fish species). A list of African threatened and endangered species was compiled from IUCN sources (Groombridge 1993). The holdings of African vertebrate species were summed across 482 facilities in the International Species Information System (ISIS) and the European Association of Zoos and Aquaria (EAZA; Table 5.1). Of Africa’s endangered or threatened amphibian species, there are none in ISIS or EAZA collections. In contrast, 14 out of 17 (82 percent) of the listed reptile species are held in captivity. Threatened and endangered bird species are poorly represented. A higher percentage of mammal species are secure in these institutions: the proportion varies from a low of 30 percent for insectivore and bat species up to 64 percent for herbivores, which are mostly antelopes; primates and carnivores rate 46 percent and 50 percent representation, respectively.
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table 5.1 The total numbers of endangered or threatened taxa of African vertebrates (excluding fish) and the numbers of each in zoos in the International Species Information System and European Association of Zoos and Aquaria. Taxon
Insectivores and bats Primates Carnivores Herbivores Birds Reptiles Amphibians
Total Number of Taxa
Number of Taxa in Zoos
10 34 6 36 29 17 7
3 12 3 23 5 14 0
A subset of the taxa in zoo collections is managed for conservation, with the stated management objectives including reintroduction where appropriate. The Species Survival Plans of the AZA cover 136 taxa, with plans for 77 mammals, 21 birds, 8 reptiles, 2 amphibians, 4 invertebrates, and 24 fish. The list is dominated by mammals (57 percent), and 27 of the 77 are large African species. Large animals undoubtedly are more attractive to the public (Ward et al. 1998), and the entertainment value and drawing power of each species and its individuals were undoubtedly criteria for their acquisition by zoos in the past. But zoos are increasingly establishing expertise and the means to display smaller animals, such as amphibians (Preece 1998), fish (Fiumera et al. 1999), and invertebrates (Mace et al. 1998), attractively to the public. Because many of the displayed taxa are threatened, total conservation responsibilities and potentials are increasing. Zoos and related facilities will be able to house only a small proportion of the world’s threatened animal species. The taxonomic representation and conservation management of threatened species in zoos are dominated by large, charismatic mammal taxa. Resources for conservation management of threatened taxa in North American and Australasian zoos are still focused on species from other regions. The African species currently held in captivity, as recorded in ISIS and EAZA facilities, do not reflect those species’ conservation status in the wild, which might form a means for assessing priority for conservation action through captive breeding. This pattern is also observed in botanic gardens. Two recent studies of representation in botanic gardens revealed that holdings of threatened plants do not fully reflect conservation priorities (Maunder et al. 2001a, 2001b).
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The Conservation Impacts of Out-of-Country Ex Situ Facilities The development of the zoo as a scientific and public display facility has led to significant gains for global conservation. For instance, early attention by zoo managers to stock pedigrees and breeding regimes to make most effective use of captive animal resources has led to great insights into the management of small populations in the wild and the need for a metapopulation approach (Flesness 1977; Chesser et al. 1980). Current zoo populations serve a wide variety of conservation functions: • The maintenance of demographically and genetically adequate • •
• •
• •
populations that are managed without further imports from the wild The co-coordinated management of both wild and captive individuals as a component of a managed species recovery program The retention and maintenance of a high or adequate proportion of the total genetic diversity within a threatened species as an integral component of a managed species recovery program Public education and awareness of local, national, and international conservation issues Establishment of research facilities to promote the health, welfare, and breeding of the species in captivity and to assist conservation efforts in situ for the same or related species Public awareness and recreation A resource and venue for fundraising for conservation activities
Many of these objectives contribute to the oft-stated goal of ex situ facilities to perform the role of Noah’s ark (Soulé et al. 1986).
Out of the Ark: Reintroductions and Repatriations In 1998 the IUCN RSG published a directory of all people willing to be listed as engaged in reintroduction, with details of the species with which they were associated (Soorae and Seddon 1998). These results can be compared with data collected 10 years earlier (Beck et al. 1994) on reintroductions recorded between 1900 and 1992. Interpretation is tentative because a time span of 90 years was compared with a snapshot view in 1998. Furthermore, RSG activities through the 1990s were specifically focused on compiling cases of reintroductions. Nonetheless, Table 5.2 indi-
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table 5.2 The number of species reintroduced in 1900–1992 and those being reintroduced in 1998. Number of Species Taxon
Invertebrates Fish Reptiles and amphibians Birds Mammals Overall
Reintroduced 1900–1992
Being Reintroduced in 1998
Percentage Increase
2 9 22 54 39 126
19 11 42 69 77 218
850 22 91 28 97 73
cates that the number of species being reintroduced in 1998 had increased greatly. The largest increase was for invertebrates. Reptile and mammal reintroductions almost doubled, and those with fish and birds had each increased by 20–25 percent. There has been no analysis as to whether these reintroductions are successful, but it may be inferred that because these efforts appear voluntarily in the directory, the attempts have been successful or are still under way, and there is optimism that they will succeed. There is also no conclusion yet as to whether the species being selected are more endangered than would be expected by chance among the total array of species in these taxa. Thus, one cannot say whether the world’s current reintroductions are addressing conservation priorities. But it is incontestable that although the absolute number of reintroductions is increasing, they involve only very small proportions of the world’s diversity. Reintroductions as a conservation tool have many values, many of which are not biological (Stanley Price 1989), but on numerical grounds they are not affecting a very large number of the world’s species. A large number of threatened animal species are bred out of range, with few individuals released back into the wild. Species may be repatriated before reintroduction to establish in-country captive stocks or reintroduced directly. Recent examples of the latter scenario include the Mhorr gazelle (Gazella dama mhorr) to Senegal (Rau and Wiesner 1999) and the black and white ruffed lemur (Varecia variegata) to Madagascar (Britt et al. 1999). In a number of cases, multiple donors support field programs for highly threatened, large flagship species such as the Sumatran rhino (Khan et al.
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1999) or taxa such as the West Indian iguana group (Hudson and Alberts 1999). For the latter, 26 zoos are listed as supporters. The release of nine black and white ruffed lemurs to reinforce a wild population is the result of collaboration and funding from 29 out-of-country sources (Britt et al. 1999). Presumably this collaboration reflects the number of captive populations and the great significance of attempting to reestablish captive-bred lemurs in the wild for the first time.
Lessons Learned and Future Directions Investments by zoos and botanic gardens often have focused on species with high display values. Thus, a single species may have managed captive and wild populations, each representing different values and resources. For instance, a zoo gorilla has a different set of economic and scientific values attached to it than a wild gorilla. A wild gorilla can be valued as a genetic and demographic contributor to wild populations, as a seed and fruit dispersal agent, as a generator of ecotourism revenue, and as measured in the market price per kilo of bushmeat. In contrast, a captive gorilla is measured in terms of its genetic and demographic contributions to the captive stock, its display and educational value, and its tangible contribution to the zoo’s public, financial, and political viability and image as a conservation facility. However, effective conservation actions must include attention to the habitat, economy, and communities in which the wild or reintroduced populations of any species have to survive. Accordingly, although the majority of zoo gorilla populations are unlikely to contribute to reintroduction activities, this does not necessarily undermine their total conservation value for zoos. Similarly, the majority of plants held in botanic gardens are unlikely to be used for reintroductions, but they can and should be used for broader conservation benefits. Our prediction is that far more conservation measures will take place in the countries that own and host threatened species. These conservation measures will have to intensify in the face of current environmental trends. There is scope for new institutions, new partnership types, and new opportunities. New technologies are available to support the joint genetic management between wild and captive populations. For instance, the Gilman International Conservation project in the Ituri Forest of the Democratic Republic of Congo has extracted okapi (Okapia johnstonii) genes from the wild without depleting the wild population: it captured wild okapi that were bred in penned conditions in-range. Their offspring were taken to captive herds in the United States, diversifying the genetically impover-
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ished stocks, while the breeders were successfully released back into the wild (J. Lukas, pers. comm., 2000). As an analogy, the Millennium Seed Bank, Royal Botanic Gardens, Kew, holds samples of seed from tropical nations currently lacking adequate storage facilities until repatriation is feasible. As technology increasingly permits effective and safe movement of genetic material between in-country and out-of-country sites, more complete metapopulation management may be possible. The rigid distinctions between in situ and ex situ conservation will break down. A number of zoos are showing leadership in managing and developing protected areas. The Wildlife Conservation Society (WCS, Bronx Zoo, New York) is playing a critical role in the development and management of tropical protected areas, far beyond the common role of zoos in identifying important areas and undertaking conservation biology research in them. For example, WCS studies of the threatened Chaco peccary in Bolivia have led to the development of the 8.5-million-acre Kaa-Iya National Park. In a parallel progression, single-species investments in Mauritius funded by the Durrell Wildlife Conservation Trust were instrumental in creating the Black River Gorge National Park. In another case, after WCS research helped identify the need for the Masoala National Park, Madagascar, the park is managed by WCS staff in preparation for a hand over to in-country management (Cohn 2000). The safeguards provided by the CBD prevent any charges of conservation colonialism, and, perhaps more importantly, such roles by outside organizations meet the critical need for cash and resources felt by many in-country conservation agencies. We see a developing role for zoos and botanic gardens, working in partnership with host country agencies and nongovernment organizations, jointly establishing, funding, and managing new protected areas. Zoos and botanic gardens want to have strong conservation records and to be able to demonstrate this success to their members, boards, and supporters. Yet zoos and botanic gardens often do not house the species of greatest ex situ conservation need, nor were the majority of institutions originally designed to be conservation organizations (Maunder et al. 2001a, 2001b; Chapter 1, this volume). One approach is to revise the display and interpretation role of zoos and botanic gardens. A future challenge for outof-country institutions will be the need to effectively support, scientifically and financially, the in situ conservation of species not displayed by that institution or not suitable for captive propagation. New technologies and the power of Internet communications will open up great opportunities for linking conservation activities for the benefit of the species and satisfying
table 5.3 Biodiversity hotspots (sensu Myers et al. 2000), with candidate flagship species for habitat conservation currently managed as ex situ zoo and botanic garden populations. Candidate Flagship Animal Species Managed through American Zoo and Aquarium Association Studbooks or Species Survival Plan Programs
Endemic Vertebrates
Endemic Plants
Tropical Andes
1,567
20,000
Mesoamerica
1,159
5,000
Caribbean
779
7,000
Brazil’s Atlantic Forest Choco, Darien, Ecuador Brazil’s Cerrado
567 418 117
8,000 2,250 4,400
Endemic iguana species, Virgin Islands boa, Cuban crocodile, Caribbean flamingo Golden lion tamarin, jaguar Jaguar Jaguar
61 71
1,600 2,125
Andean condor California condor
771
9,700
Ruffed lemurs, black lemur, mongoose lemur, Rodrigues fruit bat
Hotspot
Central Chile California Floristic Province Madagascar
Andean bear, Andean condor, Chilean flamingo Jaguar
Candidate Flagship Taxa, Often Cultivated in Botanic Gardens
Brugmansia, Puya Neotropical bromeliads and orchids, Agave Caribbean palms (e.g., Sabal and Roystonea) Neotropical bromeliads and orchids Neotropical bromeliads and orchids Brazilian cacti and terrestrial bromeliads Tecophilaea California annuals (e.g., Eschscholzia), shrubs (e.g., Ceanothus) Mauritian and Madagascan palms (e.g., Hyophorbe, Dypsis)
Eastern Arc and Coastal Forests of East Africa Western African forest
121
1,500
Sable antelope
270
2,250
Cape Floristic Province Succulent Karoo Mediterranean Basin
53 45 235
5,700 1,900 13,000
Chimpanzee, western lowland gorilla, mandrill, drill Blue crane Black rhino Waldrapp ibis
Caucasus Sundaland Wallacea
59 701 529
1,600 15,000 1,500
Philippines Indo-Burma
518 528
5,800 7,000
South-Central China
178
3,500
Western Ghats and Sri Lanka
355
2,200
Southwestern Australia New Caledonia
100 84
4,300 2,550
New Zealand Polynesia and Micronesia
136 223
1,860 3,330
Barbirusa, Bali mynah Palm cockatoo, Komodo dragon, birds of paradise Anoa Pigmy hog, white-winged wood duck, Asian elephant Chinese tiger, Chinese alligator Lion-tailed macaque, Asian elephant, Ceylon leopard Koala, western gray kangaroo Kagu Micronesian kingfisher, Hawaiian goose
African violets, Gigasiphon, succulent Euphorbia African Impatiens Protea, Strelitzia, Aloe Lithops, Welwitschia, Aloe Mediterranean bulbs (e.g., Tulipa), Canary Island flora (e.g., Echium) Pterocarya Artocarpus, Citrus, Tectona grandis Artocarpus, Citrus, Nypa palm Strongolydon, Medinilla Amherstia, Tectona Liriodendron chinense, Cunninghamia lanceolata Myristica, Mangifera, Artocarpus, Cinnamomum Eucalyptus, Banksia, Anigozanthos New Caledonian palms, Araucaria columnaris Clianthus, Nothofagus Pritchardia palms, Sophora toromiro
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the information needs of audiences in both sites. The Wild Screen project (Bristol, England) can be viewed as one example of a hybrid facility exhibiting biodiversity through live displays (a rainforest exhibit) and multimedia exhibition. One innovative response to the challenge of linking exhibition with conservation is the Congo exhibit in the Bronx Zoo, which uses live animals to tell the story of the Congo rainforest and its conservation; visitors pay an extra fee to enter it, all of which goes to a conservation project of the visitor’s choice, and a new conservation revenue of $1 million each year is predicted (W. Conway, pers. comm., 1999). We propose that out-of-country ex situ conservation will be led not only by the institution as “genetic ark” but increasingly as facilitator for collaborative research and for public display and fundraising that directly and quantifiably supports in-country activities. Increasingly, in-country ex situ activities will be directly linked to national biodiversity strategies and the global imperatives of habitat and wilderness-scale conservation. The global portfolio of zoo and botanic garden facilities and the associated broad expertise in threatened species management will be used most effectively through field-based projects linked to institution-based research and support teams. An opportunity exists for northern institutions to further use existing conservation investments to actively support conservation in biodiversity hotspots (sensu Myers et al. 2000). The success of the golden lion tamarin can be emulated. Zoos and botanic gardens are maintaining species that originate from and can represent these hotspots (Table 5.3). Some are restricted endemics, such as the African violets (Saintpaulia sp.) that represent the Eastern Arc forests of Tanzania and Kenya. Others are widely distributed and can represent a number of hotspots. We suggest that these “flagship” or “totem” species be used as foci for interpretation, using the public exhibit to explore the broader environmental issues of each region. Accordingly, a botanic garden can use an existing, perhaps poorly documented or genetically redundant collection of threatened species to promote an understanding of the in situ situation, field activities, and in situ conservation. The majority of plant collections held in botanic gardens will not contribute to reintroduction programs for a number of genetic, phytosanitary, and logistical reasons (Maunder et al. 2001a). Accordingly, the tropical plant displays of European and North American botanic gardens, representing massive capital investments, could be converted into conservation gains measured through funds generated and land secured. The Tresor
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Project, managed by the Utrecht Botanical Garden, the Netherlands, is an exciting precedent in that it uses its botanic garden collections and facilities as the basis for fundraising activities to purchase land for rainforest reserves in French Guyana (B. van Wollenberg, pers. comm., 2000). In crude terms it could be argued that this investment translates European botanic garden glasshouse infrastructure into secured tropical habitats.
Conclusions The origins and evolution of botanic gardens have been quite distinct from those of zoos, and we predict a convergence of objectives in which botanic gardens will increasingly support in situ conservation. The zoo experience suggests the following strategies for improving the use of ex situ plant conservation facilities, such as botanic gardens, to support in-country conservation: Linking in-country and out-of-country facilities. Out-of-country facilities, serving an affluent and interested visiting public, should increasingly provide scientific, financial, and managerial support to in-country facilities where indigenous biodiversity can be maintained in a more costeffective manner. Out-of-country facilities will use increasingly sophisticated display techniques and use selected species as flagships to represent the need for wilderness retention, habitat restoration, sustainable use, and the particular conservation needs of threatened taxa. Strategically identifying candidate species. The list of candidate species for ex situ management will increase as habitat areas decline in area and quality (Tilman et al. 1994); ex situ facilities both in country and overseas will need to apply scientifically based tools for identifying priority species and effectively applying their different resources to the management of these species. Conservation organizations need to take a proactive and pragmatic approach to what must be conserved. For example, the Royal Botanic Gardens, Kew, made the strategic and ambitious decision to focus on the long-term seed storage of the United Kingdom’s flora and that of the world’s arid regions. Similarly, the Sharjah Desert Park, United Arab Emirates, exhibits only indigenous Arabian species; its prime aim is to provide educational displays of regional species and to take active measures for their conservation in the wild (C. Gross, pers. comm., 2000). Incorporating species conservation with regional programs. Where possible, conservation efforts and displays for individual species should be inte-
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grated with regional programs for ecosystems and suites of species, such as the Yellowstone-to-Yukon project, which promotes the conservation of large carnivores, or the 95,000-km2 Gaza-Kruger-Gonarezhou Transfrontier Conservation Area for the Republic of South Africa, Mozambique, and Zimbabwe (Turnbull 2001). Such approaches can be developed around the regional or taxonomic initiatives of the Species Survival Commission Specialist Groups. This approach provides a cohesive and identifiable product that is readily marketed and establishes conservation tools and approaches across a wide range of stakeholders. The ex situ facility thus no longer acts as an isolated player but becomes a key player in an extensive and high-impact project. Ex situ facilities can play a more active role in supporting and undertaking ecosystem conservation and restoration. For instance, the reintroduction of the golden lion tamarin has led to a 38 percent increase in the area of protected Atlantic rainforest in the state of Rio de Janeiro, Brazil, and has prompted the development of corridors between isolated forest blocks (Mallinson 1994; Kleiman and Mallinson 1998). Similarly, work in Mauritius involving zoos (Jersey), botanic gardens (Royal Botanic Gardens, Kew), national ministries, and nongovernment organizations secured forest and island habitats, building on conservation efforts initiated for threatened bird and reptile species (Jones et al. 1999). This opens the door for new collaborative relationships with land management agencies and private landowners. Importantly, a focus on restoration can foster an ambitious and optimistic response by an institution, its partners, and local communities. This focus need not be limited to the tropics; facilities can promote the conservation of temperate ecoregions or hotspots such as the Mediterranean basin of Europe or the temperate forests of northeastern America or China. Linking ex situ conservation with the economic use of wild species. Opportunities exist to link displays and propagation facilities with ongoing debates on the economic use of wild populations. These debates may deal with apparently unpalatable messages about unsustainable wild harvests (medicinal plants or bushmeat harvests), human rights, logging, and world trade patterns. For instance, medicinal plant taxa threatened by habitat loss and overharvesting can be displayed in northern facilities to promote public education and political advocacy near the commercial market, with the front-line and in-country institutions providing a complementary and more cost-effective resource for artificial propagation that is linked to a local market or the source wild populations. They can also play an active lobbying
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role in issues pertaining to the sustainable use of wildlife resources. This will become increasingly important as developing world cities grow in size and consumer demands for wild harvested products increase. For example, the European Association of Zoos and Aquaria adopted the current excessive use of bushmeat as its annual campaign for 2001. As a consequence, its membership did the following: • Gathered 1.9 million signatures from 149 institutions in 23 countries
(and a further few associate organizations), which were presented with great publicity to the European Union in November 2001; the aim was to ensure that the unsustainable use of bushmeat was acknowledged in further development policies and overseas development assistance. • Provided an unprecedented cause around which zoos and conservation and welfare organizations rallied collaboratively and productively. • Collectively raised 70,000 Euros for direct support of field actions to reduce bushmeat consumption. • Motivated individual zoos to provide further support; for example, the Durrell Wildlife Conservation Trust won a Darwin Foundation grant for an in-depth study of protein needs, supplies, and shortages in two countries of West Africa, and Bristol Zoo, United Kingdom, raised $75,000 for its own program of wildlife sanctuary support and public education in Cameroon.
This provides a telling example of the increasing capacity of zoos to use their individual and collective “visitor power” in support of conservation causes. Developing in-country facilities for biodiversity hotspots. Conservation investment, from both zoos and botanic gardens, must be focused on the biodiversity-rich developing countries. Consistent with the CBD, there will be a diversification of in-country biodiversity facilities that benefit from inputs by ex situ organizations; these will tend to be in the less developed world (see Maunder et al. 2002). Using international support when necessary, we envisage the development of a range of facilities that exhibit indigenous biodiversity in nearly natural conditions, displaying species that are either characteristic of the country, dramatic, economically valuable, or threatened for which ex situ conservation and display can be combined effectively. These facilities may not necessarily focus only on charismatic
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species; for example, consider the display of dragonflies at the Pietermaritzburg Botanic Garden (Steytler and Samways 1995). These facilities may be sited at the edge of protected areas, with potential for collaborative management between park authorities, local communities, and entrepreneurs, or they may be sited close to urban areas where other opportunities for biodiversity-based recreation are few. In turn, this will tap into the motivation, resources, and expertise of outside institutions, which value international collaboration and partnership within their own objectives for conserving the biodiversity of other nations. Developing the whole conservation role of the institution or facility. In addition to the zoo or botanic garden as a center for the conservation of biodiversity, there is an urgent need for such facilities to engage their stakeholders and to challenge them both intellectually and emotionally in order to change behavior and values. Four roles have been identified by George Rabb (pers. comm., 2001): the institution as model citizen that operates in an environmentally friendly and sustainable manner (for instance, through adhering to Agenda 21); the institution as conservationist operating on site and in the field; the institution as agent for conservation, acting as conservation communicator, inspirer, and motivator to the broader community; and the institution as mentor and trainer, developing staff expertise. The world’s two largest ex situ networks, the zoos and botanic gardens, are playing fundamentally important roles in managing threatened species and supporting habitat conservation. The world’s zoos and botanic gardens have developed different approaches to the ex situ management of threatened species, based on their respective antecedents. Zoos have developed a conservation breeding role in response to professional and public concerns about wild collected animals and welfare; in contrast, the botanic garden community has not been subject to vigorous public debate on its role or value and has not adopted scientific management protocols for the majority of collections. The world’s zoos are maintaining global stocks of key charismatic species, flagships for the conservation world and valued display subjects. It is hoped that these species can be used to raise investments for the conservation of key habitats and wildernesses. This focus has generated expertise that should be widely applied to conservation at local and regional levels. In contrast, the world’s botanic gardens are not tied to managing a group of flagship species and, accordingly, are already devoting more resources to local and regional conservation priorities. Their challenge is to extend this strength collaboratively to use the extensive public
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displays of the northern botanic gardens to support global conservation. Both groups are facing the challenge of displaying biodiversity to an increasingly biologically illiterate population; accordingly, a major challenge is to communicate the fundamental message that biodiversity is vitally important in sustaining life. Both botanic gardens and zoos need to work collectively henceforth, bringing together complementary tools and strengths in support of the world’s areas of diversity and wilderness. There is great scope for ensuring that these activities support in-country species recovery and habitat management. The conservation responsibilities of ex situ facilities will evolve further to ensure that in-country and out-of-country facilities will play carefully integrated and complementary roles, all focused on the retention of viable wild populations and habitats.
Acknowledgments The authors would like to thank the following for their support and guidance: Dr. George Rabb, Brookfield Zoo; the late Dr. Ulie Seal, Conservation Breeding Specialist Group of the Species Survival Commission and IUCN; Dr. Clare Hankamer, Royal Botanic Gardens, Kew; Jeremy Mallinson and John Hartley, Durrell Wildlife Conservation Trust, Jersey, United Kingdom. References AZA (American Association of Zoos and Aquaria). 1999. The Collective Impact of America’s Zoos and Aquariums. Silver Spring, MD: American Association of Zoos and Aquaria. Balmford, A., N. Leader-Williams, and M. J. B. Green. 1995. Parks or arks: where to conserve threatened mammals. Biodiversity and Conservation 4:595–607. Balmford, A., G. M. Mace, and N. Leader-Williams. 1996. Designing the ark: setting priorities for captive breeding. Conservation Biology 10:719–727. Beck, B. B., L. G. Rapaport, M. R. Stanley Price, and A. Wilson. 1994. Reintroduction of captive-born animals. Pages 265–286 in P. J. S. Olney, G. Mace, and A. T. C. Feistner (eds.), Creative Conservation: Interactive Management of Wild and Captive Animals. Proceedings of the Sixth World Conference on Breeding Endangered Species. London: Chapman & Hall. BGCI (Botanic Gardens Conservation International). 2001. Botanic Garden Agenda for Conservation. London: Botanic Gardens Conservation International. Britt, A., A. Katz, and C. Welch. 1999. Project Betampona: conservation and restocking of black and white ruffed lemurs (Varecia variegata variegata). Pages 87–94 in T. L. Roth, W. F. Swanson, and L. K. Blattman (eds.), Proceedings of Seventh World Conference on Breeding Endangered Species: Linking Zoos and Field Research to Advance Conservation. Cincinnati: Cincinnati Zoo and Botanic Garden.
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Chesser, R. K., M. H. Smith, and I. L. Brisbin Jr. 1980. Management and maintenance of genetic variability in endangered species. International Zoo Year Book 20:146–154. Coe, J. C. 1985. Design and perception, making the zoo experience real. Zoo Biology 4:197–208. Cohn, J. P. 2000. Working outside the box: zoos and aquariums are shifting the focus of their conservation efforts to the wild. BioScience 50(7):564–569. Conway, W. 1995. Wild and zoo animal interactive management and habitat conservation. Biodiversity and Conservation 4:573–594. Conway, W. 1996. The in situ conservation program of the Wildlife Conservation Society. International Zoo News 43(5):274–278. Davis, S. G. 1997. Spectacular Nature: Corporate Culture and the Sea World Experience. Berkeley: University of California Press. Durrell, L., and J. J. C. Mallinson. 1998. The impact of an institutional review: a change of emphasis towards field conservation programs. International Zoo Yearbook 36:1–8. Ebenhard, T. 1995. Conservation breeding as a tool for saving animal species from extinction. Trends in Ecology and Evolution 10:438–443. Ellis, S., and U. S. Seal. 1995. Tools of the trade to aid decision making for species survival. Biodiversity and Conservation 4:553–572. Esson, M., and M. Tomlinson. 1997. Gorilla society: an educational tool to promote the development of group dynamics in teenagers. Dodo: Journal of the Wildlife Preservation Trusts 33:126–136. Fiumera, A. C., L. Wu, P. G. Parker, and P. A. Fuerst. 1999. Effective population size in the captive breeding program of the Lake Victoria cichlid Paralabidochromis chilotes. Zoo Biology 18:215–222. Flessness, N. 1977. Gene pool conservation and computer analysis. International Zoo Yearbook 17:77–81. Foose, T. J., and J. D. Ballou. 1988. Population management theory and practice. International Zoo Yearbook 27:26–41. Glowka, L., F. Burhenne-Guilman, H. Synge, J. A. McNeely, and L. Gündling. 1994. A Guide to the Convention on Biological Diversity. Environment Policy and Law paper no. 30. Gland, Switzerland: IUCN. Grajal, A. 1999. Biodiversity and the nation state: regulating access to genetic resources limits biodiversity research in developing countries. Conservation Biology 13:6–10. Groombridge, B. (ed.). 1993. 1994 IUCN Red List of Threatened Animals. Gland, Switzerland: IUCN. Hancocks, D. 2001. A Different Nature: The Paradoxical World of Zoos and Their Uncertain Future. Berkeley: University of California Press. Holdgate, M. 1999. The Green Web: A Union for World Conservation. London: Earthscan. Hudson, R., and A. Alberts. 1999. An overview of zoo supported conservation programs for West Indian iguanas. Pages 227–236 in T. L. Roth, W. F. Swanson, and L. K. Blattman (eds.), Proceedings of Seventh World Conference on Breeding Endangered Species: Linking Zoos and Field Research to Advance Conservation. Cincinnati: Cincinnati Zoo and Botanic Garden. Hutchins, M., K. Willis, and R. Wiese. 1995. Strategic collection planning: theory and practice. Zoo Biology 14:5–24.
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IUCN (International Union for the Conservation of Nature). 1987. Captive Breeding Policy. IUCN Policy Document. Gland, Switzerland: IUCN. IUCN/UNEP/WWF. 1991. Caring for the Earth: A Strategy for Sustainable Living. Gland, Switzerland: IUCN. IUDZG/CBSG/IUCN/SSC. 1993. The World Zoo Conservation Strategy: The Role of Zoos and Aquaria of the World in Global Conservation. Chicago: Chicago Zoological Society. Jones, C. G., K. Swinnerton, J. Hartley, and Y. Mungroo. 1999. The restoration of freeliving populations of the Mauritius kestrel (Falco punctatus), pink pigeon (Columba mayeri) and echo parakeet (Psittacula eques). Pages 77–86 in T. L. Roth, W. F. Swanson, and L. K. Blattman (eds.), Proceedings of Seventh World Conference on Breeding Endangered Species: Linking Zoos and Field Research to Advance Conservation. Cincinnati: Cincinnati Zoo and Botanic Garden. Khan, M. K. M., T. L. Roth, and T. J. Foose. 1999. In situ and ex situ efforts to save the Sumatran rhinoceros (Dicerorhinus sumatrensis). Pages 163–174 in T. L. Roth, W. F. Swanson, and L. K. Blattman (eds.), Proceedings of Seventh World Conference on Breeding Endangered Species: Linking Zoos and Field Research to Advance Conservation. Cincinnati: Cincinnati Zoo and Botanic Garden. Kleiman, D. G., and J. J. C. Mallinson. 1998. Recovery and management committees for lion tamarins: partnerships in conservation planning and implementation. Conservation Biology 12:27–38. Lacy, R. C. 1987. Loss of genetic diversity from managed populations: interacting effects of drift, mutation, immigration, selection, and population subdivision. Conservation Biology 1:143–158. Lefkowitz-Horowitz, H. 1996. The National Zoological Park: ‘city of refuge’ or zoo? Pages 126–136 in R. J. Hoage and W. A. Deiss (eds.), New Worlds, New Animals: From Menagerie to Zoological Park in the Nineteenth Century. Washington, DC: John Hopkins University Press. Mace, G. M., P. Pearce Kelly, and D. Clarke. 1998. An integrated conservation program for the tree snails (Partulidae) of Polynesia: a review of captive and wild elements. Journal of Conchology S2:89–96. Mallinson, J. J. C. 1994. Saving the world’s richest rainforest. Biologist 41:57–60. Maunder, M., S. Higgins, and A. Culham. 2001a. The conservation value of botanic garden plant collections: a European case study. Biodiversity and Conservation 10:383–401. Maunder, M., B. Lyte, J. Dransfield, and W. Baker. 2001. The conservation value of botanic garden palm collections. Biological Conservation 98:259–271. Maunder, M., M. R. Stanley Price, P. Soorae, and S. Mashuari. 2002. The role of tropical botanic gardens in supporting species and habitat recovery: East African opportunities. Pages 115–134 in M. Maunder, C. Hankamer, C. Clubbe, and M. Groves (eds.), Plant Conservation in the Tropics: Principles and Experiences. Kew, UK: Royal Botanic Gardens. Mittermier, R. A., and W. R. Konstant. 1999. Hotspots and wilderness areas: setting priorities in biodiversity conservation. Pages 49–62 in T. L. Roth, W. F. Swanson, and L. K. Blattman (eds.), Proceedings of Seventh World Conference on Breeding Endangered Species: Linking Zoos and Field Research to Advance Conservation. Cincinnati: Cincinnati Zoo and Botanic Garden. Mugabe, J., C. Barber, G. Henne, L. Glowka, and A. La Viña. 1996. Managing access
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to genetic resources: towards strategies for benefit sharing. Biodiversity Bulletin 1:14–15. Myers, N., R. A. Mittermeier, C. G. Mittermeier, G. A. B. da Fonseca, and J. Kent. 2000. Biodiversity hotspots for conservation priorities. Nature 403:853–858. Preece, D. J. 1998. The captive management and breeding of poison dart frogs, family Dendrobatidae, at Jersey Wildlife Preservation Trust. Dodo 34:103–114. Rabb, G. 1994. The changing roles of zoological parks in conserving biological diversity. American Naturalist 34:159–164. Rabb, G., and T. A. Sullivan. 1995. Coordinating conservation-global networking for species survival. Biodiversity and Conservation 4:536–543. Rau, B., and H. Wiesner. 1999. Captive breeding and the reintroduction of the Mhorr gazelle (Gazella dama mhorr). Pages 95–107 in T. L. Roth, W. F. Swanson, and L. K. Blattman (eds.), Proceedings of Seventh World Conference on Breeding Endangered Species: Linking Zoos and Field Research to Advance Conservation. Cincinnati: Cincinnati Zoo and Botanic Garden. Reade, L. S., and N. K. Waran. 1996. The modern zoo: how do people perceive zoo animals? Applied Animal Behavior Science 47:109–118. Robinson, M. H. 1992. Global change, the future of biodiversity and the future of zoos. Biotropica 24:345–352. Ryder, O. A., and A. T. C. Feistner. 1995. Research in zoos: a growth area in conservation. Biodiversity and Conservation 4:671–677. Sheppard, C. 1995. Propagation of endangered birds in U.S. institutions: how much space is there? Zoo Biology 14:197–210. Snyder, N. F. R., S. R. Derrickson, S. R. Beissenger, J. W. Wiley, T. B. Smith, W. D. Toone, and B. Miller. 1996. Limitations of captive breeding in endangered species recovery. Conservation Biology 10:338–348. Soorae, P. S., and P. J. Seddon (eds.). 1998. Reintroduction Practitioners Directory 1998. Published jointly by the IUCN Species Survival Commission’s Reintroduction Specialist Group, Nairobi, Kenya and the National Commission for Wildlife Conservation and Development, Riyadh, Saudi Arabia. Soulé, M., M. Gilpin, W. Conway, and T. Foose. 1986. The millennium ark: how long a voyage, how many staterooms, how many passengers? Zoo Biology 5:101–113. Spencer, L. 1993. Zoo and wildlife veterinarians examine their role in conservation. Journal of the American Veterinary Medical Association 202:714–717. Stanley Price, M. R. 1989. Animal Reintroductions: The Arabian Oryx in Oman. Cambridge, UK: Cambridge University Press. Steytler, N. S., and M. J. Samways. 1995. Biotope selection by adult male dragonflies (Odonata) at an artificial lake created for insect conservation in South Africa. Biological Conservation 72:381–386. Templeton, A. R., and B. Read. 1984. Factors eliminating inbreeding depression in a captive herd of Speke’s gazelle. Zoo Biology 3:177–199. Tilman, D., R. M. May, C. L. Lehman, and M. A. Nowak. 1994. Habitat destruction and the extinction debt. Nature 371:65–66. Turnbull, M. 2001. Breaking boundaries: game parks—the next generation. Africa, Environment and Wildlife 9(4):58–71. Ward, P. I., N. Mosberger, C. Kiestler, and O. Fischer. 1998. The relationship between popularity and body size in zoo animals. Conservation Biology 12:1404–1411. Wayne, R. K., L. Forman, A. K. Newman, J. M. Simonson, and S. J. O’Brien. 1986.
5. Ex Situ Support to the Conservation of Wild Populations and Habitats Genetic monitors of zoo populations: morphological and electrophoretic assays. Zoo Biology 5:215–232. Westley, F., and H. Vredenburg. 1997. Interorganizational collaboration and the preservation of global biodiversity. Organization Science 8(4):381–402. World Resources Institute. 1998. World Resources Report 1998–99. Environmental Change and Human Health. Oxford, UK: Oxford University Press for WRI. WRI/IUCN/UNEP. 1992. Global Biodiversity Strategy. Baltimore, MD: World Resources Institute Publications. Wyse Jackson, P. 2001. An international review of the ex situ plant collections of the botanic gardens of the world. Botanic Gardens Conservation News 3(6):22–33.
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part two
Tools of the Trade
The tools for ex situ conservation may be traced to two origins: the traditional horticultural techniques used for living collections and more modern science-based techniques derived largely from the agricultural and plant genetic resource community. Applied horicultural research seems to have had neither the cachet nor made the dramatic technical advances in recent years that germplasm management techniques for seed, pollen, and tissue have made. These recent advances make possible increasingly effective storage of plant material over extended periods of time. Nevertheless, being able to grow plants is still basic to success. The storage of seed as a conservation method is well established and is increasingly promoted as a cost-effective and efficient means of storing large numbers of genotypes over a long time period. The success of these tools can be gauged by the increasing number of working seed storage facilities maintained by plant conservation teams. The application of this approach to a wide variety of wild species has led to a greater understanding of seed storage physiology. Walters (Chapter 6) and Pritchard (Chapter 7) both review the physiology of seed storage and its implications for ex situ conservation. Walters examines the physiology of deterioration kinetics and its relevance to successful long-term storage. Pritchard reviews the types of seed behaviors and proposes a revised terminology that is derived from assessing the storage behavior of a wide variety of plants from both temperate and tropical conditions. These two chapters show that the expanded understanding of seed physiology opens the door to storing a greater number of species previously thought to be recalcitrant. Baskin and Baskin (Chapter 8) bring their extensive knowledge of the ecophysiology of seed dormancy and apply it to the practical challenge of getting the most information on dormancy-breaking and germination requirements from the least amount of seed. They show that through an understanding of a species’ ecology and the behavior of confamilial taxa it
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is possible to assign taxa to a small number of morphophysiological dormancy types. The Baskins summarize in a pair of dichotomous keys how best to determine the dormancy type of a sample. They then provide a model “move-along” experimental protocol designed to determine dormancy breaking and germination requirements of seeds using as few seeds as possible. Although it is little used in the conservation of wild plants, pollen storage has been widely used to support agricultural and horticultural breeding programs. Towill (Chapter 9) reviews application of pollen storage to threatened species management and provides practical recommendations for its storage and use. In vitro techniques, particularly micropropagation, are being widely used for threatened plant propagation and storage. Sugii and Lamoureux (Chapter 10) outline its practical application in a notorious extinction hotspot, the Hawaiian Islands. The success of their program has resulted from the development of taxon-specific protocols, meticulous hygiene, and a close working relationship with field collectors, land managers, regulatory agencies, and recovery teams. The available tools for angiosperms and gymnosperm conservation are well developed compared with those for bryophytes and pteridophytes. Pence (Chapter 11) reviews conservation issues for these two groups and assesses the available tools for ex situ management. The developing tools for ex situ conservation, particularly for seed samples, have helped reduce one of the biggest challenges to ex situ application: the management of an increasing number of species by a small cadre of institutions. Research investments made by international agencies such as the International Plant Genetic Resources Institute and by leading research facilities (the U.S. Department of Agriculture’s National Center for Genetic Resources Preservation and the Royal Botanic Gardens, Kew, Millennium Seed Bank) are, at least for temperate climates, allowing the cost-effective storage of large samples for many species.
Chapter 6
Principles for Preserving Germplasm in Gene Banks Christina Walters
Gene banks are an ex situ conservation strategy designed to capture and conserve genetic diversity within and among species. In gene banks, germplasm is placed in suspended animation so that desirable allelic combinations and rare alleles of a species are available for the future. Plant germplasm is really a collection of propagules: seeds or pollen to preserve the genetic composition of populations, and cuttings, buds, rhizomes, or cell cultures to preserve specific genetic combinations of individuals. Most gene banks have a completely utilitarian goal: plant breeders use collections to make higher-yielding, more resistant crops, and ecologists preserve threatened populations until they can be reintroduced into restored habitats. Whether for agricultural or landscape management purposes, genetic diversity is needed for a species to establish and adapt to a changing environment. To conserve genetic diversity, collections should consist of many individual propagules from several populations (Appendix 1, this volume). The appropriate number of individuals and populations depends on the genetic variability of the species within and among its populations, the life history attributes of the species, the geographic distribution of the species, and the extent to which a particular population is locally adapted. Accession sizes of 2,000 (inbred) or 5,000 (outcrossed) individuals are desirable but often impossible if propagules are harvested from wild populations and not regenerated before they are placed in storage. The number of accessions for a particular species varies widely between species and gene banks and is often a matter of convenience or opportunity rather than a science-based study of genetic composition. Which species are represented in a particular gene 113
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bank depends on the overall mission of that gene bank. For example, agricultural gene banks focus on species of present or potential economic importance and their congeners. In the United States, about 10,000 species are needed to provide food, fiber, and pharmaceuticals. If economically important is more broadly defined to encompass land management issues, several thousand more species of native plants will be included in gene banks for U.S. agriculture. Simple arithmetic shows that gene banks are often quite large. In 2002, the U.S. Department of Agriculture National Plant Germplasm System consisted of more than 450,000 accessions, mostly of seeds with 2,000 to 5,000 individuals each, representing about 10,000 species. About 85 percent of the collection is backed up as a base collection at the U.S. Department of Agriculture National Center for Genetic Resources Preservation (formerly the National Seed Storage Laboratory). Maintaining a germplasm collection containing many genetically distinct accessions of hundreds or thousands of species is the challenge for gene bank operators. Genetic shifts resulting from mortality or regeneration of small populations are minimized by ensuring that the individuals within the collection remain highly vigorous. Therefore, millions of individuals are placed under conditions where they don’t age or, more realistically, where aging is slowed so that regeneration frequency is reduced. Typically, lifespans of more than 30 years are necessary for germplasm stored in gene banks; however, lifespans of 200 years or more are desirable and, with proper storage conditions, possible.
Fundamental Principles of Preservation The lifespan that can be achieved for a propagule varies according to the species, tissue or cell type, developmental stage, initial health of propagules, and chemical composition of the cells, in addition to the storage conditions used. Despite what may appear to be unlimited variables, a set of unifying principles exists to allow curators to predict the storage physiology, the optimum conditions for storage, and the frequency of viability monitoring for a wide range of propagules. The underlying precept, controlling deteriorative reactions in biological materials, is common for all life forms. Because of different life histories, developmental physiologies, and cell characteristics, germplasm preservation technologies for vertebrates, invertebrates, microbes, and plants have sometimes been approached from different perspectives. The following discussion focuses primarily on the
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approach used to develop preservation technologies for seeds and pollen, and in these systems deteriorative reactions are defined as damage resulting from desiccation or freezing stresses or prolonged storage. The easiest way to prevent the physical and chemical reactions that cause deterioration in biological materials is to reduce the water content or the temperature, common practices in households where flour is stored in a dry place and milk is placed in the refrigerator. Lowering the moisture level and temperature of cells alters the thermodynamic status of the intracellular matrix where the reactions of life occur. Reactions are slowed, and the propagule becomes quiescent. Unfortunately, the cold, dry conditions needed to stop cellular reactions usually are lethal, and germplasm preservation requires that the vitality of the propagule be maintained. Therefore, preservation technology seeks a balance between cell damage by desiccation, freezing, and aging. The intrinsic tolerance level that some propagules have toward desiccation, freezing, and aging dictates the flexibility that a gene bank operator has in selecting preservation protocols. For example, seeds and pollen from many plant species tolerate extreme drying. Water, which might otherwise freeze and form ice crystals that damage cell membranes, can be removed from the cells of these propagules, enabling the gene bank operator to reduce the storage temperature without imposing a freezing stress. This remarkable tolerance of desiccation stress facilitates germplasm preservation and explains why gene banks are a cost-effective method to preserve genetic diversity for many plant species. Desiccation-tolerant seeds and pollen are called “orthodox” (Roberts 1973; Walters et al. 2002; Chapter 7, this volume) because their longevity in storage increases with decreasing water content and temperature. Many other propagules in the plant, animal, and microbe kingdoms do not survive removal of all freezable water and therefore must be stored at nonfreezing temperatures or under conditions in which intracellular freezing is avoided. Such seeds and pollen are called “recalcitrant” (Roberts 1973; Walters et al. 2002; Chapter 7, this volume). Propagules that overwinter in temperate climates may naturally acquire mechanisms to avoid intracellular ice formation (Thomashow and Browse 1999). Otherwise, protectants may be applied exogenously. Cryoprotective technologies were first developed in the 1950s for mammalian semen (Polge et al. 1949), adapted to mammalian embryos in the 1970s (Whittingham et al. 1972), and regularly applied to plant cell and tissue cultures in the 1980s and 1990s (Bajaj 1995). The basic princi-
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ples of freeze avoidance are discussed later in this chapter after an initial discussion of how cell constituents, moisture levels, and temperatures induce quiescence, allowing germplasm to be preserved. The rate of any reaction is described as the quotient of the thermodynamic forces that allow a reaction to occur spontaneously divided by the barriers that prevent the reaction from occurring. This relationship is described in the generic form J = ∆G R
(6.1)
where J is the rate of reaction, ∆G is the free energy difference between the reactants and products, and R is resistance to energy flow. Because reactions occur only if there is a net loss of free energy (∆G < 0), ∆G governs whether a reaction is possible. The parameter ∆G is a complex function of the concentration of reactants and products and the temperature. The free energy (G) of each chemical constituent of the deteriorative reactions is expressed as the effective concentration, or activity, of that chemical (aa, where 0 < aa < 1) and the temperature, according to the equation G/na = µa = RT ln (aa)
(6.2)
where na = the number of moles of substance a, µa is the molar free energy of substance a (or chemical potential of a), R is the ideal gas constant, and T is temperature in Kelvin. The free energy difference (∆G) is the sum of the free energies of all the products minus the sum of all the free energies of the reactants. Sometimes a reaction that is thermodynamically favored does not occur because of significant barriers that resist the reaction. Consider a sealed jar of water on a warm, dry day. The free energy difference between the water vapor outside the jar (the product) and the liquid water in the jar (the reactant) suggests that the liquid water should evaporate (the reaction). Evaporation does not occur because the lid and walls of the jar provide an effective barrier to water movement. Resistance factors may be in the form of physical barriers (e.g., a reactant molecule must penetrate a film that has extremely small pores) or motional barriers (e.g., molecules move too slowly to allow chemical reactions). The barrier or resistance term of Equation 6.1 can have a profound effect on reaction kinetics. To slow reactions, the difference in energy between products and reactants must be reduced (|∆G| → 0) or the energy change must be blocked (R → ∞). Though simple, the principles embodied in Equation 6.1 are not easy
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to apply to deteriorative reactions, mostly because the specific reactions are poorly understood (for seeds see Walters 1998b; Vertucci and Farrant 1995; Smith and Berjak 1995). We are uncertain about the reactants and products and even less certain about the relationship between the concentrations of these molecules and their chemical potential (µa). Despite our inabilities to determine precise values for Equation 6.1, we can use this important equation to provide a framework for understanding how intrinsic properties of the propagule and conditions of storage fundamentally affect rates of deterioration resulting from desiccation, freezing, or aging damage.
Cellular Constituents and Reaction Kinetics If moisture level and temperature are held constant, cells with greater concentrations of reactants—either because of genetic factors, developmental status, or growth and postharvest conditions—will be predisposed to rapid deterioration. Free radical–induced peroxidation is believed to cause significant deterioration after desiccation or freezing stresses and during storage (Priestley 1986; McKersie 1991; Smirnoff 1993; Hendrey 1993; Smith and Berjak 1995; Pammenter and Berjak 1999). Consequently, treatments that enhance free radical production are expected to enhance the rate of deteriorative reactions and reduce the overall lifespan of the propagule. Free radicals are produced in highly metabolic cells or cells previously stressed so that metabolism is unbalanced (Leprince et al. 2000; Walters et al. 2002). These types of cells are notoriously difficult to preserve and may explain why seeds that are harvested prematurely or experience adverse postharvest conditions deteriorate rapidly in cold, dry storage (Rasyad et al. 1990; Tarquis and Bradford 1992; Hay et al. 1997; Walters 1998b). Unsaturated fatty acids are particularly susceptible to free radical attack (Priestley 1986; Chan 1987), leading to the hypothesis that seeds with high lipid contents or high levels of polyunsaturated fatty acids are more susceptible to deterioration (Priestley 1986; Smith and Berjak 1995). Protective substances slow deteriorative reactions. This is accomplished by affecting the value of ∆G or R in Equation 6.1. Research over the last decade on desiccation and freezing stress in plants has focused on cellular constituents produced by the plant as it acclimates to these stresses (Thomashow and Browse 1999; Vertucci and Farrant 1995; Pammenter and Berjak 1999; Phillips et al. 2002). For some gene products, the mechanism of protection is understood. For example, antioxidants slow deterio-
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ration by reducing free radical levels (Foyer et al. 1994; Pammenter and Berjak 1999), effectively making ∆G less negative. Upon exposure to environmental cues, plant cells produce many other protectants; their mode of action remains conjectural. Sugars and hydrophilic proteins are known to accumulate in plants as they become more stress tolerant (reviewed by Thomashow and Browse 1999; Phillips et al. 2002), but the mechanism of protection may be attributed to an effect on ∆G or R. If these putative protectants bind to macromolecules, making them resistant to chemical or structural alterations (e.g., hypotheses discussed by Crowe et al. 1997 [sugars]; Dure 1993; Close et al. 1993 [LEA (late-embryonic-abundant) proteins]), the effective concentration of reactants (aa in Equation 6.2) is reduced and, consequently, ∆G becomes less negative. These highly soluble substances can also alter the effective water concentration, which, in turn, affects the relative concentration or activity of reactants (Wolfe and Bryant 1999). Finally, these substances can alter the viscosity of the aqueous matrix where deteriorative reactions occur (Slade and Levine 1991a, 1991b; Williams et al. 1993; Wolfe and Bryant 1999; Koster et al. 2000; Walters et al. 2002). Because viscosity (or the reciprocal function, fluidity) affects the ability of reactant molecules to move, then collide and react, it is an extremely important component of the resistance parameter, R, in Equation 6.1. Research in the seed literature has focused mostly on the special case in which viscosity increases so much that an aqueous glass is formed (Williams et al. 1993; Leopold et al. 1994; Walters 1998b; Chapter 7, this volume); however, there are several quasidiscrete changes in viscosity that are also used to describe stability of foods and seeds (Slade and Levine 1991a, 1991b; Buitink et al. 1998a, 2000; Walters et al. 2002). In a glass, molecular motion is so restricted that the material behaves like a solid (i.e., it doesn’t flow). Unlike crystals, molecules in glasses are not arranged in regular patterns, and so the material is technically not a true (thermodynamic) solid. Ranges of viscosity can be visualized by comparing the fluidity of corn syrup, taffy, a lollipop (a glass), and sugar granules (a true solid). Note that the sugar-water solution of the lollipop becomes fluid if the lollipop is warmed or diluted (by licking).
Temperature and Reaction Kinetics Like the chemical constituents of cells, temperature can have profound effects on kinetics of deteriorative reactions. Both the ∆G and R parame-
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ters of Equation 6.1 are affected by temperature, and so the relationship between temperature and reaction rates can be complex. The overall effect of temperature on the free energy change of a reaction is expressed by ∆G = ∆H – T∆S
(6.3)
where ∆H and ∆S describe the changes in enthalpy and entropy, respectively, resulting from the reaction, and T is temperature. Aging reactions are believed to result in a net increase in disorder (∆S > 0), so a rise in temperature makes ∆G more negative and the reaction becomes more likely. Alternatively, desiccation and freezing reactions tend to reduce entropy (e.g., fluid to gel phase changes and demixing in membranes, liquid water to ice; reviewed by Walters et al. 2002), suggesting that a rise in temperature would oppose damaging reactions. Often storage stability involves several reactions with differing temperature dependencies, and this results in a complex relationship between optimum storage temperature and shelf life, as is shown for short-term storage of recalcitrant seeds where reactions leading to germination, chilling injury, and dormancy breaking must be balanced to minimize change (Chapter 7, this volume). Temperature also affects the resistance factor, R, in Equation 6.1. A reaction that is thermodynamically favorable (negative ∆G) may not occur because there is a large energy barrier, called activation energy, to overcome. Raising the temperature increases the molecular motions of molecules, thereby increasing the number of intermolecular collisions, which in turn may allow the reaction to proceed. Temperature-induced changes in intracellular viscosity of seeds and pollen are likely to have profound effects on aging rates of stored germplasm (Buitink et al. 1998a, 2000). Near phase or state transition temperatures, small changes in temperature result in large viscosity changes; therefore, gene bank operators should seek to store germplasm well below the glass transition temperature (Chapter 7, this volume).
Water and the Nature and Kinetics of Reactions A fundamental truth in biology is the need for water, yet we have a poor understanding of the basis of this need. In thermodynamic terms, water can be regarded as a reactant (directly contributing to ∆G), as a dilutant (contributing to ∆G by affecting the chemical potential of other molecules), or as a fluid matrix allowing the diffusion of other chemical con-
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stituents (contributing to R). Changing water contents may change the effective concentrations of reactants and products in the aqueous phase, shifting the likelihood of a reaction. Complete drying may expose reactive surfaces of metals or molecules, increasing the concentration of free radicals or denaturing components. By affecting the overall viscosity, drying also mediates changes in the mobility of dissolved or suspended molecules. To a limit, the drier the medium becomes, the more viscous it becomes until it is essentially a solid matrix trapping molecules, that is, a glass (Slade and Levine 1991a, 1991b; Williams et al. 1993; Leopold et al. 1994; Buitink et al. 1998a; Wolfe and Bryant 1999; Figure 6.1). Water has a profound effect on the nature and kinetics of reactions that occur in an aqueous matrix. Model studies suggest that critical moisture levels affect enzyme activity or membrane function (Acker 1969; Rupley et al. 1983; Wolfe 1987). Labuza (1980) and Karel (1980) reviewed changes in the kinetics of different reactions that degrade foods. Clegg (1986) mapped changes in metabolism of Artemia cysts with hydration. Later, the question was extended to the study of seeds (reviewed by Vertucci and Farrant 1995), and more recent studies support the general pattern (Leprince and Hoekstra 1998; Leprince et al. 2000; Pritchard and Manger 1998; Farrant and Walters 1998; Walters et al. 1997, 2001, 2002; Chapter 7, this volume). In all the studies, the kinetics of reactions varied with water content, and some reactions were not observed when water levels were too high or too low. Critical moisture contents for physiological activity prompted the hypothesis of five hydration levels in seeds that correspond to cells’ ability to support growth (level V, Figure 6.1), affect stress-related metabolism (level IV), respire (level III), carry out catabolic reactions (level II), and be almost in stasis (level I; reviewed by Vertucci and Farrant 1995; Walters et al. 2002). In experiments conducted at temperatures between 20°C and 25°C, changes in hydration levels correspond to water activities (aw in Equation 6.2 where the subscript is changed from a to w to indicate that the chemical is water) of about 0.99, 0.97, 0.89, and 0.22, which correspond to water potentials (w = µw divided by molar volume of water) of –1.5, –4, –15, and –200 MPa (Vertucci and Farrant 1995). Pure water has a water activity of 1.0 and a water potential of 0. Within each moisture level, the rate of reactions increased with increasing water content. Critical water activities or water potentials that define hydration levels are similar between different species and tissue types (Roberts and Ellis
Figure 6.1 Hydration levels in seeds at ambient temperatures. The physical properties of water and
physiological activity of seeds change within specific water potential ranges. Map is adapted from Vertucci and Farrant (1995) and Walters et al. (2002, and references therein), with additional references from this text. Water contents corresponding to the first three hydration levels can be gleaned from isotherms in Figure 6.2.
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1989; Vertucci and Farrant 1995), and this suggests that the free energy term, ∆G in Equation 6.1, depends on the water concentration. The equilibrium relationship between water content and its effective concentration is described using water sorption isotherms. Because of the different instrumentation involved, isotherms usually are expressed in terms of water potential when w > –12 MPa or in terms of relative humidity (RH = aw 100) when w < –12 MPa (≈90 percent RH). At RH < 90 percent, isotherms of many biological materials, including desiccation-tolerant seeds, have reverse sigmoidal shapes (Walters 1998b; Figure 6.2). Water contents decrease steeply as RH decreases to about 70 percent, changes slightly as RH decreases further to about 20 percent, and then decrease sharply again as RH decreases below about 20 percent. Isotherm shape often is attributed to water interactions on macromolecule surfaces (D’Arcy and Watt 1970; Vertucci and Leopold 1987a). The composition of dry matter reserves is the predominant factor explaining isotherm shapes. Seeds containing high amounts of starch and protein have greater water contents at a given RH < 90 percent (Ellis et al. 1989; Vertucci and Roos 1990, 1993; Walters 1998b) because water is partitioned between the aqueous cytoplasm and the hydrophilic storage reserves in seeds. In lipid-rich seeds, the water is located predominantly in the aqueous matrix; thus, on a total dry matter basis, the water content at a given RH is less. Studies linking isotherm shape to sorbent properties continued through the 1980s, and it was generally agreed that there were strong and intermediate interactions of water molecules with sorbent surfaces at low humidities (Bull 1944; D’Arcy and Watt 1970; Rupley et al. 1983; Vertucci and Leopold 1987b). At higher humidities, water-sorbent interactions were sufficiently weak that water-water interactions dominated (Bull 1944; D’Arcy and Watt 1970; Rupley et al. 1983). Changes in water properties, measured using calorimetry, nuclear magnetic resonance, electron spin resonance, infrared spectroscopy, and electrical conductance, coincided with changes in isotherm slopes and also corresponded to changes in enzyme or physiological activity (Rupley et al. 1983; Clegg 1986; Vertucci and Farrant 1995; Walters 1998b). A conceptual model used to rationalize changes in water properties and water sorption onto different surfaces was proposed by Rupley et al. (1983) in which exposed charged sites, hydrophilic sites, and bridges over hydrophobic sites were progressively filled in hydration levels I, II, and III. Water filling capillary pores (level IV) and water in dilute
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Figure 6.2 Water sorption isotherms of seeds of varying lipid composition at
25°C. Lipid contents of seeds are 2, 20, 37, 45, 58, and 71 percent of dry mass for pea, soybean, lettuce, peanut, sunflower, and yew, respectively. Vertical lines represent boundaries between hydration levels given in Figure 6.1. (Data from Walters 1998.)
solutions (level V) led to complete hydration (Vertucci and Farrant 1995; Walters et al. 2002; Figure 6.1). An alternative hydration model is based on the changes in viscosity as aqueous solutions dry. Food scientists describe the viscosity of a drying solution as syrupy, rubbery, and eventually glassy (Slade and Levine 1991a, 1991b). The changes in viscosity are not discrete but roughly correspond to the sorption isotherm shape at 25°C, with 85–70 percent RH marking the transition from syrup to rubber and 50–35 percent RH marking the transition from rubber to glass (Walters 1998b). Physiological activity corresponds to changes in the viscosity of the aqueous matrix (Leprince and Hoekstra 1998; Leprince et al. 2000; Buitink et al. 1998a, 1998b, 2000), suggesting that a predominant effect of water on reaction kinetics is mediated through the resistance factor, R, in Equation 6.1. In the germplasm preservation and food stability literature, glasses have often been viewed as a static matrix, with few changes occurring once a glassy state has been achieved. “Safe” storage conditions have been assigned as some arbitrary
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temperature below the glass transition temperature, usually about 50–70°C below the glass transition (Tg; see Chapter 7, this volume). However, according to the polymer science literature, glasses are dynamic, with many forms and relaxation events. Therefore, molecular motions are only slowed, not stopped, in a glassy matrix (i.e., R in Equation 6.1 is never infinite). A more complex view of the glassy matrix in biological materials is needed to account for the discovery that viscosity in the glassy matrix actually decreases in seeds and pollen that are dried to very low water contents corresponding to hydration level I (Buitink et al. 1998a). Although increased molecular mobility under very dry conditions was predicted earlier (Labuza 1980; Clegg 1986; Vertucci and Roos 1990, 1993), its molecular basis remains conjectural. The complex interactions between temperature, water content, viscosity, and seed longevity portend the development of storage protocols that are based on physical and mechanical properties of the germplasm.
Optimum Water Contents for Storage: Orthodox Seeds Because drying reduces the kinds of reactions that occur in addition to the kinetics of those reactions, it is expected that aging is slower when seeds are stored dry. This phenomenon is generally supported for seeds stored at water contents or relative humidities corresponding to hydration level II, although exceptions have led to a suggested reclassification of seed storage physiology (Chapter 7, this volume). Some controversy arose in the 1990s regarding the longevity of seeds stored at water contents corresponding to hydration level I. One laboratory found a limit to the beneficial effects of drying on seed longevity and called the associated water content the “critical” water content (Ellis et al. 1988, 1989, 1990b). These authors argued that there was no detrimental effect of drying seeds to water contents less than this critical value, and so the critical water content marked the upper limit of a range of water contents in which longevity was maximized (the lower limit in the range being close to complete dryness). In contrast, our laboratory reported reduced longevity when seeds were progressively dried within hydration level I (Vertucci and Leopold 1987a; Vertucci and Roos 1990) and so argued for the existence of an optimum water content for storage in which longevity was maximum for a given storage temperature and aging rates increased if seeds were dried to water contents less than the optimum (Vertucci and Leopold 1987a; Vertucci and
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Roos 1990). This water content corresponded to the boundary between hydration levels I and II. The observation that overdrying seeds could be counterproductive to seed bank operators’ goals of prolonging longevity startled the seed bank community (Ellis et al. 1991; Smith 1992; Zheng et al. 1998; Walters et al. 1998). The controversy is reviewed in the International Plant Genetic Resources Institute–sponsored special issue of Seed Science Research (Walters 1998a). The presence of an optimum water content for seed storage—or, more precisely, the detrimental effect of drying—can be predicted from the hydration models discussed earlier and thought experiments on the consequences of extreme drying. According to the conceptual model of water binding, drying within the first hydration level removes water from reactive sites on macromolecular surfaces, potentially exposing macromolecules to harmful reactants and making them more susceptible to denaturation. Consistent with this idea, Vertucci and Roos (1993) hypothesized that the mechanisms of deterioration differed at supraoptimal and suboptimal water contents, implying that ∆G in Equation 6.1 was an important component driving aging reactions. Conceptual models using the resistance factor R in Equation 6.1 to explain reaction kinetics can also be invoked to explain the existence of an optimum water content. For example, removal of water molecules from the glassy matrix formed when seeds are dried increases the porosity of the matrix, allowing potentially harmful molecules to diffuse more rapidly through cells. The discovery that intracellular viscosity decreased as seed and pollen cells were dried within hydration level I provides an additional explanation for why aging rates in seeds increase with excessive drying (Buitink et al. 1998a). Although these authors did not speculate as to why molecules became more mobile with excessive drying, one can imagine that under extremely dry conditions hydrogen bonding between water molecules is diminished (i.e., less water = less bonding), thereby lessening restrictions in mobility. The strong correlation between aging rates and intracellular viscosity (Buitink et al. 1998a, 2000) argues that the resistance factor, R, has a dominant effect on aging kinetics, an expected conclusion for a reaction that is diffusion based. A relationship between viscosity and number of water molecules present further suggests that ∆G and R in Equation 6.1 are inextricably linked. Water contents that maximize storage life differ between different foods and seeds. For example, the optimum water contents for 25°C storage of yew and pea seeds is about 0.015 g H2O/g dw and 0.07 g H2O/g dw, respec-
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tively (Walters-Vertucci et al. 1996; Vertucci et al. 1994b; Figure 6.2). A large part of this variability can be explained if the lipid content of the material is taken into account: yew seeds contain about 71 percent lipid per dry mass (Walters-Vertucci et al. 1996), and pea seeds contain <2 percent lipid. Relative humidities corresponding to optimum water contents for orthodox seeds vary only slightly, from about 18 to 25 percent (0.18 ≥ aw ≥ 0.25). The similarities of optimum RH for storage between diverse seed species led to the idea that at aw between 0.18 and 0.25 the free energy differences between reactants and products of several aging reactions, |∆G| in Equation 6.1, were minimized (Vertucci and Roos 1993). Water activities describing discrete changes in the kinetics of food deterioration varied with formulation and so are not considered useful predictors of shelf life. Instead, changes in viscosity, particularly state changes from rubbers to glasses, more accurately described shelf life of foods, particularly those high in sugars (Slade and Levine 1991a, 1991b). The influence of temperature on critical or optimum water contents was an important component of the “ultradry controversy” in the 1990s (Walters 1998a). The hydration models used to explain the presence of optimum water contents for seed storage in the general sense can be used to predict the effect of changing temperature. As described previously, water sorption isotherms describe the equilibrium relationship between water content and its effective concentration (aw or w; Figure 6.2). A series of isotherms constructed at different temperatures is needed to describe the effect of temperature on water sorption characteristics (Figure 6.3A). Generally, a decrease in temperature results in an increase in water content when aw is held constant (Vertucci and Roos 1993; Walters 1998b). Classic analyses of sorption isotherms also show an increase in the number of strong binding sites or the number of water molecules associated with those binding sites with decreased temperature (Vertucci and Leopold 1987b; Buitink et al. 1996, 1998b). Intracellular viscosity is also a function of both the water content and temperature, and the water contents marking rubber-to-glass transitions or maximum viscosity increase with decreasing temperature (Slade and Levine 1991a, 1991b; Williams et al. 1993; Leopold et al. 1994; Buitink et al. 1996, 1998a; Walters 1998b; Chapter 7, this volume). Thus, using any of the three hydration models (bound water, sorption isotherms, or intracellular viscosity), we can predict that the optimum water content for seed storage increases with decreasing storage temperature (Figure 6.3B). Isotherms in Figure 6.3B show how the contro-
Figure 6.3 Water sorption isotherms of pea seeds at temperatures ranging from
–20 to 65°C. Isotherms at 5 to 35°C were measured directly, and isotherms at –20, 50, and 65°C were calculated from van’t Hoff analyses (Walters 1998b). In A, a horizontal line is drawn at 0.06 g H2O/g dw to demonstrate that the corresponding relative humidities range from 1 to 66 percent for temperatures between –20 and 65°C. In B, the vertical line is drawn at 20 percent RH, the supposed optimum water content for seed storage. Water contents corresponding to 20 percent RH range from 0.12 to 0.03 g H2O/g dw for temperatures between –20 and 65°C, suggesting that the optimum water contents for storage vary according to the storage temperature.
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versy may have arisen if temperature effects were not considered and a single critical water content was assumed. The relationship between temperature, water content, and relative humidity described by isotherms (Figure 6.3) shows that there are many ways to optimize seed water content if drying temperature, drying RH, and storage temperature are considered (Appendix 2, this volume). Experiments measuring aging rates as a function of both temperature and water contents support these predictions (Carpenter and Boucher 1992; Vertucci et al. 1994b; Buitink et al. 1998b; Walters 1998a).
The Role of Temperature in Seed Storage Revisited Generally, lowering the storage temperatures increases the longevity of stored germplasm. The benefit of reducing storage temperatures often can be predicted from temperature coefficients from various mathematical models (Dickie et al. 1990). For example, assuming that seed aging is limited by the rate of diffusion, it should follow Arrhenius kinetics with a Q10 of about 2 (Buitink et al. 1998b; Walters 1998b), suggesting that a 16-fold increase in longevity can be achieved by reducing storage temperatures from about 22°C to –18°C. Storage in liquid nitrogen vapor might result in a 20,000- to 30,000-fold increase in seed lifespans (Walters et al. 1998). The concept of ultradry storage challenges this precept by suggesting that if seeds were dried to very low water contents, aging rates could be comparable to those achieved at low storage temperatures (Ellis et al. 1990b, 1991; Zheng et al. 1998; Walters 1998a; Walters et al. 1998). The concept of ultradry storage is not supported by experimental findings, hydration models, or thought experiments on the interaction of critical (or optimum) water contents with temperature and how this interaction affects predictions of seed longevity. The following thought experiment assumes a negative exponential relationship between seed longevity and water content and a critical water content at which further drying does not increase seed longevity. The curves in Figure 6.4 examine consequences to maximum longevity if the critical water content is constant (Figure 6.4A), decreases (Figure 6.4B), or increases (Figure 6.4C) with decreasing temperature. In the simplest case, when critical water content is constant with temperature (Figure 6.4A), maximum longevity increases with decreasing temperature. Assuming no phase or state changes in water or macromolecules, maximum longevity for a temperature can be calculated
Figure 6.4 Relationships between water content, temperature, and seed
longevity in seeds as influenced by critical water content. Critical water content is constant (A), decreases (B), or increases (C) with decreasing temperature. Longevity is calculated using Harrington’s thumb rule, which assumes that shelf life doubles when water content is reduced by 0.01 g/g or temperature is reduced by 6°C. The critical water content is arbitrarily assigned at midscale for temperature (T), and effects of increasing temperature (T1) and decreasing temperature (T2) are calculated. Water content and longevity are in relative units, but the scale for each graph is the same. T, T1, and T2 are also identical in each graph.
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using existing mathematical models and temperature coefficients approximated from experiments conducted at high temperatures. This scenario shows that the lower the storage temperature, the higher the maximum longevity. Ultradry technology was developed using similar assumptions (Ellis et al. 1990b, 1991; Zheng et al. 1998). Figure 6.4a clearly shows that the promise of ultradry is not attainable: progressive drying cannot give longevities comparable to those achieved by reducing the storage temperature (Walters et al. 1998; Walters 1998a). Alternatively, if the critical water content is assumed to decrease with decreasing temperature, it would be possible to increase maximum longevity further if cold-stored seeds were dried excessively (Figure 6.4B). This scenario is not supported by experimental findings or hydration models. Finally, if the critical water content is assumed to increase with decreasing temperature, maximum longevity would also tend to increase with decreasing temperature, but the benefit of low temperature storage would be less than predicted from a constant critical water content (Figure 6.4C) and would depend on characteristics of the critical water content–temperature function. This last scenario is supported by hydration models and experimental evidence and presents the sobering idea that if there is a limit to the beneficial effect of drying, it is possible that there is also a limit to the beneficial effect of lowtemperature storage. Evidence that seeds age faster if dried below an optimum water content implies that increases in seed longevity achievable with lowtemperature storage may not be realized if seeds are dried excessively. The consequences of low-temperature storage at suboptimal water contents have not yet been experimentally determined but will probably depend on the sensitivity of the seed to overdrying and the severity of the overdrying. The severity of overdrying can be estimated using various hydration model predictions of optimum water contents. The risks of overdrying are greater according to predictions based on minimum viscosity measurements than those based on a constant w (compare optimum water contents predicted by Buitink et al. 1998a and Vertucci and Roos 1993, respectively). Based on our current understanding of the principles of dry seed storage and conservative (i.e., worst case) estimates of water content– temperature interactions, we can predict that seed longevity is indeed finite and that the costs of excessively drying seeds or excessively cooling them may be wasted because these protocols might not increase longevity sub-
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stantially. Humans are limited in their abilities to prolong seed lifespans indefinitely, and seed banks should begin to focus efforts on increasing longevity by optimizing seed quality characteristics.
Preservation of Desiccation-Sensitive Propagules: Recalcitrant and Intermediate Seeds In review, seed survival in storage is a function of the temperature and water content during storage and intrinsic properties of the seed that are under genetic and environmental control (Justice and Bass 1978). These factors affect the nature and kinetics of deteriorative reactions by altering the free energy and mobility of reactants. The interaction of factors described earlier for desiccation-tolerant, or orthodox, seeds also applies to seeds that are more sensitive to desiccation stress, or recalcitrant seeds. Intermediate seeds appear to be more tolerant of desiccation than recalcitrant seeds and less tolerant than orthodox seeds (Ellis et al. 1990a; Eira et al. 1999). Therefore, one would presume that the same principles that are described here for recalcitrant seeds are applicable to intermediate seed storage. The primary objective for preservation technology is to place cells in their lowest metabolic state without killing them. Recalcitrant seeds typically survive rapid drying to water potentials as low as –15 MPa (at the lower limits of hydration level III, Figure 6.1; Vertucci and Farrant 1995). At such low water potentials, respiration is minimal (Salmen-Espindola et al. 1994; Leprince and Hoekstra 1998; Leprince et al. 2000), but seed lifespans remain short because the metabolism is dysfunctional (Pammenter and Berjak 1999; Leprince et al. 2000; Walters et al. 2002). Quiescence can be further imposed on recalcitrant seeds by storing them at low temperatures. Orthodox seeds have an optimum moisture content corresponding to either a minimum viscosity or a specific water potential that minimizes the effects of different deteriorative reactions. Analogously, recalcitrant seeds have an optimum moisture content corresponding to either a change in viscosity (e.g., a syrup-to-rubber transition) or a water potential that minimizes the effects of dysfunctional metabolism and desiccation damage. As described earlier, water contents that correspond to changes in viscosity or specific water potentials increase with decreasing temperature (Slade and Levine 1991a, 1991b; Williams et al. 1993; Vertucci and Roos
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1993; Walters 1998b; Buitink et al. 1998a, 1998b). Consistently, the mechanism most often cited to cause desiccation damage, membrane phase behavior, is a function of the interaction between water content and temperature (Crowe et al. 1997; Hoekstra and Golovina 1999; Wolfe and Bryant 1999; Koster et al. 2000; Walters et al. 2002), and the critical water contents for desiccation damage in recalcitrant seeds increase with decreasing temperature (Kovach and Bradford 1992; Vertucci et al. 1994a, 1995; Eira et al. 1999). It follows, then, that the water content providing optimum storage for recalcitrant seeds increases with decreasing temperature. There is one problem. Water contents necessary to prevent desiccation damage at temperatures less than about –10°C are sufficiently high to allow water to freeze, with lethal consequences. The options are to increase desiccation tolerance or decrease ice formation in cells. Methods to reduce ice formation in cells can be examined in the context of Equation 6.1, where the reaction is now ice formation, and ∆G = Gice – Gliquid water. As temperature is lowered, the free energy difference (∆G) becomes more negative, and the likelihood of the reaction increases. (Note that entropy decreases [∆S < 0] when water freezes, making the second term in Equation 6.3 positive. The reaction occurs anyway because it is highly exothermic [∆H < 0]. Lowering the temperature makes the T∆S term smaller, so that the reaction becomes increasingly likely.) If solute concentration is increased by loading cells with low–molecular weight solutes, the water activity of the reactant water is reduced (becomes more negative) according to Equation 6.2, and ∆G for ice formation comes closer to 0 (i.e., the reaction is less likely). This results in a depression of the freezing temperature. Freezing point depression can reduce the likelihood of lethal ice formation only a few degrees. As temperature is lowered further, ∆G becomes more negative (Gliquid water in Equation 6.2 becomes less negative with decrease in temperature, so the value of ∆G becomes more negative), and so the freezing reaction is again promoted. Like all reactions, ice crystal formation and growth occur at a rate that is describable by Equation 6.1. The fundamental principles of cryopreservation technology are to cool cells rapidly enough that lethal ice formation does not have time to occur and then to store cells at temperatures at which the intracellular viscosity is so high that molecular motions are prevented (the R factor prohibits the reaction even though it is thermodynamically favored). At –140°C, pure liquid water forms a glass, and then there is insufficient mobility for the molecular rearrangements needed to
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form or grow ice crystals (Angell 1982). Thus, –140°C is deemed the benchmark temperature for long-term storage of hydrated samples (Bajaj 1995). Pure water must be cooled at about 20,000°C/second to achieve this theoretical vitrification (Angell 1982; Bald 1987). The necessary cooling rate in cells depends on the intracellular viscosity. For mature recalcitrant seeds, cooling rates of 200 to 1,000°C/second appear to be sufficient to prevent lethal damage when embryos are exposed to liquid nitrogen (Wesley-Smith et al. 1992, 2001). Typically, recalcitrant seeds are very large: embryonic axes can comprise 500,000 to 1 million cells. By calculating heat loads, it is clear that samples larger than about 3 mg dry mass containing more than about 1.25 g H2O/g dw cannot be cooled sufficiently rapidly to survive liquid nitrogen exposure (Wesley-Smith, pers. comm., 2002). Two alternative approaches can lead to successful cryopreservation of these materials. The first approach seeks to increase the cooling rate that can be achieved by lowering the heat capacity of the samples. This can be done by using smaller sample sizes, dehydrating samples, and allowing some extracellular ice formation in a two-step cooling process. The last procedure is commonly used in cryopreservation procedures for cell cultures and for cold-acclimated woody tissues (Sakai 1995). The second approach seeks to lower the necessary cooling rate by increasing the intracellular viscosity. This can be achieved by dehydrating samples so that endogenous protectants are concentrated (Pence 1995) and overall viscosity is reduced or by applying exogenous cryoprotectants such as glycerol, ethylene glycol, and dimethyl sulfoxide. The latter method is commonly known as a vitrification procedure (Polge et al. 1949; Whittingham et al. 1972; Bajaj 1995). The term recalcitrant is an unfortunate misnomer for desiccationsensitive seeds. The name implies that very little can be done with these seeds to preserve them. This is far from the truth (Pence 1995). Using the principles described in this chapter, successful cryopreservation has been demonstrated in numerous species, and the National Seed Storage Laboratory has initiated a pilot program to place selected species in long-term storage. Once recalcitrant germplasm has been successfully cryopreserved (i.e., survival after cryogenic exposure), questions of how long it survives in cryostorage become relevant. The principles of aging kinetics will be similar to those described previously for orthodox seeds, with the added complication of the interaction between glass properties and ice formation and
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growth. As with orthodox seeds, it is no longer possible to assume that cryogenic storage will provide infinite longevity.
Conclusions Ex situ seed banks remain the most economical and reliable method of conserving large numbers of individuals for long periods. The protocols for seed storage depend predominantly on the storage behavior of the seed (recalcitrant, orthodox, or intermediate) and the needed longevity. Extreme longevities require that moisture contents be optimized and storage temperatures be reduced. Physical principles governing deterioration kinetics can advise us on how to optimize water content and the degree to which temperature should be reduced. These principles are powerful tools that can be used to improve the efficiency of gene bank operations and to increase the number of species and tissue types that can be preserved in ex situ collections. References Acker, L. W. 1969. Water activity and enzyme activity. Food Technology 23:27–40. Angell, C. A. 1982. Supercooled water. Pages 1–77 in F. Franks (ed.), Water: A Comprehensive Treatise. Vol. 7, Water and Aqueous Solutions at Subzero Temperatures. New York: Plenum. Bajaj, Y. P. S. 1995. Biotechnology in Agriculture and Forestry 32: Cryopreservation of Plant Germplasm 1. New York: Springer-Verlag. Bald, W. B. 1987. Quantitative Cryofixation. Bristol: Adam Hilger. Buitink, J., M. M. W. E. Claessens, M. A. Hemminga, and F. A. Hoekstra. 1998a. Influence of water content and temperature on molecular mobility and intracellular glasses in seeds and pollen. Plant Physiology 118:531–541. Buitink, J., O. Leprince, M. A. Hemminga, and F. A. Hoekstra. 2000. Molecular mobility in the cytoplasm: an approach to describe and predict lifespan of dry germplasm. Proceedings of the National Academy of Sciences, USA 97:2385–2390. Buitink, J., C. Walters, F. A. Hoekstra, and J. Crane. 1998b. Storage behavior of Typha latifolia L. pollen at low water contents: interpretation on the basis of water activity and glass concepts. Physiologia Plantarum 103:145–153. Buitink, J., C. Walters-Vertucci, F. A. Hoekstra, and O. Leprince. 1996. Calorimetric properties of dehydrating pollen: analysis of a desiccation tolerant and an intolerant species. Plant Physiology 111:235–242. Bull, H. B. 1944. Adsorption of water vapor by proteins. Journal of the American Chemical Society 66:1499–1507. Carpenter, W. J., and J. F. Boucher. 1992. Temperature requirements for the storage and germination of Delphinum x cultorum seed. HortScience 27:989–992. Chan, H. W.-S. 1987. Autoxidation of Unsaturated Lipids. London: Academic Press. Clegg, J. S. 1986. The physical properties and metabolic status of Artemia cysts at low
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water contents: the “water replacement hypothesis.” Pages 169–187 in A. C. Leopold (ed.), Membranes, Metabolism and Dry Organisms. Ithaca, NY: Comstock. Close, T. J., R. D. Fendon, A. Yang, R. Asghar, D. A. DeMason, D. E. Crone, N. C. Meyer, and F. Moonan. 1993. Dehydrin: the protein. Pages 104–118 in T. J. Close and E. A. Bray (eds.), Current Topics in Plant Physiology v10: Plant Responses to Cellular Dehydration during Environmental Stress. Rockville, MD: American Society of Plant Physiologists. Crowe, J. H., L. M. Crowe, J. F. Carpenter, S. J. Prestrelski, F. A. Hoekstra, P. S. de Araujo, and A. D. Panek. 1997. Anhydrobiosis: cellular adaptations to extreme dehydration. Pages 1445–1477 in W. H. Dantzler (ed.), Handbook of Physiology, Sec. 13, Comparative Physiology, Vol. 2. Oxford, England: Oxford University Press. D’Arcy, R. L., and I. C. Watt. 1970. Analysis of sorption isotherms of non-homogeneous sorbents. Transactions of Faraday Society 66:1236–1245. Dickie, J. B., R. H. Ellis, H. L. Kraak, K. Ryder, and P. B. Tompsett. 1990. Temperature and seed longevity. Annals of Botany 65:197–204. Dure, L. III. 1993. Structural motifs in LEA proteins. Pages 91–103 in T. J. Close and E. A. Bray (eds.), Current Topics in Plant Physiology v10: Plant Responses to Cellular Dehydration during Environmental Stress. Rockville, MD: American Society of Plant Physiologists. Eira, M. T. S., C. Walters, L. S. Caldas, L. C. Fazuoli, J. B. Sampaio, and M. C. L. L. Dias. 1999. Tolerance of Coffea spp. seeds to desiccation and low temperature. Revista Brasileira de Fisiologia Vegetal 11:97–105. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1988. A low-moisture-content limit to logarithmic relations between seed moisture content and longevity. Annals of Botany 61:405–408. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1989. A comparison of the low-moisturecontent limit to the logarithmic relation between seed moisture and longevity in twelve species. Annals of Botany 63:601–611. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1990a. An intermediate category of seed storage behaviour. I. Coffee. Journal of Experimental Botany 41:1167–1174. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1991. Seed moisture content, storage, viability and vigour [correspondence]. Seed Science Research 1:275–279. Ellis, R. H., T. D. Hong, E. H. Roberts, and K. L. Tao. 1990b. Low-moisture-content limits to relations between seed longevity and moisture. Annals of Botany 65:493–504. Farrant, J. M., and C. Walters. 1998. Ultrastructural and biophysical changes in developing embryos of Aesculus hippocastanum in relation to the acquisition of tolerance to drying. Physiologia Plantarum 104:513–524. Foyer, C. H., M. Lelandais, and K. J. Kunert. 1994. Photooxidative stress in plants. Physiologia Plantarum 92:696–717. Hay, F. R., R. J. Probert, and R. D. Smith. 1997. The effect of maturity on the moisture relations of seed longevity in foxglove (Digitalis purpurea L.). Seed Science Research 7:341–349. Hendrey, G. A. F. 1993. Oxygen, free radical processes and seed longevity. Seed Science Research 3:141–153. Hoekstra, F. A., and E. A. Golovina. 1999. Membrane behavior during dehydration: implications for desiccation tolerance. Russian Journal of Plant Physiology 46:295–306.
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Justice, O. L., and L. N. Bass. 1978. Principles and Practices of Seed Storage. Agriculture Handbook no. 506. Washington, DC: U.S. Government Printing Office. Karel, M. 1980. Lipid oxidation, secondary reactions and water activity of foods. Pages 191–206 in M. G. Simic and M. Karel (eds.), Autoxidation in Food and Biological Systems. New York: Plenum. Koster, K. L., Y. P. Lei, M. Anderson, S. Martin, and G. Bryant. 2000. Effects of vitrified and non-vitrified sugars on phosphatidylcholine fluid-to-gel phase transitions. Biophysical Journal 78:1932–1946. Kovach, D. A., and K. J. Bradford. 1992. Imbibitional damage and desiccation tolerance of wild rice (Zizania palustris) seeds. Journal of Experimental Botany 43:747–757. Labuza, T. P. 1980. The effect of water activity on reactions kinetics of food deterioration. Food Technology 34:36–59. Leopold, A. C., W. Q. Sun, and I. Bernal-Lugo. 1994. The glassy state in seeds: analysis and function. Seed Science Research 4:267–274. Leprince, O., F. J. M. Harren, J. Buitink, M. Alberda, and F. A. Hoekstra. 2000. Metabolic dysfunction and unabated respiration precede the loss of membrane integrity during dehydration of germinating radicles. Plant Physiology 122:597–608. Leprince, O., and F. A. Hoekstra. 1998. The responses of cytochrome redox state and energy metabolism to dehydration support a role for cytoplasmic viscosity in desiccation tolerance. Plant Physiology 118:1253–1264. McKersie, B. D. 1991. The role of oxygen free radicals in mediating freezing and desiccation stress in plants. Pages 107–118 in E. Pell and K. Staffen (eds.), Current Topics in Plant Physiology, Vol. 8: Active Oxygen and Oxidative Stress in Plant Metabolism. Rockville, MD: American Society of Plant Physiologists. Pammenter, N. W., and P. Berjak. 1999. A review of recalcitrant seed physiology in relation to desiccation tolerance mechanisms. Seed Science Research 9:13–37. Pence, V. C. 1995. Cryopreservation of recalcitrant seeds. Pages 29–50 in Y. P. S. Bajaj (ed.), Biotechnology in Agriculture and Forestry 32: Cryopreservation of Plant Germplasm 1. New York: Springer-Verlag. Phillips, J. R., M. J. Oliver, and D. Bartels. 2002. Molecular genetics of desiccation tolerant systems. Pages 319–342 in M. Black and H. W. Pritchard (eds.), Desiccation and Survival in Plants: Drying without Dying. Wallingford, UK: CABI Publishing. Polge, C., A. U. Smith, and A. S. Parkes. 1949. Revival of spermatozoa after vitrification and dehydration at low temperature. Nature 164:666. Priestley, D. A. 1986. Seed Aging. Ithaca, NY: Comstock Publishing Associates. Pritchard, H. W., and K. R. Manger. 1998. A calorimetric perspective on desiccation stress during preservation procedures with recalcitrant seeds of Quercus robur L. CryoLetters 19(supplement 1):23–30. Rasyad, A., S. S. VanSanford, and D. M. TeKrony. 1990. Changes in seed viability and vigor during wheat seed maturation. Seed Science and Technology 18:259–267. Roberts, E. H. 1973. Predicting the storage life of seeds. Seed Science and Technology 1:499–514. Roberts, E. H., and R. H. Ellis. 1989. Water and seed survival. Annals of Botany 63:39–52. Rupley, J. A., E. Gratton, and G. Careri. 1983. Water and globular proteins. Trends in Biochemical Science 8:18–22. Sakai, A. 1995. Cryopreservation of germplasm of woody plants. Pages 53–69 in Y. P. S.
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Bajaj (ed.), Biotechnology in Agriculture and Forestry 32: Cryopreservation of Plant Germplasm 1. New York: Springer-Verlag. Salmen-Espindola, L., M. Noin, F. Corbineau, and D. Côme. 1994. Cellular and metabolic damage induced by desiccation in recalcitrant Araucaria angustifolia embryos. Seed Science Research 4:193–201. Slade, L., and H. Levine. 1991a. Beyond water activity: recent advances based on an alternative approach to the assessment of food quality and safety. Critical Reviews in Food Science and Nutrition 30:115–360. Slade, L., and H. Levine. 1991b. A food polymer science approach to structureproperty relationships in aqueous food systems: nonequilibrium behaviour of carbohydrate-water systems. Pages 29–101 in H. Levine and L. Slade (eds.), Water Relationships in Food. New York: Plenum. Smirnoff, N. 1993. The role of active oxygen in the response of plants to water deficit and desiccation. Tansley Review no. 52. New Phytologist 125:27–58. Smith, M. T., and P. Berjak. 1995. Deteriorative changes associated with the loss of viability of stored desiccation-tolerant and desiccation-sensitive seeds. Pages 701–746 in J. Kigel and G. Galili (eds.), Seed Development and Germination. New York: Marcel Dekker. Smith, R. D. 1992. Seed storage, temperature and relative humidity [Correspondence]. Seed Science Research 2:113–116. Tarquis, A. M., and K. J. Bradford. 1992. Prehydration and priming treatments that advance germination also increase the rate of deterioration of lettuce seeds. Journal of Experimental Botany 43:307–317. Thomashow, M. F., and J. Browse. 1999. Plant cold tolerance. Pages 61–80 in K. Shinozaki and K. Yamaguchi-Shinozaki (eds.), Molecular Responses to Cold, Drought, Heat and Salt Stress in Higher Plants. Austin, TX: RG Landes Company. Vertucci, C. W., J. Crane, R. A. Porter, and E. A. Oelke. 1995. Survival of Zizania embryos in relation to water content, temperature and maturity status. Seed Science Research 5:31–40. Vertucci, C. W., and J. Farrant. 1995. Acquisition and loss of desiccation tolerance. Pages 237–272 in J. Kigel and G. Galili (eds.), Seed Development and Germination. New York: Marcel Dekker. Vertucci, C. W., and A. C. Leopold. 1987a. Relationship between water binding and desiccation tolerance in tissues. Plant Physiology 85:232–238. Vertucci, C. W., and A. C. Leopold. 1987b. Water binding in legume seeds. Plant Physiology 85:224–231. Vertucci, C. W., and E. E. Roos. 1990. Theoretical basis of protocols for seed storage. Plant Physiology 94:1019–1023. Vertucci, C. W., and E. E. Roos. 1993. Theoretical basis of protocols for seed storage II. The influence of temperature on optimal moisture levels. Seed Science Research 3:201–213. Vertucci, C. W., and J. Crane, R. A. Porter, and E.A. Delke. 1994a. Physical properties of water in Zizania embryos in relation to maturity status, water content and temperature. Seed Science Research 4:211–224. Vertucci, C. W., J. Crane, R. A. Porter, and E. A. Delke. 1994b. Theoretical basis of protocols for seed storage III. Optimum moisture contents for pea seeds stored at different temperatures. Annals of Botany 74:531–540. Walters, C. (ed.). 1998a. Ultra-dry seed storage. Seed Science Research 8 (supplement).
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Walters, C. 1998b. Understanding the mechanism and kinetics of seed aging. Seed Science Research 8:223–244. Walters, C., J. M. Farrant, N. W. Pammenter, and P. Berjak. 2002. Desiccation stress and damage. Pages 263–292 in M. Black and H. W. Pritchard (eds.), Desiccation and Survival in Plants: Drying without Dying. Wallingford, UK: CABI Publishing. Walters, C., N. W. Pammenter, P. Berjak, and J. Crane. 2001. Desiccation damage, accelerated aging and metabolism in desiccation tolerant and sensitive seeds. Seed Science Research 11:135–138. Walters, C., J. L. Ried, and M. K. Walker-Simmons. 1997. Heat-soluble proteins extracted from wheat embryos have tightly bound sugars and unusual hydration properties. Seed Science Research 7:125–134. Walters, C., E. E. Roos, D. H. Touchell, P. C. Stanwood, L. E. Towill, L. Wiesner, and S. A. Eberhart. 1998. Refrigeration can save seeds economically. Nature 395:758. Walters-Vertucci, C., J. Crane, and N. C. Vance. 1996. Physiological aspects of Taxus brevifolia seeds in relation to seed storage characteristics. Physiologia Plantarum 98:1–12. Wesley-Smith, J., C. W. Vertucci, P. Berjak, N. W. Pammenter, and J. Crane. 1992. Cryopreservation of desiccation-sensitive axes of Camellia sinensis in relation to dehydration, freezing rate and the thermal properties of tissue water. Journal of Plant Physiology 140:596–604. Wesley-Smith, J., C. Walters, N. W. Pammenter, and P. Berjak. 2001. Interactions of water content, rapid (non-equilibrium) cooling to –196°C and survival of embryonic axes of Aesculus hippocastanum seeds. Cryobiology 42:196–206. Whittingham, D. G., S. P. Leibo, and P. Mazur. 1972. Survival of mouse embryos frozen to –196 and –269°C. Science 178:411–414. Williams, R. J., A. G. Hirsh, T. A. Takahashi, and H. T. Meryman. 1993. What is vitrification and how can it extend life? Japanese Journal of Freezing and Drying 39:1–10. Wolfe, J. 1987. Lateral stresses in membranes at low water potential. Australian Journal of Plant Physiology 14:311–318. Wolfe, J., and G. Bryant. 1999. Freezing, drying and/or vitrification of membranesolute-water systems. Cryobiology 39:103–129. Zheng, G. H., X. M. Jing, and K. L. Tao. 1998. Ultradry seed storage cuts costs in gene bank. Nature 393:223–224.
Chapter 7
Classification of Seed Storage Types for Ex Situ Conservation in Relation to Temperature and Moisture Hugh W. Pritchard
Seed banks are an effective form of ex situ conservation. The technology (drying and cooling) can be applied to a wide range of species, in a simple way and with little intervention, to maintain large amounts of intraspecific diversity. The popularity and effectiveness of seed banking are such that of the 6 million accessions of plant genetic resources held globally, about 90 percent are held in seed banks (Linington and Pritchard 2001). Most accessions are cultivars and landraces, with far fewer nondomesticated species conserved (perhaps 15 percent of accessions). Given concerns about the loss of plant diversity globally, there is clearly an urgent need to advance programs on the ex situ conservation of wild plant diversity to complement in situ conservation efforts. There are encouraging signs that effort is being channeled in this way. For example, a number of local (e.g., Berry Botanic Garden), regional (e.g., Department of Conservation and Land Management, Western Australia), national (National Seed Storage Laboratory, United States), and global (Millennium Seed Bank Project, United Kingdom) seed banks have been established. However, not all species are amenable to routine seed banking, and as the number and diversity of seeds entering storage facilities increase dramatically, there is a need to understand better the range of storage behaviors they can exhibit. For many years seed storage behavior was partitioned into two main classes based on the seeds’ response to dehydration: desiccation-tolerant and long-lived (orthodox) and desiccation-sensitive and short-lived (recalcitrant; Roberts 1973). A third class of seed response, intermediate, was proposed for seeds that tolerated the removal of water to a 139
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moisture content equivalent to about 20–50 percent RH (a water potential of –100 to –250 MPa), but further dehydration was detrimental to longevity (Ellis et al. 1990, 1991b). Storage response is also a consequence of storage temperature and to a lesser extent light and the gaseous environment. Low oxygen tensions improve longevity of dry orthodox seeds (Roberts and Ellis 1989) and reduce longevity in recalcitrant seeds (Tompsett 1983) and hydrated orthodox seeds (Roberts and Ellis 1989). The recommended temperature for the long-term storage of orthodox seeds is close to –18°C (Genebank Standards 1994), although cryopreservation (usually less than –100°C) can also be used for such seeds (Stanwood 1985; Vertucci 1989a, 1989b; Pritchard 1995). Intermediate seeds exhibit reduced longevity when dry stored at cold temperatures of –20°C and 0°C, and as a consequence storage temperatures of 15°C have been recommended (Ellis et al. 1990, 1991a, 1991b). Similarly, hydrated recalcitrant seeds cannot be kept at subzero temperatures because of the risk of ice nucleation and crystal growth. Consequently, it has been proposed that the minimum temperature for germination in these seeds might be used as an indirect marker of maximizing longevity in the wet state (Pritchard et al. 1995b; Pritchard and Tompsett 1998). Alternatively, cryopreservation can be used for recalcitrant seed embryonic axes after rapid, partial drying as long as cooling and rewarming rates are rapid (Pence 1995; Dumet et al. 1997). In effect, seed conservationists are challenged with coping with a diverse range of seeds with varying sensitivities to final moisture content, drying rate, cooling and warming rates, and final temperature for storage. This may mean theoretically that there is an infinite number of optimum conditions for seed processing and storage. However, gene banking resources are limited, and simpler practical recommendations are needed that are likely to enhance (if not optimize in all cases) the storage opportunities for the germplasm. The aim here is to review some of the underlying principles of biological specimen lifespan to see whether clear recommendations on seed storage temperature can be made. The concept is introduced of a practical system of seed types delineated by the sorption properties of water in seeds and the seeds’ potential storage at different subzero temperatures in relation to the biophysical properties of the water contained therein. In addition, the notion of optimum temperatures for the moist storage of desiccation-sensitive seeds is advanced in relation to their germination characteristics.
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Seed Types in Relation to Moisture Content In terms of the water relations of survival, orthodox seeds tolerate removal of both “free” water and most “bound” water (Vertucci and Farrant 1995). As the latter is reduced, longevity is increased, at least down to moisture contents in equilibrium with around 15–20 percent RH (Roberts and Ellis 1989). Other seeds clearly lose viability as bound water (less than about 85 percent RH) is being removed, such as some species of Araucaria (Araucariaceae; Tompsett 1984a) and Coffea (Rubiaceae; Dussert et al. 1999), and they tend to be short-lived in the dry state (Bonner 1990; Ellis et al. 1990). In contrast, recalcitrant seeds may tolerate the loss of free water to around 20 percent moisture content (85–90 percent RH), but usually only after rapid desiccation (Pence 1995; Vertucci and Farrant 1995). Thus, at the species level desiccation tolerance and longevity tend to partition into three types: desiccation-tolerant, inherently long-lived seeds in the dry state; short-lived, partially desiccation-tolerant seeds; and essentially desiccation-sensitive seeds. Detailed studies since the 1990s have revealed that the level of desiccation tolerance within a species is a consequence of a number of factors that affect a specific seedlot, especially in relation to the developmental age of the seeds and the method of handling (see chapters in Black and Pritchard 2002). Clearly, then, it is possible that seedlots of the same species harvested, for example, in different years could respond to desiccation (and other storage treatments) in a subtly different way (Pammenter and Berjak 1999). In addition, the apparent level of desiccation tolerance appears to vary within a seedlot. The response of an individual seed to desiccation is allor-nothing in the sense that it either does or does not subsequently germinate (i.e., it passes or fails the viability test). However, the measure of desiccation tolerance is applied at the population level. Most commonly this is assessed via co-plots of germination (on a probit or probability scale) against moisture content (Figure 7.1). Comparison of such co-plots for recalcitrant seeds of three temperate species, using whole seed and embryo–embryonic axis moisture content (MC) values down to about 20 percent, reveals that the dependency of survival on moisture contents is approximately 0.2 probits MC–1 (Pritchard and Prendergast 1986; Pritchard 1991; Tompsett and Pritchard 1998). Interpolation of data for six Coffea species with desiccation sensitivities down to around 10 percent moisture
Figure 7.1 Classification of seed types in relation to moisture content (fresh mass basis) and storage temperature. A, Three seed types (I, II, and III) are marked on the upper axis and relate to the water sorption properties in seeds (dash-dot line) and known storage responses (i.e., short-term [III] to long-term [I]; see Table 7.1). Note that the sorption isotherm in zone II would be at about 2 percent lower and higher moisture contents for oily and nonoily seeds and that RH is shown as the dependent variable for presentation reasons. The slope of the relationship between moisture content and probability of seed viability (thin lines) is typical of desiccation sensitivity studies on a broad range of species (see text for full explanation), with the relative position of each line reflecting the
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content (Dussert et al. 1999) yield a similar value. This means that for any 10 percent fall in moisture content, 2 probits of viability would be lost. This is equivalent to viability falling from 84 to 16 percent, that is, the sort of range of viability often measured in desiccation sensitivity experiments. Such variability in response to drying within the population could result from two factors: seed-to-seed variation in desiccation tolerance per se (e.g., as a feature of differential developmental age) and the consequences of a normal distribution in seed moisture contents within the population masking a single or narrow range of critical moisture contents. Unraveling the contribution of each factor remains a challenge (Black and Pritchard 2002). A practical consequence of such variation in response to desiccation is that a certain level of survival may be retained at any given moisture content within broad ranges, as shown in Figure 7.1A. Orthodox seeds tolerate drying to moisture contents at the water sorption zone I (high-affinity sorption sites) and II (weak-affinity sorption sites) interface, around 15–20 percent RH. Partially desiccation-tolerant seeds survive (to varying degrees) removal of water in sorption zone II, down to around 50–60 percent RH. Finally, under optimal drying conditions (i.e., rapid, enforced drying of isolated axes), recalcitrant seeds partially tolerate removal of sorption zone III (multimolecular, “free” water), down to approximately 85 percent RH. It is suggested that these three broad groups of species’ seed responses be described as types I, II, and III, respectively. The sequence of steps in the process of seed banking is drying, packaging, and transferring to low temperature for storage (Linington and
variability in response observed either between species or within species, the latter being dependent on seed developmental age, size/mass processing, and storage treatments. The dashed lines signify the moisture contents at which a typical representative of each type would have 50 percent viability during drying. Their corresponding RHs can be read off the graph as circa 10 percent, 50–60 percent, and 90 percent for types I, II, and III, respectively. B, Typified phase diagram for seeds showing the relationship between the glass transition temperature (Tg) and moisture content. Proposed safe temperatures for storage (Tg – 70) are shown (dashed line), as is the method of handling drying followed by cooling for representative seeds of types I, II, and III, in relation to the responses presented in A. In B, the best predicted response of type III seeds to drying and cooling is shown, which is usually achieved with isolated embryonic axes.
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Pritchard 2001). Given that seedlot desiccation tolerance levels vary both within and between species, is it necessarily the case that all seeds should be transferred to the same temperature for potential long-term storage?
Seed Types, Moisture, and Storage Temperature Table 7.1 outlines a proposed practical classification system for seed storage types in relation to phenology, temperature conditions for storage, and prestorage moisture status.
Type I Seeds The recognized international standard temperature for long-term orthodox seed storage is –18°C after drying to 3–7 percent moisture content (Genebank Standards 1994). Hundreds of thousands of accessions (seedlots) are held worldwide in this way (Linington and Pritchard 2001). Concerns have been expressed about the potential for overdrying seeds before storage based on thermodynamic considerations and molecular mobility, such that the optimum moisture content for storage increases as temperature decreases (Vertucci and Roos 1993; Walters 1998; Buitink 2000). Evidence of reduced longevity of ultradry seeds at warmer temperatures has been provided for some seeds and pollen (Vertucci 1998; Buitink 2000), and small reductions in longevity have been observed for very dry orchid seeds held at 40°C (Pritchard et al. 1999a). In contrast, the seeds of many species show little effect of ultradrying with respect to subsequent longevity at warm temperatures (see Ellis et al. 1995 and references therein). Moreover, ultradry (about 3.1 percent moisture) barley and oat seeds survived 110 years at 10–15°C (Steiner and Ruckenbauer 1995). The continuing debate around optimum conditions for storage revolves around the biophysical state of the seeds under combinations of temperature and moisture content, the so-called phase or state diagram. A representative diagram is shown in Figure 7.1B, based on axes of three legume species (modified from Sun 1997). Seeds dried to the interface of sorption zone I and II (i.e., at 3–5 percent moisture content depending on chemical composition) exhibit a glass transition, Tg, at around 45°C. This means that the seed constituents form an amorphous solid with high viscosity, which reduces the kinetics of chemical reactions, enhancing seed longevity (Walters 1998). Because the glassy state is a nonequilibrium state with an
table 7.1 A proposed practical classification system for seed storage types in relation to phenology, temperature conditions for storage, and prestorage moisture status. Seed Storage Types Includes Category
Type I
Fully orthodox seeds
Type II
Essentially orthodox seeds, including orthodox with limited desiccation ability (OLDA1), suborthodox2, and intermediate 3
Type III
Recalcitrant
Phenological Condition
Isotherm Limits
I–II sorption zone Undergo maturation drying after interface mass maturity or have a lengthy ripening period after mass maturity. Sorption zone II As type I, but heterogeneity of development yields mixed population of seeds with varying desiccation tolerance and temperature sensitivity. Often of tropical origin, exacerbating storage at conventional seed bank temperatures because of chilling-related injury. II–III sorption zone Abscised before mass maturity is interface achieved and development ex planta through desiccationmaturation is truncated by desiccation sensitivity.
Sources: 1Tompsett and Kemp (1996); 2Bonner (1990); 3Ellis et al. (1990, 1991b).
Recommended Conditions for Storage and Handling
Prestorage desiccation to ca. 15– 20% RH, storage at 20C (i.e., Tg 70). Store seeds at 70C (i.e., Tg 70) or through cryopreservation after predrying to ca. 50% RH at room temperature. May require high temperatures during rehydration or other modifications to handling procedures to maximize survival or germination after drying and storage. Store seeds (at harvest moisture content but after surface drying) close to the minimum temperature for germination; cryopreserve (i.e., Tg 70) partially dried isolated axes.
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effectively infinite response time, its determination is strongly influenced by rate (experimental) effects. Nonetheless, similar state diagrams have been determined for bean axes (Leprince and Walters-Vertucci 1995) and Typha latifolia pollen (Buitink et al. 1996). However, it is not possible to be unequivocal about the state diagram in seeds because Tg for seeds at the same moisture content can vary by about 25°C (Sun 1997). Such variation may be a consequence of differing seed chemical composition and methods of determination (e.g., equipment) or identification of the Tg in relation to the onset, middle, or end of the process. Nonetheless, considering seed aging in relation to the glassy state has been shown to be a valuable innovation in seed science in the last decade (Williams and Leopold 1989; Sun and Leopold 1994; Leprince and Walters-Vertucci 1995; Sun 1997; Buitink et al. 1998; Walters 1998; Buitink 2000). Accepting these limitations, where in relation to Tg should type I seeds be stored? Ultradry seeds (1.2–6.1 percent moisture) of 17 accessions of crucifers survived long-term storage (24–25 years) at –5°C to –10°C with little or no loss of germinability (Ellis et al. 1993). Such dry seeds are predicted to have a Tg of around 60°C, suggesting that storage at about Tg – 70 (in this instance, about –10°C) is acceptable. Seeds in gene banks, for which there is evidence of considerable longevity (Hong et al. 1998), generally have moisture contents in the range 3–7 percent. If the average Tg of such seeds were close to 50°C, then storage at Tg – 70 would be around –20°C (Figure 7.1B), close to the current international standard for storing seeds (Genebank Standards 1994). This recommended temperature standard for storage (i.e., Tg – 70) is somewhat lower than that suggested to be optimal of Typha latifolia pollen (i.e., around Tg – 40), where there is an upward shift in heat capacity of the glass (Buitink et al. 1996; Walters 1998; Buitink 2000). This rationale for seed storage does not mean that type I seeds cannot be stored at lower temperatures. Indeed, there is good practical evidence that type I seeds can be stored in liquid nitrogen chambers, which is roughly Tg – 200. One argument for using such low temperatures is based on the extrapolation of the temperature effect on longevity from 90°C to –13°C and below. Over this temperature range, desiccation-tolerant type I seed longevity in the dry state increases in a predictable way as temperature decreases (Dickie et al. 1990). This analysis, which is based on empirical data sets for eight diverse species, also suggests that the temperature response is universal (i.e., the relative effect is the same). It is possible to
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describe this response mathematically using two functions: quadratic and exponential. Both are equally appropriate, statistically, but the former is favored due to precedent (Ellis and Roberts 1980). In practice, though, the quadratic term is inappropriate when extrapolations are made to ultralow temperatures, below about –70°C. For example, seed longevity at about –150°C is predicted to be short (e.g., days, depending on moisture content), yet type I (orthodox) seeds survive storage in liquid nitrogen at –150 to –196°C (Stanwood 1985; Pritchard 1995). The theoretical benefit to longevity of using such low temperatures compared to about –20°C is considerable (Dickie et al. 1990; Pritchard 1995), and analysis of longterm storage data is needed. There are some inherent risks associated with the preservation of orthodox seeds at ultralow temperatures, some of which may relate to the fragility of systems cooled to temperatures below about Tg – 100. Hard-seeded legumes and some oilseeds, in particular, can be susceptible to physical stresses unless cooling and warming rates to and from liquid nitrogen are below about 10°C min–1 (Sakai and Noshiro 1975; Stanwood 1985; Pritchard et al. 1988; Vertucci 1989a; Pritchard 1995), possibly as a result of lipid glass cracking (Vertucci 1989b). Nonetheless, studies on a diverse range of species from various floras, such as Ohio State, United States (Pence 1991), Western Australia (Touchell and Dixon 1993), and Spain (Gonzalez-Benito et al. 1999), indicate that the application of this technology to type I (orthodox) seeds can be straightforward. The longevity of dry seed at –18°C in conventional seed banks is predicted to be in the region of hundreds of years for many species (Medeiros et al. 1998). Therefore, it is probably necessary only to consider the largescale use of liquid nitrogen in special circumstances, such as the preservation of inherently short-lived seeds (Pritchard 1995). Indeed, over 6 years seeds of the orchid Dactylorhiza fuchsii stored better in liquid nitrogen than at –20°C (Pritchard and Seaton 1993). For such material, the lower ease of access and generally higher operating costs associated with liquid nitrogen storage compared with conventional storage facilities may be worth the potential gains in longevity.
Type II Seeds There is a small but increasing number of species whose long-term storage appears to be less easily attainable at conventional storage temperatures,
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presumably because the optimal combination of desiccation and temperature for storage has not been elucidated. Such seeds exhibit varying levels of tolerance of the removal of sorption type II water (in the range about 20–85 percent RH) and are referred to here as type II seeds. Dehydration may initially improve longevity, but some of the seeds can be desiccation sensitive, and dry seeds may exhibit sensitivity at some subzero temperatures. Examples include some Araucaria species (Tompsett 1984a, 1984b), orchid species (Pritchard et al. 1999a), coffee (Coffea arabica, Rubiaceae; Ellis et al. 1990, 1991b; Dussert et al. 1999), oil palm (Elaeis guineensis, Arecaceae; Grout et al. 1983; Ellis et al. 1991c), and papaya (Carica papaya, Caricaceae; Ellis et al. 1991a; Wood et al. 2000a). In the case of some orchids, seeds pretreated to about 30 percent RH tolerated 1 year of storage at –30°C and –50°C less well than at –20°C, –70°C, and –196°C, with the temperature sensitivity response being evident by 1 month of storage (Pritchard et al. 1999a). Moreover, some seeds of three dryland palms showed varying levels of desiccation sensitivity to 21–30 percent RH and rapid (7 days) sensitivity to –20°C (Davies and Pritchard 1998). Such results indicate that cold temperature sensitivity in type II seed populations may have dual kinetics: part of the population responds negatively and rapidly to cooling, whereas the other seeds remain unaffected over the short or longer term. A similar interpretation can be made for the response of dry seeds of Araucaria columnaris (Araucariaceae; Tompsett 1984b) held at –75°C for more than 600 days and another gymnosperm, Agathis macrophylla (Araucariaceae), held at both –13°C and –75°C for up to 2,500 days (Dickie and Smith 1995). These types of response are not consistent with the earlier definition of intermediate seeds, which specified an accelerating rate of viability loss, over 3 to 6 months, when temperature was reduced from 15°C to 0°C to –20°C (Ellis et al. 1990, 1991b). Clearly, variation in response is possible because some seedlots of Coffea arabica have been stored at –20°C for about 3 years (Hong and Ellis 1992), suggesting that coffee seeds can exhibit type I storage behavior. Similarly, some batches of another putative intermediate seed, neem (Azadirachta indica, Meliaceae), can be dried to low moisture content (Leprince et al. 1998) and stored at –20°C (Tompsett and Kemp 1996; Sacandé 2000) as long as particular care is taken to rehydrate the seeds at an appropriately high temperature, thereby reducing the risk of imbibitional chilling injury (Sacandé 2000). A similar type of temperature dependence of the desiccation response was previously observed in seeds of Ziza-
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nia palustris (Poaceae; Kovach and Bradford 1992), resulting in a reconsideration of its storage behavior from type III (recalcitrant) toward type I. A modification in handling procedures has also been necessary for type II papaya (Carica papaya) seed. Both desiccation (Wood et al. 2000a) and cooling to some but not all subzero temperatures (Pritchard et al. 1999c) appear to rapidly induce dormancy in the seeds, which can be alleviated in the majority of seeds by applying heat shock after rehydration. Therefore, it is clear that there are a range of species whose seeds are essentially desiccation tolerant (i.e., certainly not recalcitrant) but nonetheless display a diverse range of physiological responses that might affect an interpretation of their storage behavior, and this means that maximizing their storage potential through improved handling is particularly challenging. Type II seeds dehydrated to a moisture content in equilibration with 50–60 percent RH (i.e., about 12 percent moisture), when at least some viability is likely to remain, are expected to have a Tg of around 5°C (Figure 7.1B). On the premise that a safe temperature for seed storage may be around Tg – 70 (see earlier comments for type I seed), it is possible that such seeds could be stored at around –70°C. Some practical evidence to support this is provided by the orchid seed study discussed earlier (Pritchard et al. 1999a). Maintaining such seeds at warmer temperatures (e.g., –20°C to –50°C) means that the seeds are close to the predicted glass transitions and therefore may be subject to an elevated rate of aging as a result of minor equipment fluctuation or the longer-term loss of the glassy state (devitrification). It is conceivable, then, that such effects explain the observation of an intermediate type of seed at –20°C (Ellis et al. 1990, 1991b). It is implicit in this argument that type II seeds are not sensitive to all subzero temperatures, however. Indeed, many can be cryopreserved, including the partially desiccation-tolerant (about 10 percent moisture content, about 70 percent RH) seeds of Warburgia salutaris (Canellaceae), a highly threatened medicinal plant indigenous to tropical Africa (Daws et al. 1999; Kioko et al. 1999), and slender naiad (Najas flexilis, Najadaceae; Hay and Muir 2000). In addition, seeds of papaya (Becwar et al. 1983) and neem (Berjak and Dumet 1996), at close to 6 percent moisture content, are known to tolerate cryopreservation, in the latter case for 4 months. The problems of storing type II seeds at subzero temperatures appear to be species-specific. The response is not linked necessarily to –20°C; some temperatures are more pernicious than others (Grout et al. 1983; Ellis et
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al. 1990, 1991b; Pritchard et al. 1999a). If, as assumed, the seeds at warmer temperature have exited the glassy state, then the higher the temperature above Tg, the faster the aging rate should be (Sun 1997; Walters 1998). However, this is contrary to many observations (Ellis et al. 1990, 1991b; Pritchard and Seaton 1993). Other mechanisms therefore must contribute to viability loss in this nonglassy state. Primary targets for stress are cellular membranes. Seeds of the tropical species neem (Azadirachta indica) have a high membrane transition temperature that predisposes the seeds to imbibitional chilling stress (Sacandé 2000). Subtropical papaya (Carica papaya) seeds exhibit greatest stress at about –13°C (Pritchard et al. 1999c). The corollary is that more temperate species may show some temperature sensitivity over a slightly lower range of temperatures. Although there is some evidence for this in orchid seeds (Pritchard and Seaton 1993; Pritchard et al. 1999a), such a hypothesis needs rigorous testing.
Type III Seeds Recalcitrant (type III) seeds (as opposed to parts thereof) are incapable of tolerating the removal (by conventional means) of all free water (i.e., within zone III of the water sorption isotherm). This essentially means that seed viability is at least partially, and more likely wholly, lost on drying to 15–25 percent moisture content (about 85 percent RH and –11 to –20 MPa; Pritchard 1991; Vertucci and Farrant 1995).
Long-Term Storage Recently, some advances have been made toward the development of longterm ex situ conservation methods for such material using cryopreservation of the isolated axis or embryo after rapid, partial desiccation (Pritchard and Prendergast 1986; Pence 1990, 1995; Fu et al. 1993; Dumet et al. 1997). The optimum moisture content for predrying varies somewhat between species and in relation to the developmental age of the material (Dumet et al. 1997), but it is often close to 17 percent (e.g., for Araucaria hunsteinii [Araucariaceae] embryos [Pritchard and Prendergast 1986] and Coffea liberica [Rubiaceae] embryos [Normah and Vengdasalam 1992]). To date, this approach, which includes a recovery phase in vitro, represents the best generic preservation method for the long-term storage of type III seeds (for recent reviews see Pence 1995; Dumet et al. 1997).
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Partial drying to around 20 percent moisture should decrease Tg in seed material close to –60°C (Figure 7.1B). Therefore, storing seeds at temperatures close to this runs the risk of devitrification from the glassy state and ice nucleation and growth. Work on a model system, consisting of encapsulated (in alginate beads) seeds of orchids and fungal hyphae dehydrated to such a moisture content, revealed the risks of storage at –70°C. Seeds and fungal tissue in about half the beads died within 1 day, and the rest survived for a month (Wood et al. 2000b). In contrast, more than 85 percent of beads had surviving seeds and hyphae after cryopreservation (Wood et al. 2000b). Thus, in the case of type III seeds, cryopreservation equals Tg – 70 (i.e., about –130°C). The application of cryopreservation to type III seeds, which generally involves the excision of axes, drying, cryopreservation, and in vitro growth, may be too costly and time-consuming for a majority of such species. Therefore, it seems likely that in the future the focus will remain on species of high socioeconomic value, conservation threat status, or endemism.
Short-Term Storage A water relations framework for the physiological responses in a recalcitrant seed has been established for Aesculus hippocastanum (Hippocastanaceae) seeds (Tompsett and Pritchard 1998). It appears that over a range of axis water potentials extending down to about –3 MPa, at least four physiological processes can progress. These are desiccation-maturation, the onset of desiccation intolerance, longevity, and germination. Controlling storage environments to maintain specific water potentials in this range, where the seeds are approaching the fully hydrated state, is difficult. As a consequence, recent attention has focused on influencing two of these responses, longevity and germination, via temperature control, which is more easily achievable. Reducing temperature for the wet storage of Araucaria hunsteinii seeds had opposite effects on seed germination rate and longevity, decreasing the former and increasing the latter (Pritchard et al. 1995b). Although the optimum temperature for storage was not identified, it was noted that many months of storage were achievable when the seeds were held at 6°C, which was also the estimated minimum temperature for germination. Subsequent investigations on Aesculus hippocastanum seeds showed that 3 years of storage were possible when the seeds were held about 8°C below the estimated
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minimum temperature for germination (Pritchard et al. 1996, 1999b). The objective of such studies was to improve the short-term handling opportunities for recalcitrant seed material by reducing seedlot spoilage as a result of sprouting (germination). But the studies also suggested a physiological basis to improved longevity, leading to the hypothesis that the minimum temperature for germination (radicle emergence) provides adequate to ideal conditions for their short-term hydrated storage. In parallel to these early studies, the storage longevity and germination characteristics were recorded for recalcitrant seeds of a diverse range of species, providing a combined data set for 29 species from six families on which to more stringently test this hypothesis. The data in Table 7.2 relate mainly to the hydrated storage of seeds at or close to their harvest moisture content. The seeds were maintained in inflated polyethylene bags, which were regularly ventilated, and the seeds subsequently germinated on 1 percent (weight/volume) agar-water. Exceptions to this were Aesculus hippocastanum and Quercus robur (Fagaceae) seeds, which were stored on agar-water and in peat, respectively, and the storage studies on Inga species, which involved embryos isolated from the succulent sarcotesta. The temperature conditions under which the highest level of survival has been recorded are described as “best storage.” However, it should be noted that this does not necessarily identify the optimum conditions for storage because most storage studies have not used a wide range of temperatures. The information does represent the best available data. The data sources are listed in Table 7.2. The minimum temperature for germination varied from about 1°C in Quercus robur to about 24°C in Aesculus hippocastanum. In these mainly temperate species, these temperature limits encourage germination either side of winter in the case of the former species and mainly after winter in the case of the latter species. Not surprisingly given the tropical provenance of the majority of the species listed in Table 7.2, the most frequently observed minimum temperature range for germination was 10–16°C, which applied to 17 species. Ten species had minimum temperatures for germination between 5 and 9°C. Minimum temperatures varied between species, and there appeared to be little intraspecific variation. Two or more seedlots were investigated for Aesculus hippocastanum, Araucaria hunsteinii, Inga punctata (Fabaceae), and Quercus robur, and their minimum temperatures were found to vary by only 1–4°C (Table 7.2). However, temperature minima for seed germination are not fixed at the species level but
table 7.2 Base temperature for germination rate and best storage performance for recalcitrant seeds of 29 species from six families. Species (Family)
Aesculus hippocastanum L. (Hippocastanaceae) Anisoptera costata Korthals (Dipterocarpaceae) Araucaria angustifolia Kuntze (Araucariaceae) Araucaria hunsteinii K. Schum. (Araucariaceae) Cotylobium burckii (Heim) Heim (Dipterocarpaceae) Cotylobium melanoxylon (Hook. f.) Pierre (Dipterocarpaceae) Dipterocarpus obtusifolius Teysm. Ex. Miq. (Dipterocarpaceae) Dipterocarpus turbinatus Gaertn. f. (Dipterocarpaceae) Dipterocarpus zeylanicus Thwaites (Dipterocarpaceae) Guarea thompsonii Sprague and Hutch (Meliaceae) Hopea foxworthyi Elmer (Dipterocarpaceae) Hopea odorata Roxburgh (Dipterocarpaceae) Inga marginata Willdenow (Fabaceae) Inga punctata Willdenow (Fabaceae) Inga vera Willdenow ssp. spuria (Bentham) J. Leon (Fabaceae) Parashorea smythiesii Wyatt-Smith ex Ashton (Dipterocarpaceae) Parashorea tomentella (Sym.) Meijer (Dipterocarpaceae) Quercus robur L. (Fagaceae) Shorea affinis (Thwaites) Ashton (Dipterocarpaceae) Shorea amplexicaulis Ashton (Dipterocarpaceae) Shorea argentifolia Symington (Dipterocarpaceae) Shorea contorta Vidal (Dipterocarpaceae) Shorea ferruginea Dyer ex Brandis (Dipterocarpaceae)
Base Temperature for Germinationa (C SE)
Best Storage Recorded (% viability/days/C)
Source
23.4–25.2 1.0
40/1150/16
1, 2
11.4 1.2
44/30/18
3
10.1 7.8
67/496/2
3
5.6–6.7 0.5
96/183/2
3, 4
13.0 0.6
52/28/21
3
15.4 1.0
46/67/21
3
12.4 1.4
20/60/18
3
15.2 2.9
20/177/16
3
16.2 0.6
53/100/21
3
13.0 3.0
85/44/18
3
6.4 1.0
68/365/18
3
7.1 0.6
48/93/16
3
9.1 1.0 6.7–10.9 1.0
10/35/21 75/63/16
5 5
8.5 0.2
10/98/16
5
9.7 1.6
50/317/18
3
10.5 1.2 0.8–2.4 1.0
40/91/16 52/1280/2
3 6, 7
9.4 3.1
56/253/21
3
15.8 1.4
60/71/21
3
16.4 2.6
60/45/21
3
8.8 1.0
35/32/21
3
12.1 0.7
36/77/21
3
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table 7.2 (continued) Base temperature for germination rate and best storage performance for recalcitrant seeds of 29 species from six families. Species (Family)
Shorea leprosula Miq. (Dipterocarpaceae) Shorea parvifolia Dyer (Dipterocarpaceae) Trichilia emetica Vahl (Meliaceae) Trichilia megalantha Harms (Meliaceae) Trichilia pieureana Jussieu (Meliaceae) Vatica mangachapoi (Dipterocarpaceae)
Base Temperature for Germinationa (C SE)
Best Storage Recorded (% viability/days/C)
Source
7.5 2.2
45/30/21
3
8.3 1.0 11
40/57/18 71/124/16
3 8
13.1 5.7
87/221/18
3
9.1 2.1
50/106/18
3
13.0 1.5
24/85/21
3
Sources: 1Pritchard et al. (1996); 2Pritchard et al. (1999b); 3Tompsett and Kemp (1996); 4Pritchard et al. (1995b); 5Pritchard et al. (1995a) and Pritchard, Davies, and Haye, unpublished; 6Holmes and Buszewicz (1955); 7Pritchard and Manger (1990); 8Daws et al. (1999). a Base (minimum) temperatures were determined by linear regression of germination rate data against temperature for different percentiles of the population (for explanation, see Pritchard and Manger 1990), except for Trichilia emetica, where the datum is based on a radicle extension rate of 0.1 mm/day1 compared with an optimum of 4.6 mm/day1 at 31C (Daws et al. 1999).
depend on the seeds’ environmental origins and postharvest history (Baskin and Baskin 1998). Indeed, a systematic change in temperature minimum has been observed in recalcitrant seeds of Aesculus hippocastanum on chilling-induced dormancy loss (Pritchard et al. 1999b). It is noteworthy that the only other species listed in Table 7.2 for which there is a record of seed dormancy is Quercus robur. The temperature for “best storage” varied from 2 to 21°C, which is very similar to the observed range of germination temperature minima (Table 7.2). The numbers of species that could be stored for about 1, 2, and 3 months were seven, six, and six, respectively. Seeds of three and four species survived 4–6 months and about 7–12 months, respectively. The remaining three species could be stored for 16–43 months. Twenty-six of the 29 species stored best at 16–21°C, probably delimited at the lower temperature by the onset of chilling injury. The remaining three species stored well at 2°C, even though two of these (Araucaria spp.) probably could be considered subtropical or warm temperate rather than temperate. When the minimum temperature for germination was regressed against storage temperature, there was a significant (p = 0.026) relationship (y =
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Figure 7.2 Effects of temperature on radicle emergence (A) and epicotyl emergence (B) for three batches of Inga punctata embryos (from type III seeds). The dashed and dotted lines indicate the temperatures at which 50 percent germination is interpolated. (Data are replotted from Pritchard et al. 1995a).
5.52 + 0.35x; r = 0.45; n = 24). The results suggest that the current best practice for storing recalcitrant seeds is to use temperatures from about 2°C below to 7°C above the base temperature for germination, depending on species. Although the data provide general support of the hypothesis being tested, there are obvious limitations to the model. For example, there are no data in the middle of the temperature range under consideration,
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around 10°C. Moreover, the more tropical material clearly responds better to storage above the minimum temperature for germination (radicle emergence), but why? Figure 7.2 shows the effect of temperature on the emergence of both the radicle and epicotyl of three lots of Inga punctata embryos. Compared with epicotyl emergence, a greater level of radicle emergence was possible at lower temperatures. The temperatures interpolated for 50 percent emergence were 11–13°C and 13–15°C for the radicle and epicotyl, respectively. The difference in value probably reflects differential chilling sensitivities between the tissues of the axis and may explain why there is a less than perfect match between the minimum temperature for germination (radicle emergence) and the “best storage” temperature (Table 7.2 and Figure 7.2). Further studies are needed to better quantify the temperature dependency of recalcitrant seed survival in relation to the competence of the whole embryo for growth. Another challenge is to elucidate the mechanism of chilling injury and seek ways to alleviate such stress, thereby improving the short-term hold capability of tropical recalcitrant seeds.
A Shortcut to Types Is there any quick and easy way to characterize an unknown seed into one of the three types, without carrying out potentially protracted physiological studies? Not yet. It would be convenient if desiccation sensitivity could be predicted on taxonomic grounds. Certainly, there are taxonomic hotspots for type III seeds, such as Fagaceae, Moraceae, and Sapotaceae (Hong et al. 1998; Dickie and Pritchard 2002). However, no family appears to have exclusively species with desiccation-sensitive seeds. Moreover, as discussed earlier, there are many examples of congeneric species with varying seed storage types. There are some ecological correlates that may indicate likely seed response to dehydration. For example, an analysis of seed mass for 1,080 species against storage types showed that type I, II, and III seeds have mean masses of 329, 900, and 3,958 mg, respectively. However, for each type seed masses are distributed over at least three orders of magnitude, leading to considerable overlap in masses across types (Dickie and Pritchard 2002). Of course, simple associations with single factors (ecological, biochemical, or other) may not have broad applicability, and the use of multiple criteria keys has been suggested by Hong and Ellis (1998) for Meliaceae
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species, using four characteristics (seed size, seed shape, seed moisture content at maturity, and plant ecology).
Conclusions Current evidence strongly suggests that seeds can be simply classified into three types in relation to their tolerance of equilibration to three main areas of the water sorption isotherm and storage in the glassy state. It is proposed that the three types replace the terms orthodox, intermediate, and recalcitrant. Seeds of type I can be stored at Tg – 70 (i.e., around –20°C) or under cryopreservation conditions. Similarly, type II seeds may be stored at Tg – 70 (about –70°C, at odds with current international recommendations for the storage of dry seeds) or under cryopreservation conditions. For type III (desiccation-sensitive) seeds, Tg – 70 equals cryopreservation, although such storage may be an option only for the isolated embryo or part thereof. Alternatively, type III seeds can be stored in the short term at above-zero temperatures close to their minimum temperature for the progression of germination.
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Sakai, A., and M. Noshiro. 1975. Some factors contributing to the survival of crop seeds cooled to the temperature of liquid nitrogen. Pages 317–326 in O. H. Frankel and J. Hawkes (eds.), Crop Genetic Resources for Today and Tomorrow. Cambridge, UK: Cambridge University Press. Stanwood, P. C. 1985. Cryopreservation of seed germplasm for genetic conservation. Pages 199–226 in K. K. Kartha (ed.), Cryopreservation of Plant Cells and Organs. Boca Raton, FL: CRC Press. Steiner, A. M., and P. Ruckenbauer. 1995. Germination of 110-year-old cereal and weed seeds, the Vienna sample of 1877. Verification of effective ultra-dry storage at ambient temperature. Seed Science Research 5:195–199. Sun, W. Q. 1997. Glassy state and seed storage stability: the WLF kinetics of seed viability loss at T > Tg and the plasticization effect of water on storage stability. Annals of Botany 79:291–297. Sun, W. Q., and A. C. Leopold. 1994. Glassy state and seed storage stability: a viability equation analysis. Annals of Botany 74:601–604. Tompsett, P. B. 1983. The influence of gaseous environment on the storage life of Araucaria hunsteinii seed. Annals of Botany 52:229–237. Tompsett, P. B. 1984a. Desiccation studies in relation to the storage of Araucaria seed. Annals of Applied Biology 105:581–586. Tompsett, P. B. 1984b. The effect of moisture content and temperature on the seed storage life of Araucaria columnaris. Seed Science and Technology 12:801–816. Tompsett, P. B., and R. Kemp. 1996. Database of Tropical Tree Seed Research (DABATTS). Database Contents. Kew, UK: Royal Botanic Gardens Kew. Tompsett, P. B., and H. W. Pritchard. 1998. The effect of chilling and moisture status on the germination, desiccation tolerance and longevity of Aesculus hippocastanum L. seeds. Annals of Botany 82:249–261. Touchell, D. H., and K. W. Dixon. 1993. Cryopreservation of seed of Western Australian native species. Biodiversity and Conservation 2:594–602. Vertucci, C. W. 1989a. Effects of cooling rate on seeds exposed to liquid nitrogen temperatures. Plant Physiology 90:1478–1485. Vertucci, C. W. 1989b. Relationship between thermal transitions and freezing injury in pea and soybean seeds. Plant Physiology 90:1121–1128. Vertucci, C. W., and J. M. Farrant. 1995. Acquisition and loss of desiccation tolerance. Pages 237–271 in J. Kigel and G. Galili (eds.), Seed Development and Germination. New York: Marcel Dekker. Vertucci, C. W., and E. E. Roos. 1993. Theoretical basis of protocols for seed storage II. The influence of temperature on optimal moisture levels. Seed Science Research 3:201–213. Walters, C. 1998. Understanding the mechanisms and kinetics of seed aging. Seed Science Research 8:223–244. Williams, R. J., and A. C. Leopold. 1989. The glassy state in corn embryos. Plant Physiology 89:977–981. Wood, C. B., H. W. Pritchard, and D. Amritphale. 2000a. Desiccation-induced dormancy in papaya (Carica papaya L.) is alleviated by heat shock. Seed Science Research 10:135–145. Wood, C. B., H. W. Pritchard, and A. P. Miller. 2000b. Simultaneous preservation of orchid seed and its fungal symbiont using encapsulation-dehydration is dependent on moisture content and storage temperature. CryoLetters 21:125–136.
Chapter 8
Determining Dormancy-Breaking and Germination Requirements from the Fewest Seeds Carol C. Baskin and Jerry M. Baskin
A small number of seeds greatly limits the number, kind, and size of experiments that can be conducted to determine the dormancy-breaking and germination requirements of a species. For many species, problems related to a low number of seeds can be solved simply by returning to the field and collecting additional seeds. However, in some rare species (and sometimes also in common, widely distributed ones) with low seed production, it is undesirable or impractical to collect large numbers of seeds. However, even with a small number of seeds, it is possible to learn much about the germination biology of a species. In this chapter, we show how information on seeds of other members of the family and on the life cycle (especially the phenology of seed maturation, dispersal, and germination) of the species under study may suggest the kind of dormancy present and how and when it is broken in nature. To facilitate seed germination studies, we describe how to differentiate the various general kinds of dormancy (or lack thereof). Because physiological dormancy is the most common and morphophysiological dormancy is the most difficult to break, much attention is devoted to these types of dormancy in this chapter. We have designed a move-along experiment involving a small number of seeds to determine the sequence of environmental conditions required to break dormancy in seeds with physiological or morphophysiological dormancy. We present our key for the eight known types of morphophysiological dormancy and discuss the use of data from the move-along experiment in identifying these types. 162
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Identifying Dormancy Types At the time of maturation, seeds of many species including Chrysanthemum leucanthemum L. (Baskin and Baskin 1988), Agropyron repens (L.) Beauv. (Williams 1971), and Rumex obtusifolius L. (Steinbauer and Grigsby 1960) germinate over a wide range of environmental conditions; these seeds are nondormant (sensu Baskin and Baskin 1985) or nearly so. The seeds of concern to us in this chapter do not germinate at any conditions when they are freshly matured and thus are dormant. Although it may not be too difficult to distinguish dormant from nondormant seeds, identifying the kind of seed dormancy can be difficult. One of the best clues to the kind of dormancy in seeds of a given species comes from information in the literature about other members of the family to which the species in question belongs.
Family-Level Dormancy Patterns Physical Dormancy Seeds of some species fail to germinate because the seed (or fruit) coat is impermeable to water; this is called physical dormancy. Physical dormancy occurs in members of several families, including the Anacardiaceae, Bixaceae, Cannaceae, Cistaceae, Cochlospermaceae, Convolvulaceae (including Cuscutaceae), Cucurbitaceae, Dipterocarpaceae (subfamilies Monotoideae and Pakaraimoideae but not subfamily Dipterocarpoideae), Fabaceae, Geraniaceae, Malvaceae (including Bombacaceae, Sterculiaceae, and Tiliaceae), Nelumbonaceae, Rhamnaceae, Sarcolaenaceae, and Sapindaceae (Baskin et al. 2000). However, it should be noted that in some of these families, such as the Anacardiaceae, Fabaceae, Malvaceae, and Rhamnaceae, not all taxa have physical dormancy (Baskin and Baskin 1998). For example, in the Anacardiaceae only Rhus, Cotinus, and a few of the other 70 or so genera have physical dormancy (Baskin and Baskin, unpublished data). The way to determine whether seeds or fruits are impermeable to water is to weigh them, place them on a moist substrate for 24 hours, blot them dry, and reweigh. If seeds or fruits are impermeable to water, the surest way to break dormancy is to cut a small hole in the seed or fruit coat, preferably on the cotyledon end so as not to accidentally damage the radicle. Acid scarification or heat treatments often are used when it is
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desirable to break physical dormancy in large quantities of seeds. Freshly matured seeds or fruits of some tropical members of the Anacardiaceae, Cucurbitaceae, Fabaceae, Malvaceae, and Sapindaceae are not only permeable to water but recalcitrant. That is, if water content of the seed or fruit decreases to less than about 25 percent of its air-dry weight (depending on the species), it will lose viability (Baskin and Baskin 1998). In addition to an impermeable seed or fruit coat, the embryo in seeds of some species, including Ceanothus sanguineus Pursh, Cercis spp., Rhus aromatica Ait., and Tilia spp. (see Table 6.10 in Baskin and Baskin 1998 for complete list), is physiologically dormant. Therefore, germination does not occur until the seed or fruit coat becomes permeable and dormancy of the embryo has been broken. See Baskin and Baskin (1998) for a discussion of how physical dormancy is broken in nature. The remainder of this chapter is devoted to seeds and fruits whose coats are permeable to water.
Morphological Dormancy This type of dormancy occurs in seeds with an undifferentiated embryo and in those with a differentiated but very small (underdeveloped) embryo. One or more (sometimes all) genera in the Balanophoraceae, Burmanniaceae, Ericaceae, Gentianaceae, Hydnoraceae, Lennoaceae, Monotropaceae, Orchidaceae, Orobanchaceae, Pyrolaceae, and Rafflesiaceae have either dwarf or micro seeds with small, undifferentiated embryos consisting of two or more cells, depending on the species (Baskin and Baskin 1998). In the presence of appropriate environmental stimuli, which may include exudates from roots of potential host plants (Parker and Riches 1993), cells of the embryo divide, and eventually a tissue emerges from the seed. Depending on the species, the “germinating” seed produces a tubercle, haustorium, or protocorm but not cotyledons or a radicle per se. Because germination of seeds with undifferentiated embryos often requires special media and/or stimulatory compounds (e.g., orchids and parasitic species), consultation with a specialist on the propagation of the genus or family in question increases the chance of growing the species from seeds. In at least 55 plant families, including the Apiaceae, Araceae, Araliaceae, Berberidaceae, Illiciaceae, Liliaceae, Magnoliaceae, Papaveraceae, Ranunculaceae, Taxaceae, and Winteraceae (see Table 3.3 in Baskin and Baskin 1998 for a complete list of families), seeds have a fully differentiated (cotyledons and radicle present) but underdeveloped (small) embryo. The embryo must undergo elongation or growth before germination (i.e., radi-
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cle emergence) occurs. Seeds with differentiated, underdeveloped embryos may not require any special dormancy-breaking treatment to promote germination, and embryos begin to grow as soon as seeds are placed on a moist substrate at appropriate temperature and light (or dark) conditions, depending on the species; these seeds have morphological dormancy (Baskin and Baskin 1998). After seeds are imbibed, the time required for completion of embryo growth and emergence of the radicle varies from 6 (Jacobsen and Pressman 1979) to 30–45 (Baskin and Baskin 1986a) days.
Physiological Dormancy Dormancy (lack of germination under otherwise favorable conditions) in seeds of many species is attributed to a physiological inhibiting mechanism in the embryo (Nikolaeva 1969, 1977); this is called physiological dormancy. Physiological dormancy can be found in seeds with undifferentiated embryos; differentiated, underdeveloped (small) embryos; or differentiated, fully developed embryos. Physiological dormancy is the most common type of seed dormancy, and it occurs in numerous plant families whose seeds have differentiated, fully developed embryos, including the Amaranthaceae, Asteraceae, Boraginaceae, Brassicaceae, Caryophyllaceae, Chenopodiaceae, Cyperaceae, Euphorbiaceae, Lamiaceae, Poaceae, Rosaceae, and Scrophulariaceae (Baskin and Baskin 1998). There are three levels of physiological dormancy: nondeep, intermediate, and deep (Nikolaeva 1969, 1977). Nondeep physiological dormancy is broken by 1–8 weeks of warm (≥15°C) or cold (0–10°C) stratification, depending on the species, and gibberellic acid (GA3) promotes germination (Baskin and Baskin 1998). Intermediate physiological dormancy is broken by 8–14 weeks of cold stratification, but a period of dry storage at room temperatures or warm stratification may reduce the length of the cold stratification period required to break dormancy; GA3 can promote germination. Deep physiological dormancy is broken by 10–16 weeks of cold stratification, but neither warm pretreatment nor GA3 promotes germination.
Morphophysiological Dormancy When physiological dormancy occurs in seeds with undifferentiated embryos or in those with differentiated, underdeveloped embryos, the seeds have morphophysiological dormancy. In the remainder of this chapter,
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however, morphophysiological dormancy is used to refer only to seeds with differentiated, underdeveloped embryos.
Key to the General Types of Seed Dormancy Although information about the types of dormancy found in a plant family can be very useful, germination studies of a specific species are aided by knowledge of the kind of dormancy occurring in the seeds of that species. To facilitate identification of the kind of dormancy, a key has been constructed (Figure 8.1). This key is based on the permeability of the seed or fruit coat to water; the characteristics and size of the embryo, which often may be obtained from the literature (e.g., see Martin 1946); and whether freshly matured seeds germinate within about 30–45 days at temperatures simulating those in the habitat at the time of seed maturation. It should be noted that freshly matured seeds of some species can germinate at temperatures higher or lower than those in the habitat at the time of seed maturation. Furthermore, in some species treatments that overcome physiological dormancy result in a decrease and/or increase in the temperature range for germination. A change in temperature requirements for germination means that the freshly matured seeds were in conditional dormancy. Conditional dormancy occurs in seeds with nondeep physiological dormancy, and it represents an intermediate state between dormancy and nondormancy (see “Dormancy Continuum” in Baskin and Baskin 1985).
Breaking Physiological and Morphophysiological Dormancy Physiological and morphophysiological dormancy are the types of greatest concern (i.e., they can be the most difficult to break) in propagating many species from seeds. If seeds have either fully developed or underdeveloped embryos with physiological dormancy, they may require warm and/or cold stratification treatments before they will germinate. In both kinds of treatments, seeds are placed on a moist substrate. The range of effective temperatures for warm stratification is about 15–35°C (Baskin and Baskin 1986b), with 20–25°C being optimal for many species (Nikolaeva 1969). Many seeds that require exposure to high summer temperatures before they can germinate in autumn (especially those of winter annuals) also germinate after 1–3 months of dry storage at ambient room temperatures (Baskin and Baskin 1983). The range of effective temperatures for cold
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figure 8.1. Simplified key to general kinds of dormancy (or lack thereof) in freshly-matured seeds. 1. Seed or fruit coat not permeable to water. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .2 2. Germination occurs within about 2 weeks when seed or fruit coat is scarified. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .PHYSICAL DORMANCY 2. Germination does not occur within about 2 weeks when seed or fruit coat is scarified. . . . . . . . . . . . . . . . . . . . . . .COMBINATION OF PHYSICAL AND PHYSIOLOGICAL DORMANCY 1. Seed or fruit coat permeable to water. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .3 3. Embryo not differentiated, or if differentiated it is underdeveloped (small). . . . .4 4. Embryo not differentiated. . . . . . . . . . . . . . . . . . . .SPECIALIZED TYPE OF MORPHOLOGICAL DORMANCY 4. Embryo differentiated but underdeveloped (small). . . . . . . . . . . . . . . . . . . . . .5 5. Seeds germinate within about 30 days at simulated autumn, spring, and summer habitat temperatures.a . . . . .MORPHOLOGICAL DORMANCY 5. Seeds do not germinate within about 30 days at simulated autumn, spring, and summer habitat temperatures.a . . . . . .MORPHOPHYSIOLOGICAL DORMANCY 3. Embryo differentiated and fully developed (elongated). . . . . . . . . . . . . . . . . . . . .6 6. Seeds germinate within about 30 days at simulated autumn, spring, and summer habitat temperatures.a . . . . . . . . . . . . . . . . . . . . . . .NONDORMANT 6. Seeds do not germinate within about 30 days at simulated autumn, spring, and summer habitat temperatures.a . . . . . . . . .PHYSIOLOGICAL DORMANCY aIn regions where winter temperatures are seldom or never below freezing, simulated winter habitat temperature also should be included.
stratification is about 0–10°C, with about 5°C being optimal for seeds of many species (Stokes 1965; Nikolaeva 1969). Depending on the species, some (rather slow) loss of dormancy may occur if seeds that normally come out of dormancy during a cold stratification treatment are stored dry at room temperatures (Baskin and Baskin 1998). If it is concluded or suspected that seeds have physiological dormancy of a fully developed or of an underdeveloped embryo, the next step is to determine what dormancy-breaking treatments to use. These decisions are greatly facilitated by data on the phenological life cycle of the species, especially the timing of seed maturation, dispersal, and germination, and on environmental conditions in the habitat from the time of seed maturation until germination.
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Physiological Dormancy in Seeds with Fully Developed Embryos Summer is the natural time for loss of seed dormancy in winter annuals, and germination occurs in autumn. Seeds of various winter annuals have been shown to require exposure to high summer temperatures before they will germinate at autumn temperatures in autumn (Baskin and Baskin 1986b). As seeds come out of dormancy, the maximum temperature at which germination is possible increases (Baskin and Baskin 1985). Therefore, seeds of winter annuals subjected to natural (or simulated) summer temperatures for 2–3 months germinate at natural (or simulated) autumn temperature regimes. In some species, maximum germination does not occur until seeds are exposed to temperature regimes simulating those of late autumn and early winter (e.g., 15/6°C; Baskin and Baskin 1973). If the species is a summer annual, the natural time for loss of seed dormancy is winter, and germination occurs in spring and/or summer. Seeds of various summer annuals have been shown to require exposure to cold stratification before they will germinate at spring temperatures in spring (Baskin and Baskin 1987). As seeds come out of dormancy, the minimum temperature at which germination is possible decreases (Baskin and Baskin 1985). Therefore, seeds of summer annuals subjected to 2–3 months of cold stratification germinate in spring and/or summer. If the species is a perennial whose seeds mature in spring, the dormancy-breaking and germination requirements may be like those of a winter annual (Baskin et al. 1998). That is, the seeds require high summer temperatures for loss of dormancy, and nondormant seeds germinate in autumn. On the other hand, many spring-produced seeds of perennials require a cold stratification treatment for loss of dormancy and therefore do not germinate until the subsequent spring, such as Mertensia virginica (L.) Pers. (Baskin and Baskin 1998, unpublished data). Most autumn-produced seeds of perennials also require a cold stratification treatment for dormancy loss to occur (Baskin et al. 1993a, 1993b); therefore, nondormant seeds germinate in spring and/or summer (Baskin and Baskin 1988).
Morphophysiological Dormancy in Seeds with Underdeveloped Embryos Germination does not occur in seeds with morphophysiological dormancy until physiological dormancy has been broken and the embryo has grown to some critical, species-dependent length, which may or may not equal the
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figure 8.2. Key to kinds of morphophysiological dormancy in seeds with differentiated, underdeveloped embryos. 1. Cold stratification (12–14 weeks) of freshly matured seeds results in emergence of radicle and epicotyl or only the radicle at simulated spring (e.g., 20/10°C, 15/6°C) temperatures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .2 2. After cold stratification, both radicle and epicotyl emerge. . . . . . . . . . . . . . . . . .3 3. Gibberellic acid substitutes for cold stratification in promoting germination. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .INTERMEDIATE COMPLEX 3. Gibberellic acid does not substitute for cold stratification in promoting germination. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .DEEP COMPLEX 2. After cold stratification, only the radicle emerges. Shoot (epicotyl) emerges after a period of warm stratification followed by a second period of cold stratification (i.e., shoot emerges the second spring). . . . . . . . . . .DEEP SIMPLE DOUBLE 1. Cold stratification (12–14 weeks) of freshly matured seeds does not result in emergence of radicle or epicotyl. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .4 4. Warm stratification (8–12 weeks) of freshly matured seeds results in emergence of epicotyl and radicle or only the radicle at simulated autumn (e.g., 20/10°C, 15/6°C) temperatures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .5 5. After warm stratification, radicle and epicotyl emerge at autumn temperatures. NONDEEP SIMPLE 5. After warm stratification, only the radicle emerges at autumn temperatures (epicotyl emerges in spring). . . . . . . . . . . . . . . . . . . .DEEP SIMPLE EPICOTYL 4. Warm stratification (8–12 weeks) of freshly matured seeds results in no emergence of radicle or epicotyl at autumn temperatures. . . . . . . . . . . . . . . . . . . . . . . . . . . .6 6. Embryo growth (but not emergence of radicle or epicotyl) occurs at autumn temperatures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .7 7. After embryo has grown, gibberellic acid promotes germination. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .INTERMEDIATE SIMPLE 7. After embryo has grown, gibberellic acid does not promote germination; seeds require cold stratification before they will germinate. . . . . . .DEEP SIMPLE 6. Embryo growth does not occur at autumn temperatures but does occur during a subsequent period of exposure to winter temperatures; seeds require cold stratification before they germinate (i.e., seeds require warm followed by cold stratification for germination). . . . . . . . . . . . . . . . . . . . . .NONDEEP COMPLEX
total length of the seed. Eight types of morphophysiological dormancy have been distinguished, based on the environmental conditions required for loss of physiological dormancy and growth of the embryo and on responses of seeds to GA3 (Baskin and Baskin 1998). To facilitate identification of the various kinds of morphophysiological dormancy, we have developed a key (Figure 8.2).
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Winter annuals whose seeds have morphophysiological dormancy germinate in autumn, like those of winter annuals whose seeds have only physiological dormancy (Baskin and Baskin 1990, 1994). Seeds of winter annuals have nondeep simple morphophysiological dormancy (Figure 8.2), and loss of physiological dormancy occurs while seeds are exposed to high temperatures in summer. However, loss of morphological dormancy (i.e., the embryo elongation that must precede radicle emergence) does not take place until physiological dormancy is broken and imbibed seeds are exposed to autumn temperatures. Furthermore, seeds of some species, such as Chaerophyllum tainturieri Hook., require light for embryo growth in autumn (Baskin and Baskin 1990). If seeds of C. tainturieri are in darkness in autumn, embryo growth does not occur, and seeds reenter physiological dormancy (secondary dormancy) as habitat temperatures decrease in late autumn (Baskin and Baskin 1990). In contrast, seeds of the winter annual Corydalis flavula (Raf.) DC. do not require light for embryo growth in autumn (Baskin and Baskin 1994). Therefore, after physiological dormancy is broken in summer, a high percentage of C. flavula seeds germinate even if they are buried. Not much is known about morphophysiological dormancy in seeds of summer annuals, probably because few summer annuals are known to have morphophysiological dormancy. Seeds of the summer annual Aethusa cynapium L. are dormant at the time of maturation in autumn in England (Roberts and Boddrell 1985). Therefore, because seeds of A. cynapium have underdeveloped embryos (Martin 1946), it has been concluded that the seeds have morphophysiological dormancy (Baskin and Baskin 1998). However, the type of morphophysiological dormancy in A. cynapium seeds has not been determined. Cold stratification at 4°C or warm stratification at 30°C promotes germination of A. cynapium seeds. However, coldstratified seeds germinated over a wide range of low to high temperatures (10/4°C, 20/4°C, 20/10°C, and 30/10°C), but warm-stratified ones did not germinate at low temperatures (Roberts and Boddrell 1985). The environmental conditions required for embryo growth in A. cynapium seeds are unknown. In the field, seeds of A. cynapium germinate primarily in spring, with some germination (≤10 percent) occurring in autumn (Roberts 1979). It is not known whether plants from autumn-germinating seeds of A. cynapium survive; therefore, we do not know whether this species can behave as a facultative summer annual. In some winter annuals, such as Papaver spp. (Roberts and Boddrell 1984), germination occurs mostly in
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autumn with plants behaving as winter annuals, but some seeds germinate in spring with plants behaving as summer annuals; these species are facultative winter annuals. Seeds of Papaver spp. have underdeveloped, physiologically dormant embryos and thus morphophysiological dormancy (Baskin and Baskin 1998). Numerous perennial species have seeds with morphophysiological dormancy. Depending on the species, a period of exposure to conditions suitable for warm and/or cold stratification (hereafter these periods of exposure will be called warm stratification or cold stratification) may be required to break dormancy (Figure 8.2). Decisions about which dormancy-breaking protocol to use for seeds of a given species are aided by information on seed dispersal and germination of the species in the field. For example, • If seeds are dispersed in spring and germinate in autumn, they may
have nondeep simple morphophysiological dormancy and therefore require only warm stratification for dormancy loss (e.g., Chaerophyllum tainturieri; Baskin and Baskin 1990). It should be noted that another explanation for delay of germination until autumn is that seeds have only morphological dormancy, but they have a low temperature requirement for germination (e.g., Isopyrum biternatum [Raf.] T. & G.; Baskin and Baskin 1986a). • If seeds are dispersed in autumn and germinate in spring, they may have deep complex morphophysiological dormancy and therefore require only cold stratification for dormancy break (e.g., Heracleum sphondylium L.; Stokes 1952). However, seeds might have deep simple double morphophysiological dormancy, with only the radicle emerging in spring after a period of cold stratification; shoot growth would not occur until the second spring (e.g., Trillium grandiflorum [Michx.] Salisb.; Barton 1944). • If seeds are dispersed in spring and germinate the next spring, they may have deep complex morphophysiological dormancy and therefore require only cold stratification for dormancy break (e.g., Delphinium tricorne Michx.; Baskin and Baskin 1994a). The warm period to which seeds are exposed in summer is not required to break dormancy. • If seeds that mature in summer are dispersed over a period of many months and germinate only in spring, they may have nondeep complex morphophysiological dormancy. These seeds would require
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both warm and cold stratification to break dormancy. Seeds dispersed in summer and early autumn would be warm stratified before being cold stratified in winter and therefore would germinate in spring (e.g., Osmorhiza longistylis [Torr.] DC.; Baskin and Baskin 1984). Seeds dispersed too late in autumn to be warm stratified would not germinate until the second spring, after they had been warm stratified in summer and cold stratified in the subsequent winter. Cold stratification is effective in promoting germination of seeds with nondeep complex morphophysiological dormancy only if it follows warm stratification. • If seeds mature in early autumn but do not germinate until the second spring (e.g., Panax spp.; Baskin and Baskin 1998), they may have deep simple morphophysiological dormancy. Seeds with this type of dormancy require three treatments (in sequence) before they will germinate: warm stratification in summer, a period at autumn temperatures for embryo growth, and cold stratification in winter. Seeds do not germinate in the field in the first spring after dispersal because they are dispersed too late in autumn to be exposed to a long enough period of warm stratification to complete the first phase of dormancy loss.
Move-Along Experiment Over the years, we have developed an experimental design that allows one to learn much about the germination ecology of a species, even if little or nothing is known about its life cycle (Table 8.1). Eighteen dishes of seeds ([two treatments + four controls] three replications) are used in this experiment, and seeds are placed on wet sand or soil. We prefer to use 50 seeds per dish, but the number per dish can be reduced if seed supplies are limited. This technique also is a good way to learn something about a species before a lot of time, energy, materials, and seeds is invested in large experiments. In our laboratory, seeds are exposed to 14 hours of light per day (40 µmol m–2s–1, 400–700 nm, cool-white fluorescent light). We use 30/15°C to simulate summer, 20/10°C and then 15/6°C to simulate decreasing temperatures in autumn, a constant temperature of 1 or 5°C (or sometimes 5/1°C) for winter, and 15/6°C and then 20/10°C to simulate increasing temperatures in spring. These temperature regimes generally approximate seasonal temperature changes in much of temperate eastern
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table 8.1 Design for move-along experiment to determine dormancy-breaking and germination requirements of seeds; seeds are placed on a wet substrate and given a 14-hour daily photoperiod at each temperature regime. Temperature Regime (C) Series A
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5 ↓ 15/6b ↓ 20/10 ↓ 30/15 ↓ 20/10 ↓ 15/6b ↓ 5 ↓ 15/6b ↓ 20/10 ↓ 30/15
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a Controls are seeds that remain on a wet substrate at 5C, 15/6C, 20/10C, and 30/15C for the duration of the experiment. b If number of seeds is limited, 15/6C can be omitted and time at 20/10C increased to 6 weeks.
North America (Wallis 1977), but the temperatures easily can be modified to simulate conditions in other parts of the world. For example, to simulate temperatures for the boreal region, 15/10°C or 20/10°C might be used for the highest temperature regime, but to simulate temperatures for the Mediterranean region 15/10°C might be used for the lowest temperature regime. In our studies, daily temperature regimes are 12/12 hours, and lights come on in the incubators 1 hour before the beginning of the hightemperature period and remain on for 1 hour after the beginning of the low-temperature period. Controls for the experiment are seeds incubated continuously at each temperature regime. If seeds are nondormant or if they have morphological dormancy, they will germinate at one or more of the temperature regimes. Also, seeds of some species may require a long period at a partic-
table 8.2 Germination percentages (mean percentage SE) of seeds of Zigadenus leimanthoides and Zigadenus densus moved through two series of temperature regimes. Imbibed seeds were exposed to 14 hours of light each day. Control seeds were kept continuously at 5C, 20/10C, and 30/15C. Zigadenus leimanthoides Time (weeks)
12 6 12 6
Moved
Zigadenus densus
Controls
Moved
Controls
Series A
Series B
5
20/10
30/15
Series A
Series B
5
20/10
30/15
30/15 0 ↓ 20/10 0 ↓ 5 97 2 ↓ 20/10 97 2
5 12 2 ↓ 20/10 97 2 ↓ 30/15 97 2 ↓ 20/10 97 2
5 42 ↓ 5 71 2 ↓ 5 99 1 ↓ 5 99 1
20/10 31 ↓ 20/10 13 1 ↓ 20/10 36 3 ↓ 20/10 41 3
30/15 0 ↓ 30/15 0 ↓ 30/15 0 ↓ 30/15 11
30/15 0 ↓ 20/10 0 ↓ 5 81 ↓ 20/10 97 1
5 51 ↓ 20/10 92 6 ↓ 30/15 92 6 ↓ 20/10 92 6
5 62 ↓ 5 65 7 ↓ 5 98 1 ↓ 5 99 1
20/10 0 ↓ 20/10 51 ↓ 20/10 28 9 ↓ 20/10 34 9
30/15 0 ↓ 30/15 0 ↓ 30/15 0 ↓ 30/15 21
Source: Data modified from Baskin et al. (1993b).
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ular temperature for dormancy break and germination. For example, seeds of Ceratiola ericoides begin to germinate after about 90 days at 30/15°C, and continuous incubation on a wet substrate at high temperatures is optimal for dormancy break and germination of seeds of this species (Baskin et al., unpublished data). During an experiment, seeds are moved from one temperature regime to the next in each of the two series of temperature regimes (Table 8.1). Therefore, the experiment is called a move-along experiment. Series A tells us whether warm stratification alone is sufficient for dormancy break, and Series B tells us whether cold stratification alone is sufficient for dormancy break. By moving seeds through Series A and B concurrently, it is possible to determine whether warm stratification must precede cold stratification before seeds can germinate. For example, seeds of Zigadenus leimanthoides Gray and Zigadenus densus (Desr.) Fernald require cold stratification for loss of dormancy, but warm stratification does not have to precede cold stratification (Table 8.2). Seeds kept at 5°C for the duration of the experiment eventually germinated at 5°C, but germination was faster for seeds moved from 5°C to 20/10°C than it was for those kept continuously at 5°C (Table 8.2). Embryo growth occurred while seeds were at 5°C (Baskin et al. 1993b). The information obtained from transferring seeds through Series A and B allows one to use the key for types of morphophysiological dormancy (Figure 8.2); however, additional information is needed for final decisions about some types of dormancy. If seeds germinate after cold stratification, their response to GA3 must be determined to know whether seeds have intermediate or deep complex morphophysiological dormancy. Fresh seeds (i.e., seeds that have not been cold stratified) can be placed on filter paper moistened with water or with a solution of 100 or 1,000 mg/L GA3 and distilled water and incubated at 20/10°C for 12 or more weeks (Baskin et al. 1992). To help distinguish between intermediate simple, deep simple, and nondeep complex morphophysiological dormancy, we need to know whether embryos grow in autumn or in winter. Also, if embryos grow in autumn, will the seeds germinate when treated with GA3? Embryo growth in autumn (after seeds are warm stratified in summer) but lack of germination in autumn indicates that seeds have either intermediate simple or deep simple morphophysiological dormancy, depending on their response to GA3. Seeds (with elongated embryos) can be transferred to dishes containing filter paper moistened with 1,000 mg/L GA3 (GA4/7 may work as
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well or better) to determine whether GA3 will substitute for cold stratification in promoting germination. If GA3 promotes germination, seeds have intermediate simple morphophysiological dormancy, but if GA3 does not promote germination, seeds have deep simple morphophysiological dormancy. However, it should be noted that GA3 promotes embryo growth (but not germination) in seeds with deep simple morphophysiological dormancy (Baskin and Baskin 1989). If seeds are warm stratified in summer and embryos fail to grow in autumn but do grow in winter, seeds have nondeep complex morphophysiological dormancy. If seeds are moved through Series A and B concurrently and no germination occurs, there are several things to consider: • The seeds may not be viable. A few seeds could be removed from the
dishes and examined or tested to determine whether they are viable. We recommend excising the embryo and determining its degree of firmness and color. A firm, white embryo probably is alive; a soft, slightly tan or gray one is dead. In endospermous seeds, it is useful to compare the color of the embryo with that of the endosperm. If the embryo is darker than the white endosperm, the embryo is nonviable. Visual examination of embryos can be followed by tetrazolium tests (Grabe 1970). In our experience, firm, white embryos give a positive tetrazolium test, indicating viability, but soft, gray ones give a negative test. Furthermore, it should be noted that if seeds are dead or have low vigor, they often are attacked by fungi. • Four weeks at 20/10°C may not be long enough for the embryo to become fully elongated. After 12 weeks of warm stratification at 30/15°C, seeds of Jeffersonia diphylla (L.) Pers. required 6 weeks at 20/10°C for completion of embryo growth (Baskin and Baskin 1989). • A winter temperature of 5°C may be too high for effective cold stratification to occur (Baskin et al. 1995); therefore, 1°C or 5/1°C may be required to break dormancy. • Seeds of some species may germinate to higher percentages in darkness than in light (Baskin and Baskin 1998).
We have emphasized the usefulness of the move-along experiment in determining the kind of morphophysiological dormancy; however, it can be helpful in studying seeds with fully developed but physiologically dormant embryos. For example, seeds of Floerkea proserpinacoides Willd. (Baskin et al. 1988) and Cardamine concatenata (Michx.) O. Schwarz
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(Baskin and Baskin 1994b) have fully developed embryos with physiological dormancy. Seeds of these two species need cold stratification before they will germinate. However, a period of warm stratification before the cold stratification treatment reduced the length of the cold stratification period required for 50 percent germination from 19 to 8 weeks in F. proserpinacoides and from 19 to 13 weeks in C. concatenata seeds. Thus, using Series A and B concurrently permits detection of species whose seeds have fully developed physiologically dormant embryos in which warm stratification reduces the cold stratification requirement for germination. If seed supplies are limited, perhaps only Series A or B can be used. However, it sometimes takes longer to obtain seedlings using only one of the two series than it does when seeds are moved through Series A and B concurrently. Also, if Series A is used alone and seedlings are obtained after seeds have been exposed to 5°C for 12 weeks, one does not know whether warm stratification is a necessary part of the dormancy-breaking protocol. Thus, when Series A or B is used alone, seeds may germinate, but one is not conducting an experiment per se. If the number of seeds is very limited, one could use only three dishes of seeds and move them each season of the year to simulated habitat temperatures (i.e., summer, autumn, winter, spring). We suggest starting with a temperature regime simulating temperatures in the habitat at the time of seed maturation and dispersal. However, this approach is not experimental. If seeds are locally produced, the same germination results may be obtained just as easily by planting seeds outdoors, where they would be exposed to natural temperature changes.
Conclusions By knowing the family to which a species belongs, one can immediately obtain information about whether the seeds might have seed or fruit coats that are impermeable to water (= physical dormancy); an undifferentiated embryo (= specialized morphological dormancy); differentiated, underdeveloped (small) embryo (= morphological or morphophysiological dormancy); or a fully developed embryo (nondormancy or physiological dormancy). Regardless of the type of embryo, data from a move-along experiment make it possible to learn much about the dormancy-breaking and germination requirements of a species by using only a few hundred seeds.
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References Barton, L. V. 1944. Some seeds showing special dormancy. Contributions from Boyce Thompson Institute 13:259–271. Baskin, C. C., and J. M. Baskin. 1988. Germination ecophysiology of herbaceous plant species in a temperate region. American Journal of Botany 75:286–305. Baskin, C. C., and J. M. Baskin. 1994a. Deep complex morphophysiological dormancy in seeds of the mesic woodland herb Delphinium tricorne (Ranunculaceae). International Journal of Plant Sciences 15:738–743. Baskin, C. C., and J. M. Baskin. 1994b. Warm plus cold stratification requirement for dormancy break in seeds of the woodland herb Cardamine concatenata (Brassicaceae), and evolutionary implications. Canadian Journal of Botany 73:608–612. Baskin, C. C., and J. M. Baskin. 1998. Seeds: Ecology, Biogeography, and Evolution of Dormancy and Germination. San Diego, CA: Academic Press. Baskin, C. C., J. M. Baskin, and M. A. Leck. 1993a. Afterripening pattern during cold stratification of achenes of ten perennial Asteraceae from eastern North America, and evolutionary implication. Plant Species Biology 8:61–65. Baskin, C. C., J. M. Baskin, and W. W. McDearman. 1993b. Seed germination ecophysiology of two Zigadenus (Liliaceae) species. Castanea 58:45–53. Baskin, C. C., J. M. Baskin, and O. W. Van Auken. 1998. Role of temperature in dormancy break and/or germination of autumn-maturing achenes of eight perennial Asteraceae from Texas, U.S.A. Plant Species Biology 13:13–20. Baskin, C. C., E. W. Chester, and J. M. Baskin. 1992. Deep complex morphophysiological dormancy in seeds of Thaspium pinnatifidum (Apiaceae). International Journal of Plant Sciences 153:565–571. Baskin, C. C., S. E. Meyer, and J. M. Baskin. 1995. Two types of morphophysiological dormancy in seeds of two genera (Osmorhiza and Erythronium) with an ArctoTertiary distribution pattern. American Journal of Botany 82:293–298. Baskin, J. M., and C. C. Baskin. 1973. Studies on the ecological life cycle of Holosteum umbellatum. Bulletin of the Torrey Botanical Club 100:110–116. Baskin, J. M., and C. C. Baskin. 1983. Germination ecology of Veronica arvensis. Journal of Ecology 71:57–68. Baskin, J. M., and C. C. Baskin. 1984. Germination ecophysiology of the woodland herb Osmorhiza longistylis (Umbelliferae). American Journal of Botany 71:687–692. Baskin, J. M., and C. C. Baskin. 1985. The annual dormancy cycle in buried weed seeds: a continuum. BioScience 35:492–498. Baskin, J. M., and C. C. Baskin. 1986a. Germination ecophysiology of the mesic deciduous forest herb Isopyrum biternatum. Botanical Gazette 147:152–155. Baskin, J. M., and C. C. Baskin. 1986b. Temperature requirements for after-ripening in seeds of nine winter annuals. Weed Research 26:375–380. Baskin, J. M., and C. C. Baskin. 1987. Temperature requirements for after-ripening in buried seeds of four summer annual weeds. Weed Research 27:385–389. Baskin, J. M., and C. C. Baskin. 1989. Seed germination ecophysiology of Jeffersonia diphylla, a perennial herb of mesic deciduous forests. American Journal of Botany 76:1073–1080. Baskin, J. M., and C. C. Baskin. 1990. Germination ecophysiology of seeds of the winter annual Chaerophyllum tainturieri: a new type of morphophysiological dormancy. Journal of Ecology 78:993–1004.
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Baskin, J. M., and C. C. Baskin. 1994. Nondeep simple morphophysiological dormancy in seeds of the mesic woodland winter annual Corydalis flavula (Fumariaceae). Bulletin of the Torrey Botanical Club 121:40–46. Baskin, J. M., C. C. Baskin, and X. Li. 2000. Taxonomy, anatomy and evolution of physical dormancy in seeds. Plant Species Biology 15:139–152. Baskin, J. M., C. C. Baskin, and M. R. McCann. 1988. A contribution to the germination ecology of Floerkea proserpinacoides (Limnanthaceae). Botanical Gazette 149:427–431. Grabe, D. F. (ed.). 1970. Tetrazolium Testing Handbook for Agricultural Seeds. Contribution no. 29 to the Handbook on Seed Testing. Association of Official Seed Analysts. Jacobsen, J. V., and E. Pressman. 1979. A structural study of germination of celery (Apium graveolens L.) seed with emphasis on endosperm breakdown. Planta 144:241–248. Martin, A. C. 1946. The comparative internal morphology of seeds. The American Midland Naturalist 36:513–660. Nikolaeva, M. G. 1969. Physiology of deep dormancy in seeds. Leningrad: Izdatel’stvo Nauka. [Translated from Russian by Z. Shapiro, National Science Foundation, Washington, DC.] Nikolaeva, M. G. 1977. Factors controlling the seed dormancy pattern. Pages 51–74 in A. A. Khan (ed.), The Physiology and Biochemistry of Seed Dormancy and Germination. Amsterdam: North-Holland. Parker, C., and C. R. Riches. 1993. Parasitic Weeds of the World: Biology and Control. Wallingford, UK: CAB International. Roberts, H. A. 1979. Periodicity of seedling emergence and seed survival in some Umbelliferae. Journal of Applied Ecology 16:195–201. Roberts, H. A., and J. E. Boddrell. 1984. Seed survival and periodicity of seedling emergence in four weedy species of Papaver. Weed Research 24:195–200. Roberts, H. A., and J. E. Boddrell. 1985. Temperature requirements for germination of buried seeds of Aethusa cynapium L. Weed Research 25:267–274. Steinbauer, G. P., and B. Grigsby. 1960. Dormancy and germination of the docks (Rumex spp.). Proceedings of the Association of Official Seed Analysts 50:112–117. Stokes, P. 1952. A physiological study of embryo development in Heracleum sphondylium L. I. The effect of temperature on embryo development. Annals of Botany 16:441–447. Stokes, P. 1965. Temperature and seed dormancy. Pages 746–803 in W. Ruhland (ed.), Encyclopedia of Plant Physiology, Vol. 15/2. Berlin: Springer-Verlag. Wallis, A. L. Jr. (ed.). 1977. Comparative Climatic Data through 1976. Asheville, NC: U.S. Department of Commerce, National Oceanic and Atmospheric Administration, Environmental Data Service, National Climatic Data Center. Williams, E. D. 1971. Germination of seeds and emergence of seedlings of Agropyron repens (L.) Beauv. Weed Research 11:171–181.
Chapter 9
Pollen Storage as a Conservation Tool Leigh E. Towill
Pollen storage creates the opportunity to make controlled crosses between individual plants widely separated in space and time. Though widely used in crop breeding, the technology has yet to find much application in plant conservation. Because pollen, like sperm and eggs, is haploid, pollen storage is not a substitute for seed or clonal preservation. To use the genetic information in pollen, another compatible reproductive individual must be available. This chapter provides a brief overview of pollen collection, handling, and storage (see also Akihama and Omura 1986; Barnabas and Kovacs 1997; Hanna and Towill 1995; Hoekstra 1995; Towill 1985). Pollen is the male gametophyte, or gamete-producing generation, of flowering and coniferous plants, and as such is haploid. In angiosperms, the pollen is either two-celled (binucleate or bicellular) or three-celled (trinucleate or tricellular) when it released from the anther of the flower. This developmental stage can have importance for storability and longevity. Most angiosperms produce bicellular pollens. Some families, most notably the Poaceae, produce tricellular pollens that tend to be short-lived (Brewbaker 1967). Grass pollens generally have a high water content and live only a few minutes to hours under ambient conditions, whereas other pollens can last days to months. Longevities can be extended in many pollens to months or years with storage under controlled humidities and temperatures. Historically pollen storage has been used for agricultural plant improvement and regeneration strategies. The use of stored and transported pollen for controlled breeding has become a regular tool for a number of plant groups. Hobbyists and breeders regularly exchange pollen for breeding of palms, cycads, gesneriads, Lilium, Rhododendron, and orchids. In fact, it 180
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is strange that the techniques are so infrequently used by botanic gardens and other ex situ plant conservation facilities. Pollen is easy to transport and can be stored, albeit for short periods, without sophisticated facilities. It has a number of potential uses for ex situ plant conservation: it allows use of pollen when male and female inflorescences are not synchronous, it allows controlled pollination between individuals, and it allows gene exchange between geographically isolated individuals, both wild and cultivated. In conservation programs, this can facilitate the effective use of founders and increase effective population size (Ne). For germplasm preservation, in which the time of use is not easily foreseen, pollen storage must be designed for indefinite periods. Although advocated in the literature (Bajaj 1987), systematic pollen banks have not been routinely developed except in some forest tree improvement programs (Mercier 1995). The development of a bank requires effective pollen collection, processing, storage, and use (Connor and Towill 1993). Many conservation scenarios exist in which pollen collection and storage are beneficial. For instance, if large numbers of seeds are to be generated from individuals flowering at different times, short-term storage facilitates the process. If subpopulations have become reproductively isolated, pollen manipulation, including storage, may facilitate crossing between fragments and increase fitness. For example, the Lakeside daisy (Hymenoxys acaulis var. glabra), a federally listed threatened species, was reduced to a single self-incompatible clone in Illinois. Part of the recovery effort involved the transfer of seeds and pollen from populations in other states to produce a reproductively viable population in Illinois (DeMauro 1994). Pollen exchange has been used for controlled pollination of geographically isolated individuals of threatened cycads in botanic gardens (Crosiers and Malaisse 1995) and for overcoming timing problems for the self-pollination of Amorphophallus titanium at the Fairchild Tropical Garden in Coral Gables, Florida, where pollen was stored until the female flowers were receptive. At the National Tropical Botanical Garden, Hawaii, wild populations of Brighamia insignis were artificially pollinated when it was observed that wild plants were exhibiting low levels of seed set after the decline of the wild pollinator. Pollen has been collected from wild individuals of two highly threatened plant species, Brighamia rockii and Pittosporum halophilum, and used to pollinate cultivated populations. As the management of cultivated plant populations becomes more sophisticated, it is likely that pollen will become an increasingly used tool for breeding.
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Collection Poor-quality pollen does not store well; that is, it loses both viability and vigor (pollen growth rate) quickly under conventional storage, such as in a refrigerator or freezer. Because immature and aged pollen usually are of poor quality, it is best to collect pollen just after anthesis. Pollen deteriorates quickly when left in the anther, probably because it experiences higher temperatures and humidity. Pollen collected from plants subjected to temperature stress or herbivory often is of poor quality (Delph et al. 1997). Collection timing can also influence quality; for example, grass species (Poaceae) extrude anthers early in the morning and shed short-lived pollen as the anther desiccates. As with seeds, to ensure adequate genetic representation, pollen should be collected from many individuals per population. Similar amounts should be collected from each individual, and samples (paternal lines) should be kept separate to help in equalizing founder representation. The amount of pollen collected is species and project dependent. Some wind-pollinated species produce prodigious amounts of pollen that can be collected easily. Insect-pollinated species typically have small anthers and produce little pollen per anther. Collecting from these can be tedious; often anthers are obtained using combs and dried as a unit. Small whole flowers may also be dried. Briefly crushing the anthers frees the pollen, and it is not necessary to separate the pollen from the debris. Other species such as orchids and asclepiads produce pollen in pollinia that may necessitate other collecting techniques.
Viability and Vigor Assaying viability is crucial for efficient storage and use of pollen. Pollen should be assayed soon after collection to determine initial viability. If it is low, additional samples can be taken while individuals are still flowering. The most common assays include seed or fruit set, in vitro germination, and staining. Seed set is a definitive test of functionality but is timeconsuming and cannot easily provide percentages of viability. Pollen tube growth in isolated stigma-styles works for many taxa, but incompatibility relationships must be known and a supply of stigma-styles available. A detailed assessment of these methods is provided by Stone et al. (1995). In vitro germination on agar or in liquid medium can assay for both viability and pollen growth rate. However, existing media do not work for all
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taxa (i.e., some Poaceae). Many in vitro germination assays have been developed with crop systems, but detailed information for wild species is scanty. The medium of Brewbaker and Kwack (1963) has been useful for many species if sucrose and boron concentrations are optimized. The pollen germination rate varies tremendously between species, ranging from minutes for some Poaceae to days for some conifers. Although it is generally assumed that in vitro germination is correlated with seed set, few references provide critical data. Pollen vigor can affect this relationship because low-vigor pollen may germinate but not be able to sire seed. Given these limitations, staining tests for viability are simple, quick, and broadly applicable and use little pollen. Staining tests can be useful to monitor viability decline during storage. However, these tests do not always correlate with seed set, but data are sparse. The most common stain is the fluorochromasia test (Heslop-Harrison and Heslop-Harrison 1970), which generally correlates with in vitro germination. After staining, live pollen fluoresces under ultraviolet light, whereas nonviable pollen does not. A suitable fluorescent microscope is needed to count the grains. Tetrazolium tests are also widely used (Shivanna and Johri 1989). Pollen that is not viable does not stain, but distinguishing between stained and nonstained grains can be somewhat subjective. The concept of pollen vigor, which reflects quality, is less well defined than that for seed (Hampton and TeKrony 1995), and there is no simple measure of it (Heslop-Harrison et al. 1984; Shivanna et al. 1991). More vigorous pollen germinates more quickly and often produces a longer pollen tube. Within a collection, vigor of the pollen declines in storage before loss of viability.
Storage Once pollen has been collected, processing for storage should be undertaken as quickly as possible. Processing consists of desiccation, packaging, and storing at a desired temperature. The major factors that influence longevity of collected pollen are moisture content, storage atmosphere, and storage temperature.
Developmental Stage At anthesis pollen is either bicellular or tricellular, and this is often correlated with storability (Brewbaker 1967). Pollen that is bicellular, especially
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if it is not large, may be more tolerant to desiccation than tricellular types. In tricellular pollen the generative nucleus has already undergone a division and undergoes rapid growth when placed on a stigma (or in a germination medium). Tricellular pollen can be considered to have proceeded in development and may not tolerate extensive desiccation. This is analogous to a recalcitrant seed quickly entering a germination phase and losing desiccation tolerance (Vertucci and Farrant 1995; Pammenter and Berjak 1999). There are many exceptions to this generalization; for example, large bicellular pollen from cucurbits has a very short expected longevity (Digonnet-Kerhoas et al. 1989), and tricellular pollen from sugar beet (Hecker et al. 1986) and Pennisetum sp. (Hanna et al. 1986; Hanna 1990) tolerates desiccation and has a longer expected longevity. In some families, all taxa have exclusively bicellular or tricellular pollen, whereas other families contain both pollen types (Brewbaker 1967). It should be recognized that desiccation tolerance is not discrete; a range of tolerances exist. As with seed, it is expedient to consider tolerant, intermediate, and sensitive classes.
Storage Conditions Given the aforementioned caveats, one can generalize about storage capability. Pollen from the Poaceae and Asteraceae often is short-lived, whereas that from the Solanaceae and Rosaceae is longer-lived. Pollen longevity under a given set of storage conditions must be empirically determined. There is an urgent need to develop species-specific protocols that conservationists can apply to effectively manage pollen holdings. Both moisture content and storage temperature influence viability.
Desiccation-Tolerant Pollen Partial desiccation enhances longevity at any given storage temperature. It is recommended for effective storage that pollen be dehydrated to a moisture content of 5–10 percent on a fresh weight basis. This can be achieved by desiccation of the pollen in a thin layer at ambient conditions if humidity is low or by equilibration in chambers with a known relative humidity. The latter is recommended because overdrying is avoided and pollen is not expended solely to determine the moisture content. Pollen moisture roughly equilibrates to a given relative humidity over a 1- to 4-day period.
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Equilibration over a saturated calcium nitrate salt solution (ca. 61 percent RH) gives a moisture content of roughly 10 percent, and equilibration over a saturated magnesium chloride solution (ca. 33 percent RH) gives roughly 7 percent; these values are influenced by the chemical composition of the pollen. At these moisture contents all liquid water has been removed from the pollen, so freezing damage to the cells is minimized. Examples of isotherm data for Typha pollen can be found in Buitink et al. (1998). These examples demonstrate the relationships between equilibrium moisture content after exposure to a given relative humidity and temperature. To reduce damage during rehydration, storage at moisture contents of about 5 to 10 percent is desirable (Hoekstra 1995). Generally, lower storage temperatures lead to greater longevity. Thus, longevity is short at room temperature (minutes to hours, but months for conifer pollen) and much longer at –18°C (refrigerator freezer temperature—months to years). Longevity is highly species-specific; conifers retain high levels of pollen viability for years, whereas some solanaceous pollen may survive for only 1–2 years at –18°C (for examples, see Towill 1985 and Hong et al. 1999). There is an increasing appreciation of the complexity of the problem and a better understanding of the biophysical and biochemical factors and interactions that occur; these might ultimately be used to predict longevity, as described for seed (Walters 1998). Storage atmosphere also influences longevity. Research has shown that freeze-dried (or vacuum-dried) pollen stored in a vacuum and dried pollen stored in a nitrogen atmosphere have greater longevity than samples stored in an air atmosphere (Jensen 1964, 1970). Improvements in viability usually are greater at higher storage temperatures, and the extra effort to control storage atmosphere may not be warranted if lower temperatures are used. For long-term storage, the inclusion of vacuum or nitrogen atmosphere storage may be desirable if storage temperatures lower than –20°C are not available.
Desiccation-Sensitive Pollen Not all pollens tolerate the same extent of desiccation. Pennisetum pollen can tolerate desiccation to below 5 percent moisture (fresh weight basis) content (Hanna and Towill 1995), and Zea pollen can tolerate about 10 percent (Barnabas et al. 1988), but Triticum pollen tolerates much less desiccation (Barnabas and Kovacs 1997). It is difficult to define desiccation
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limits in species such as Triticum, where pollen metabolizes rapidly and longevity is short. In general, for desiccation-sensitive pollen, effective storage depends on balancing the moisture content with potential desiccation injury. Because samples stored at temperatures down to about –20°C show very short longevities (hours to weeks), cryogenic storage is generally used. Fortunately, many desiccation-sensitive pollens survive liquid nitrogen exposure if some adjustment of moisture content is made. This has been most successful for species in which desiccation injury does not occur until well below the limit of freezable water. In all cases, pollen should be stored in a vial that does not allow moisture loss or gain.
Retrieval, Distribution, and Use When pollen is needed, samples are removed from storage and warmed to room temperature. If a desiccation-tolerant pollen is stored at a low moisture, the warming rate does not influence viability, and samples can be taken directly to room temperature. Although pollen can be refrozen, this is not recommended. Avoid opening cold vials; this may allow condensation on the pollen. Once the pollen is at room temperature, deterioration proceeds as before storage, and the sample should be used for pollination as soon as possible. If pollen is to be exchanged between locations, shipment with ice packs to slow deterioration is recommended. Rehydrating pollen that has been desiccated to low moisture levels (below about 20 percent moisture content, fresh weight basis) usually is necessary before viability can be estimated using in vitro germination tests but may not be needed for pollinations. Pollen exhibits imbibitional damage if rapidly exposed to water or if exposed to water vapor at low temperatures (Crowe et al. 1989). Therefore, rehydration is accomplished by placing the sample at a high humidity, such as over water, for a few hours at room temperature before use.
Conclusions Effective pollen storage entails planning and knowledge of conditions that influence viability and vigor. Available information about related species, genera, or families can guide the practitioner. Control of moisture content is critical for both desiccation-tolerant and desiccation-sensitive types. Generally, storage at lower temperatures is beneficial, but deterioration, even
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at temperatures as low as –80°C, may still occur. Nevertheless, the establishment of a pollen storage program can be a valuable tool to support the conservation and management of wild species. References Akihama, T., and M. Omura. 1986. Preservation of fruit tree pollen. Pages 101–112 in Y. P. S. Bajaj (ed.), Biotechnology in Agriculture and Forestry, Vol. 1: Trees I. Berlin: Springer-Verlag. Bajaj, Y. P. S. 1987. Cryopreservation of pollen and pollen embryos, and the establishment of pollen banks. International Review of Cytology 107:397–420. Barnabas, B., and G. Kovacs. 1997. Storage of pollen. Pages 293–314 in K. R. Shivanna and V. K. Sawhney (eds.), Pollen Biotechnology for Crop Production and Improvement. Cambridge, UK: Cambridge University Press. Barnabas, B., G. Kovacs, A. Abrany, and P. Pfahler. 1988. Effects of pollen storage by drying on the expression of different agronomic traits in maize (Zea mays L.). Euphytica 39:221–225. Brewbaker, J. L. 1967. The distribution and phylogenetic significance of binucleate and trinucleate pollen grains in the angiosperms. American Journal of Botany 54:1069–1083. Brewbaker, J. L., and B. H. Kwack. 1963. The essential role of calcium ion in pollen germination and pollen tube growth. American Journal of Botany 50:859–865. Buitink, J., C. Walters, F. A. Hoekstra, and J. Crane. 1998. Storage behavior of Typha latifolia pollen at low water contents: interpretation on the basis of water activity and glass concepts. Physiologia Plantarum 103:145–153. Connor, K. F., and L. E. Towill. 1993. Pollen-handling protocol and hydration/dehydration characteristics of pollen for application to long-term storage. Euphytica 68:77–84. Crosiers, C., and F. P. Malaisse. 1995. Ex situ pollination and multiplication of Encephalartos laurentianus De Wild. (Zamiaceae-Cycadales). Biodiversity and Conservation 4:767–775. Crowe, J. H., F. A. Hoekstra, and L. M. Crowe. 1989. Membrane phase transitions are responsible for imbibitional damage in dry pollen. Proceedings of the National Academy of Sciences, USA 86:520–523. Delph, L. F., M. H. Johannsson, and A. G. Stephenson. 1997. How environmental factors affect pollen performance: ecological and evolutionary perspectives. Ecology 78:1632–1639. DeMauro, M. M. 1994. Development and implementation of a recovery program for the federal threatened Lakeside daisy (Hymenoxys acaulis var. glabra). Pages 298–321 in M. L. Bowles and C. J. Whelan (eds.), Restoration of Endangered Species. Cambridge, UK: Cambridge University Press. Digonnet-Kerhoas, C., G. Gay, J. C. Duplan, and C. Dumas. 1989. Viability of Cucurbita pepo pollen: biophysical and structural data. Planta 179:165–170. Hampton, J. G., and D. M. TeKrony. 1995. Handbook of Vigour Test Methods. 3rd edition. Zurich: The International Seed Testing Association. Hanna, W. W. 1990. Long-term storage of Pennisetum glaucum (L.) R.Br. pollen. Theoretical and Applied Genetics 79:605–608.
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Hanna, W. W., G. W. Burton, and W. G. Monson. 1986. Long-term storage of pearl millet pollen. Journal of Heredity 77:361–362. Hanna, W. W., and L. E. Towill. 1995. Long-term pollen storage. Pages 179–207 in J. Janick (ed.), Plant Breeding Reviews, Vol. 13. New York: Wiley. Hecker, R. J., P. C. Stanwood, and C. A. Soulis. 1986. Storage of sugarbeet pollen. Euphytica 35:777–783. Heslop-Harrison, J., and Y. Heslop-Harrison. 1970. Evaluation of pollen viability by enzymatically induced fluorescence: intracellular hydrolysis of fluorescein diacetate. Stain Technology 45:115–120. Heslop-Harrison, J., Y. Heslop-Harrison, and K. R. Shivanna. 1984. The evaluation of pollen quality, and a further appraisal of the fluorochromatic (FCR) test procedure. Theoretical and Applied Genetics 67:367–375. Hoekstra, F. A. 1995. Collecting pollen for genetic resources conservation. Pages 527–550 in L. Guarino, V. R. Rao, and R. Reid (eds.), Collecting Plant Genetic Diversity. Technical Guidelines. Wallingford, UK: CAB International. Hong, T. D., R. H. Ellis, J. Buitink, C. Walters, F. A. Hoekstra, and J. Crane. 1999. A model of the effect of temperature and moisture on longevity in air-dry storage experiments. Annals of Botany 83:167–173. Jensen, C. J. 1964. Pollen storage under vacuum. Pages 133–146 in Royal Veterinary and Agricultural College Yearbook. Copenhagen: Royal Veterinary and Agricultural College. Jensen, C. J. 1970. Some Factors Influencing Survival of Pollen on Storage Procedures. FAO/IUFRO Working Group meeting on sexual reproduction of forest trees, reprint no. 148. Mercier, S. 1995. The role of a pollen bank in the tree genetic improvement program in Quebec (Canada). Grana 34:367–370. Pammenter, N. W., and P. Berjak. 1999. A review of recalcitrant seed physiology in relation to desiccation-tolerance mechanisms. Seed Science Research 9:13–37. Shivanna, K. R., and B. M. Johri. 1989. The Angiosperm Pollen: Structure and Function. New York: Wiley. Shivanna, K. R., H. F. Linskens, and M. Cresti. 1991. Pollen viability and vigor. Theoretical and Applied Genetics 81:38-42. Stone, J. L., J. D. Thomson, and S. J. Dent-Acosta. 1995. Assessment of pollen viability in hand-pollination experiments: a review. American Journal of Botany 82:1186–1197. Towill, L. E. 1985. Low temperature and freeze-/vacuum-drying preservation of pollen. Pages 171–178 in K. K. Kartha (ed.), Cryopreservation of Plant Cells and Organs. Boca Raton, FL: CRC Press. Vertucci, C. W., and J. Farrant. 1995. Acquisition and loss of desiccation tolerance. Pages 237–272 in J. Kigel and G. Galili (eds.), Seed Development and Germination. New York: Marcel Dekker. Walters, C. 1998. Understanding the mechanisms and kinetics of seed aging. Seed Science Research 8:223–244.
Chapter 10
Tissue Culture as a Conservation Method: An Empirical View from Hawaii Nellie Sugii and Charles Lamoureux
Hawaii has been called the Endangered Species Capital of the World, with more than one-half of the native Hawaiian flora at risk (U.S. Fish and Wildlife Service 1999). The flora of Hawaii has the highest percentage of endemism found in a single large island group in the world, with 90 percent of all native Hawaiian angiosperms and 70 percent of the pteridophytes being endemic (Wagner et al. 1999). Although Hawaii contains less than 0.2 percent of the total land area of the United States, its native plants make up approximately one-third of the United States’ federally listed endangered and threatened (302 endangered, 10 threatened) plant species. Another 25 percent (approximately 250) are in a significant state of decline and depleted enough to be listed as species of concern (U.S. Fish and Wildlife Service 1999). More than 100 species are known from 20 or fewer remaining wild individuals. Eleven of these species are so rare that they are currently known only from a single plant remaining in the wild. In addition, there are at least four others that no longer exist in the wild but survive only in tissue culture or as cultivated plants in greenhouses or gardens (Center for Plant Conservation 1999). In 1991, Lyon Arboretum initiated the Rare Hawaiian Plant Program (RHPP), using micropropagation as a tool for plant genetic conservation. The objective was to prevent further extinction of Hawaiian plant taxa by propagating plants for use in restoration and reintroduction and initiating and maintaining an in vitro germplasm collection of critically endangered plants included in the Genetic Safety Net Listing (GSNL). Lyon Arboretum works cooperatively with four other Hawaiian botanical gardens in the Center for Plant Conservation network, various state and federal agencies, 189
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private conservation agencies, environmental organizations, and major private landowners. To date, more than 130 federally listed endangered taxa have been successfully grown at Lyon Arboretum using micropropagation techniques, and plants have been supplied to restoration and reintroduction projects. The Lyon Arboretum Micropropagation Laboratory also maintains a large in vitro germplasm collection of ethnobotanically important crops such as taro (Colocasia esculenta) and banana (Musa). Plant micropropagation has become an indispensable tool for plant genetic conservation, especially where conventional propagation efforts have failed or proven difficult. Micropropagation is particularly useful in situations in which collected seed propagules are immature, tiny, or recalcitrant or are in short supply. The seeds are germinated in vitro and stored as living germplasm collections or are germinated for future conservation projects. When seeds are unavailable, clonal propagules can be initiated, propagated, and maintained in vitro. Preserving the integrity of the original plant genotype and genetic stability is of utmost importance. The selection of suitable plant material and explants in the field and proper surface disinfestation, plant medium, and culture conditions are all considered crucial for successful establishment of in vitro cultures. Hawaii has the highest percentage of angiosperm endemism found in a single large island group anywhere in the world. The uniqueness of Hawaii’s native flora is attributed to its isolation together with the various combinations of topography, weather, and geology, creating a diverse range of microclimates (Sohmer and Gustafson 1996; Wagner et al. 1999). Seed from ancestors of the Hawaiian flora carried by birds, wind, or water colonized the islands (at a rate of approximately one new successful seed plant introduction per 100,000 years) and over time spread, adapted to their microenvironments, and eventually evolved into distinct species (Grierson and Green 1996). After human discovery and colonization of the islands, significant decline and in many cases extinction of species have occurred. About 1,000 nonindigenous plants are well established, and many have become pests and noxious weeds in the Hawaiian landscape, often replacing native ecosystems. Additional biological pressures include introduced diseases, ungulates, rats, and insects. Urban and agricultural expansion has also contributed to the rapid decline of the native flora (Cuddihy and Stone 1990; Wester 1992; Gon and Matsuwaki 1999).
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The GSNL, generated by the Hawaii Rare Plant Restoration Group (HRPRG), currently lists 103 federally endangered Hawaiian taxa that are in imminent danger of extinction and possess 20 or fewer plants in the wild. The State of Hawaii’s Department of Land and Natural Resources (DLNR) has adopted this list and uses it to direct its collections, monitoring and propagation activities. The Lyon Arboretum is part of the HRPRG, a collaborative network of several organizations, agencies, and private landowners. This network includes Waimea Arboretum, National Tropical Botanical Garden (NTBG), Amy Greenwell Ethnobotanical Gardens, Honolulu Botanical Gardens, DLNR, U.S. Fish and Wildlife Service, U.S. Army, Hawaii Army National Guard, Center for Plant Conservation Hawaii (CPCH), Hawaii Volcanoes National Park, the Nature Conservancy Hawaii, and private landowners. Through this concerted effort, threatened Hawaiian plants can be identified, collected, monitored, propagated, and stored for possible restoration efforts in the future. This group identifies the critically at risk Hawaiian plant species and develops and initiates collection strategies for the landowners and propagators by identifying deficiencies in sampling and ex situ germplasm inventories. The Lyon Arboretum Micropropagation Laboratory is the designated in vitro propagation facility for the critically endangered Hawaiian plant species collected as part of the GSNL project. It also maintains a large in vitro germplasm collection of Colocasia esculenta and Musa, particularly Hawaiian cultivars that are of ethnobotanical importance to the indigenous people of Hawaii. Many of the original, introduced cultivars are in danger of disappearing because of a lack of cultural use and because of diseases such as the taro leaf blight (Jackson 1999). In vitro methods are especially important for the conservation of these plant species, which are clonally propagated and are difficult to conserve as seed (IBPGR 1984; Vuylsteke 1989).
Plant Micropropagation Plant micropropagation technology has been developed and redefined continuously over the past 30 years and has received an increasing amount of interest as a tool for plant genetic conservation (Dodds 1991). It is an indispensable tool in the areas of biotechnology, genetic engineering, and plant
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propagation for the production of large quantities of clonal material for the commercial agricultural industry (Kyte and Kleyn 1996). Added benefits of tissue culture include disease indexing and elimination of viruses by using meristem-tip culture and the ease of transport of in vitro cultures (Ashmore 1997; Toll 1999). In Hawaii, micropropagation has become a common method of propagation and storage for many Critically Endangered plant taxa. It is especially common when the species are difficult to propagate using conventional methods; viable propagules are scarce because of reproductive problems or difficulty in collecting plants; plants have very small, recalcitrant, or immature seeds or spores; plant numbers are low, reducing the amount of material available for propagation; and propagules are of poor quality because of disease or nutritional deficits. Theoretically, entire plants can be produced from a single plant cell, and many species can regenerate into plants when grown in vitro, free of microorganisms and on an appropriate nutrient medium. However, some species still cannot be propagated using known techniques. Explants for micropropagation are either vegetatively or sexually derived, and the most commonly used vegetative explants include apical, axillary, and root meristems, stem internodes, leaves, and inflorescences. They produce plantlets called clones, which share an identical genotype with each other and with the original parent. Clones eventually can be multiplied to produce more plants in a process known as cloning or clonal propagation. Sexually derived explants such as seed, embryos, ovules, spores, and pollen produce plantlets with unique genotypes (George 1993; Kyte and Kleyn 1996). Once the clones or seedlings are established in culture, the plantlets continue to grow miniaturized leaves and stems and remain in a juvenile or juvenile-like state. It is possible to maintain these plants for several months to years in small vessels within a controlled, sterile growing environment (Withers 1985; George 1993). Tissue culture of vegetative explants may also have a rejuvenating effect on mature, senescent plants and restore juvenility to the cultured explants (Pierik 1987; Ashmore 1997), an extremely beneficial outcome for plants that have lost their vegetative vigor and reproductive capacity. One of the goals for the RHPP is to clone every individual of all Hawaiian taxa having 20 or fewer plants in the wild for the purpose of germplasm storage (Sugii and Lamoureux 1998; HRPRG 1999a). Mature seeds are also collected from these individuals when available, but in many of these
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species, reproductive problems are common and result in plants that produce no viable seeds or seeds that abort spontaneously before they reach maturity. In this case, immature seeds are collected and embryo or ovule culture is attempted (George 1993; Dodds 1991; U.S. Fish and Wildlife Service 1998). For recalcitrant seed, where standard seed storage techniques have proved inadequate, in vitro germination and storage of the seedlings is currently the only efficient way to preserve the genetic diversity contained in the seeds (Callow et al. 1997). Micropropagation and in vitro storage provide good short- to medium-term storage of plant germplasm, but cryopreservation may eventually be able to offer a long-term storage solution (Dodds 1991). It is expected that a proportion of the Hawaiian taxa will be amenable to cryopreservation, provided that the tissue culture protocols are adequate for the species (Sakai 2000; Engelmann 2000). HRPRG does not intend to use this technique in the near future because of financial and infrastructural constraints. Genetic variation in tissue culture has been recognized for many years in many plant species (Ashmore 1997). Tissue culture instability, or somaclonal variation, is recorded at the karyotypic, morphological, biochemical, and molecular level and can be generated at any time during the tissue culture process. These modifications may manifest themselves as heritable mutations, which can be passed to the progeny of the regenerated plants. Some of these mutations are obvious and directly affect the physical aspect of the plant’s growth; the majority are detrimental to the plant. Other mutations are difficult or impossible to interpret, especially when they occur on the genetic level (Scowcroft 1984; Skirvin et al. 1994; Ronchi 1995; Villordon and Labonte 1996). When working with threatened Hawaiian plants, it is of the utmost importance to preserve the integrity of the original plant genotype when using micropropagation. Any variable that may jeopardize genetic stability during the culturing process must be minimized. The ability to establish in vitro cultures of plants that can regenerate normally and maintain a high degree of genetic stability depends on several factors, such as selection of suitable explants, adequate postharvest handling, proper preparation of plant material for culture (e.g., surface decontamination), and the optimization of culture and storage conditions (Pierik 1987; George 1993). The selection of suitable plant material pertains not only to the type of explant used for culturing but also to the time of harvest, the juvenility of
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the material, and the general health of the plant (Pierik 1987; Bonga and Von Aderkas 1992). Vegetative plants generally are more responsive to in vitro culture than flowering plants, so cuttings from actively growing plants are preferred (Pierik 1987). Because of the lack of adequate information on the plant’s phenology, several field visits may be necessary before propagules in the proper physiological stage can be collected. Even when the time of vegetative growth is known, collections are hampered by several factors: adverse weather conditions, arduous or dangerous collection conditions necessitating the services of professional collectors, limited accessibility because of private landownership, lack of suitable propagules because of poor health, and lack of funding for collecting expeditions. Usually adult plants are used for explant collection for propagation. In some cases, it is almost impossible to induce adventitious organ formation in vitro in adults, particularly for trees and woody shrubs (Pierik 1987; Dodds 1991; Bonga and Von Aderkas 1992). Sometimes this difficulty can be overcome by collecting explants from the basal part, or juvenile zone of the tree. Juvenile plant tissues generally are found on the lower branches of the plant and as suckers arising from the stem, root, or stump (Bonga and Von Aderkas 1992). Many of the threatened Hawaiian natives, especially the woody species, survive only as adult specimens. Juvenile explant tissue can be difficult to collect because of a lack of basal suckering or sparse branching due to the plant’s natural habit or poor health. For these plants, explant material is still harvested and propagation attempted when it can be sacrificed with minimal effect to the plant. Also, for plants experiencing rapid decline and possibly death, explants are collected as an emergency salvage attempt. Seeds are optimally collected just before fruits dehisce or fall from the plant (Guarino et al. 1995). The maternal parentage of seeds collected in this manner is ensured, but, more importantly, they remain protected and generally sterile within the intact fruit (Dodds 1991). Disinfestation is needed only for the exterior of the fruit before seed removal. When intact fruit cannot be collected, the seed propagules harvested should show no visible signs of microbial growth or other damage (HRPRG 1999b). In vitro cultures usually are more successful if explants are taken from plants that are in reasonably good health. Any decline either by disease or senescence tends to be irreversible and will hinder progress of the explants during micropropagation (Pierik 1987; George 1993). Many Critically Endangered Hawaiian plant populations are very small, dispersed, senes-
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cent, and in decline. These problems often result in the premature death of existing plants and a lack of seedling regeneration in the natural population (Falk and Holsinger 1991). The poor condition of the remaining individuals renders collection of viable material suitable for in vitro culture and successful micropropagation extremely difficult. The postharvest handling of plant propagules can have a direct influence on the successful establishment of in vitro cultures. Collectors are encouraged to submit plant samples as soon as possible because delays may seriously reduce sample viability. Loss of sample viability commonly results from several factors, including exposure to excessive temperatures, creation of anaerobic conditions through prolonged storage in plastic bags or tight packing, storage under conditions conducive to fungal and bacterial growth, and physical damage during transport (Guarino et al. 1995; HRPRG 1999b). Plant propagules submitted to the Lyon Arboretum Micropropagation Laboratory are almost always field collected, often from plants in poor condition and heavily contaminated. As a result, it tends to be exceedingly difficult to produce viable aseptic plant cultures from these explants. Sometimes the disinfestment treatment can damage or destroy the explant. In the case of field-collected vegetative cuttings, whenever possible the cuttings are established in the greenhouse before micropropagation is attempted. The new growth on these plants is used as starting material for tissue culture because it tends to be cleaner and easier to disinfect (Bonga and Von Aderkas 1992). Once in the laboratory, the plant materials are cleaned of debris, trimmed (vegetative cuttings), rinsed with water, and prepared for disinfestation. The treatments and soaking times vary according to the plant material and the sensitivity of the plant tissue to the cleaning agent. Also, morphological differences (seed, leaf, meristem, root) and their physical characteristics (hirsute, glabrous) play a large part in determining how surface disinfestation proceeds. Careful monitoring of the plant material throughout the disinfestation process must be done to prevent damage from oversterilization. Highly contaminated material may need one or more pretreatments before the actual sterilization of the final explant. The pretreatments may consist of long water rinses, 3 percent hydrogen peroxide dips, 70 percent ethanol dips, or 0.1 percent Physan 20™ soaks. Physan 20™(Benzylkonium chloride, Maril Products, Inc., Tustin, California) is a general bactericide, fungicide, viricide, and algaecide that has been found to be useful in the decontamination of plant propagules.
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The most common disinfectant used is sodium hypochlorite, found as a 5 percent solution in household bleach. For disinfestation, the soaking solutions are prepared fresh by diluting the bleach to a 5 percent and 10 percent (volume to volume in water) solution. One drop of Tween 20™ is added for every 100 mL of solution as a surfactant. Tween 20™ (Polyoxyethylenesobitan monolaurate, Unigema, a business unit of ICI Americas, Inc., New Castle, Delaware) is a non-ionic, nontoxic, wetting agent that enhances the cleaning action of the sterilizing solution by allowing the sterilant to have good contact with the surfaces of the plant tissue. For highly contaminated or hard-to-clean material, an additional disinfestation is needed, and a 5-ppm solution of Plant Preservative Mixture (PPM, Plant Cell Technologies, Washington, D.C.) is used as a 2-hour to 2-day soak. The length of soaking is determined by the sensitivity of the plant tissue to PPM. PPM may also be included in the initial culture medium at a rate of 0.5 ppm. Vegetative propagules are soaked initially in a 10 percent bleach solution and Tween 20™ for 15–30 minutes. The explants are trimmed further into final explant pieces with the aid of a dissecting microscope in a petri dish containing a 5 percent bleach solution. The final explants are placed in a 5 percent bleach solution for 1–15 minutes, rinsed well with sterile water, and placed on the appropriate plant medium. For highly contaminated, hirsute or woody tissue, the disinfestation procedure remains the same except that the initial soak in 10 percent bleach is no more than 15 minutes and the final explant is placed in 5 ppm PPM overnight. Intact immature and mature fruits usually are soaked in a 10 percent bleach solution for 30 minutes to 3 hours, depending primarily on the size and thickness of the exocarp. The containers holding the fruit are brought into the transfer hood and the seeds excised under aseptic conditions. Fruits with thick exocarps, especially with fuzzy, pitted, or grooved surfaces, may be dipped in 95 percent ethanol and flamed briefly for an additional sterilization step. The seeds are left whole (aseptic seed culture), or ovule or embryo culture is performed. Loose mature seeds are washed thoroughly in water then soaked in a 10 percent bleach solution for 15 minutes to overnight, depending on the morphology of the seed coat. To enhance germination, the seeds may be soaked, hot water treated, scarified, or dipped in 95 percent ethanol and quickly flamed before they are placed on the germination medium. Fern spores are placed in filter paper packets, soaked in water for 30 minutes, then subjected to a 30-minute sterilization with a 10 percent
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bleach solution and Tween 20™. After sterilization, the packets are given a thorough soaking in sterile water. The packets are then removed from the water, opened, and inverted onto petri dishes containing half-strength Knudson medium (Knudson 1946). The media formulations commonly used at Lyon Arboretum for micropropagation are based on the Murashige and Skoog (MS; 1962), Woody Plant Medium (WPM; Lloyd and McCown 1980), and Knudson (1946) formulas. The MS medium originally was developed for the tissue culture of tobacco but has become one of the most widely used medium formulations in the culture of many kinds of plant tissue worldwide (Kyte and Kleyn 1996). Different variations of this medium are made not only for the purpose of inducing specific plant responses but also to accommodate the nutrient needs of the plants. In Hawaii, the plant genetic conservation effort encompasses a large and diverse number of plant species for which very little or no propagation information exists. In almost all cases, there is neither time nor sufficient plant material to optimize the culture conditions. The type of plant propagules submitted and related past experience usually determine protocol and medium selection. Four basic types of tissue culture are performed at the Lyon Arboretum Micropropagation Laboratory: seed, embryo, ovule, and organ culture. Mature seeds generally are placed in seed storage or propagated in the greenhouse using conventional seed-sowing practices. They are placed into in vitro culture if the seeds are small, recalcitrant, rare, or difficult to germinate or if germplasm storage is requested. The germinated seedlings arising from in vitro seed sowing may be placed in short- or medium-term germplasm storage or grown for future conservation projects. In most cases, seeds are germinated and maintained on a medium containing only one-half of the MS macronutrients and micronutrients (1/2MS) with no hormones. Immature seeds benefit the most from in vitro germination and are suitable for in vitro seed sowing, ovule, and embryo culture. In ovule culture, the ovary tissue and embryo are left intact, whereas in embryo culture the embryo is excised from the seed. Embryo culture is also used to overcome seed dormancy and shorten the germination time (Bridgen 1994). For ovule and embryo culture, a 1/2MS medium is used, which may contain a higher or lower sugar concentration, gibberellic acid, polyvinylpyrrolidone (PVP), auxin, or cytokinin. Germinated seedlings usually are maintained on 1/2MS or MS medium.
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The regeneration of plants from vegetative explants that possess either meristematic or nonmeristematic tissue is accomplished through organ culture or organogenesis. Organogenesis is the development of plant organs directly on the explant surface or on an intervening callus phase (Pierik 1987; George 1993; Kyte and Kleyn 1996). Usually a 1/2MS medium supplemented with hormones such as auxin and cytokinin is used to induce shoot and root growth. The ratio of auxin to cytokinin often is manipulated to promote a particular type of growth (Coenen and Lomax 1997). In some cases, auxin is omitted during culture initiation to promote shoot growth and is added later to induce root formation. Once whole plants are regenerated, they are maintained on hormone-free 1/2MS, MS, or WPM medium. Apical and axillary meristems induced to produce direct shoot regeneration have been known to possess greater genetic stability (Ng and Ng 1991). Plants derived from unorganized callus cultures originating from nonmeristematic tissue such as stem internodes and leaves are generally thought to be more genetically unstable, which may manifest as abnormal phenotypes. During callus induction, differentiated cells, usually from the parenchyma, must be de-differentiated before cell division can occur. This is accomplished usually through the use of plant growth regulators. The unnatural manipulation of plant cells in tissue culture may result in the unintentional selection of rapidly growing somaclonal variants (Pierik 1987; Skirvin et al. 1994). Because of these problems, callus culture generally is not used routinely to preserve Hawaiian plant tissue. Sometimes, however, only nonmeristematic propagules are collected because they are the only kind of explant that can be sacrificed for propagation. Also, in some cases callus induction is the most responsive to in vitro propagation and the only way to regenerate plants. In this case, organogenic or shoottype callus, as opposed to somatic or undifferentiated-type callus, is preferred (Dodds 1991). If the explant contains meristematic tissue, such as apical and lateral meristems, cell division can occur without the cells going through intermediate de-differentiation, thereby reducing the risk of somaclonal variants (Pierik 1987). The acclimatization, or weaning, of micropropagated plants to the greenhouse is the next challenge. Plants undergoing this transition are subjected to high stress levels caused by lower relative humidities, higher light levels, and a septic environment. Plants grown in vitro are morphologically and functionally different from those grown in the greenhouse or field. As
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a result of high humidity in the culture vessel, the leaves of in vitro grown plants do not possess an adequate cuticle or fully functional stomata. These plants are also heterotrophic, not dependent on photosynthesis because of the nutrients provided in the culture medium and the lower gas exchange. Roots are also specialized for the nutrient uptake in the culture medium (Preece and Sutter 1991; Hartmann et al. 1997). Tissue-cultured plants are established initially on the mist bench, with plants gradually exposed to lower humidity levels. Dilute foliar feeding is used to encourage plant growth, and fungicide drenches are intermittently applied for disease prevention. Once new growth is observed, the plants are gradually moved to lower humidity levels until they can be placed on the regular greenhouse bench. During this hardening-off period, the plants become autotrophic and develop leaves and roots better suited to the greenhouse environment (Hartmann et al. 1997).
Examples of Endangered Hawaiian Species That Have Benefited from Tissue Culture Lobelia monostachya, Campanulaceae This species was thought to be extinct, having been last seen and collected in the 1920s, when only four plants were known to be growing on the dry exposed slopes in the mesic forests of the southern Ko’olau Mountains on the island of Oahu. Since then it was thought to be extinct and is listed as such in the Manual of the Flowering Plants of Hawaii (Wagner et al. 1999). In 1994 Joel Lau from the Nature Conservancy–Hawaii discovered a single plant at Wailupe, Oahu, growing precariously on a vertical rock face. Since then, eight plants have been found growing in the same general area. Immature fruits from a plant higher up from the main population were collected and aseptic seed culture successfully performed. Currently, approximately 20–30 seedlings derived from a single founder genotype are growing in the laboratory.
Schiedea adamantis, Caryophyllaceae This species was first collected on the slopes of Diamond Head Crater on Oahu by Charles Lamoureux and Earl Ozaki in 1955 and formally described in 1970. It is a small shrub known only from one population and has survived in the midst of this urban area largely because access to the site
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is limited by proximity to the Federal Aviation Administration facilities. S. adamantis exists under harsh conditions, buffeted by strong winds, high light intensity, low precipitation, and high temperatures. Since 1998, unusually prolonged drought conditions at Diamond Head have caused resulted in the collapse of the Schiedea population. Of an original 200 individuals, only two plants are known to be alive, although a few more may recover with increased precipitation. Fortunately, seed collections were made earlier by Dr. Steven Weller of the University of California–Irvine. Seedlings have been propagated in vitro for the Hawaii Army National Guard. Approximately 200 plants have been reintroduced on Diamond Head, where they are being maintained and monitored.
Cyanea longifolia, Campanulaceae C. longifolia, an endangered lobeliad, is known from three populations consisting of 120 plants. In the last 20 years this species has seen rapid decline, especially in the population along the Waianae Kai Trail on Oahu. It has become extremely rare in this area, with only two remaining individuals that are separated by a ridge. Extensive collections were made from both plants in 2000 by various concerned organizations such as the NTBG, CPCH, U.S. Army, and Lyon Arboretum. To date, more than 500 seedlings have been propagated in vitro from these two remaining plants. Approximately one-half of the seedlings have been sent to Hawaii’s State Department of Forestry and Wildlife Pahole midelevation nursery to be grown for future reintroduction projects.
Tetraplasandra flynnii, Araliaceae In 1998, this new species of Tetraplasandra was discovered by Ken Wood (NTBG) on the steep slopes of Kalalau of Kauai. Only one population of this species is known, consisting of only five mature trees approximately 5 m tall. Browsing goats are a constant threat, and as a result there is no evidence of regeneration. Immature fruit from all five trees was collected and sent to the Lyon Micropropagation Laboratory for ovule culture. So far, approximately 20 percent (25/125) of the immature embryos cultured have germinated and have been propagated for the purpose of germplasm storage and restoration work. Recently, one of the trees was found to have been deliberately and maliciously girdled. Attempts are being made to save this
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tree by bridge grafting over the wound using cambial tissue from Tetraplasandra kavaiensis, a related species.
Kokia cookei, Malvaceae K. cookei is considered one of the rarest and most threatened plant species in the world. It was discovered in the 1860s on the western end of Molokai by Mr. R. Meyer. This find consisted of three trees, which were not relocated on subsequent visits a few years later. In 1910, a single living tree was discovered in the general area of the initial sighting and may have been one of the original trees. In 1915, this last remaining wild specimen was found in extremely poor condition, but a few seeds were found and collected. K. cookei was extirpated from the wild in 1918. Seeds from this collection produced only one seedling that survived past 1933. This seedling was planted at a Kauluwai residence on Molokai and produced viable seed from the 1930s through the late 1950s. More than 130 seedlings were germinated and planted about Kauluwai, Molokai, in the Wai’anae Mountains and at Wa’ahila on Oahu, but none of these plants have persisted. In the late 1950s, the single plant at Kauluwai died, and the species was presumed extinct. In 1970, a single plant of the species was discovered at the Molokai residence, probably a relict of the previous cultivated plant. In 1978, a fire destroyed the last remaining rooted plant of K. cookei. Fortunately, before it was destroyed a branch was removed and later grafted onto a related species, Kokia kauaiensis, at the Waimea Arboretum (Oahu); as a result, K. cookei exists as approximately 23 grafted plants. K. cookei currently lacks viable seed production, with the last batch of viable seeds collected in 1974–1975. In 1993, two seedlings were produced through embryo culture but subsequently died, possibly because of poor growing conditions. In 2000 and 2001, six seedlings were germinated through embryo culture at Lyon Arboretum. Three of them are being grown in the greenhouse at Volcano Rare Plant Facilities (Volcano, Hawaii), and three are still in in vitro culture at the Lyon Arboretum Micropropagation Laboratory.
Conclusions These examples demonstrate that native species can be rescued from the brink of extinction, multiplied using various micropropagation techniques,
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and ultimately reintroduced in the wild when appropriate management of their habitats can be undertaken. To date, more than 300 Hawaiian plant species of which more than 130 are federally listed as Endangered or Threatened have been successfully grown at Lyon Arboretum using micropropagation techniques. The majority of these taxa from more than 14,000 accessions are being maintained as in vitro germplasm collections. For several taxa, the members of HRPRG have used specimens produced by micropropagation for restoration and reintroduction projects. The Lyon Arboretum Micropropagation Laboratory has successfully propagated 50 (82 percent) of the 61 GSNL species that have been submitted for propagation. To save the disappearing Hawaiian endemic flora, much more public and government commitment is needed. Limited funding for plant conservation continues to be problematic worldwide, and sustainable funding sources must be established for both existing and new programs (Toll 1999; Ashmore 1997). Many of the conservation programs in Hawaii, including the RHPP, are funded through species-specific projects required of federal agencies under the Endangered Species Act or receive only limited support, even though they have been successful in saving some of Hawaii’s most endangered and threatened species. The in vitro germplasm collection must be properly managed and accurately documented with the aid of a database that can be networked to collaborators and supporting organizations. Such a database, including plant logistics, propagation techniques and protocols, propagation and storage success rates, and in situ and ex situ plant inventory can be used to coordinate future research, collections, monitoring, propagation, and restoration work. The RHPP has just overhauled its micropropagation germplasm database. Various other Hawaii government and private conservation organizations are revising their databases in the hope that sometime soon, they can be networked. It is not an easy task because of differing mandates and needs. The tendency has always been to try to include too much information rather than a basic backbone on which interactions can be based. In vitro storage and propagation techniques will continue as important components of conservation strategies. In vitro conservation in Hawaii must be expanded to include long-term storage through cryopreservation, but a new infrastructure and long-term funding are needed. Collaborative research, information sharing, and technology transfer between conservation programs must be strengthened and new conservation programs estab-
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lished at national, regional, and international levels. Only through concerted efforts such as this can we optimize the limited resources available to conservation programs. References Ashmore, S. 1997. Status Report on the Development and Application of In Vitro Techniques for the Conservation and Use of Plant Genetic Resources. Rome: IPGRI. Bonga, J. M., and P. Von Aderkas. 1992. In Vitro Culture of Trees. Dordrecht, the Netherlands: Kluwer. Bridgen, M. 1994. A review of plant embryo culture. HortScience 29:1243–1246. Callow, J. A., B. V. Ford-Lloyd, and J. J. Newbury. 1997. Overview. Pages 1–7 in J. A. Callow, B. V. Ford-Lloyd, and H. J. Newbury (eds.), Biotechnology and Plant Genetic Resources: Conservation and Use. Wallingford, UK: CAB International. Center for Plant Conservation. 1999. Hawaii Task Force Species of Concern List. Unpublished report, December 3, 1999. Honolulu: Center for Plant Conservation. Coenen, C., and T. Lomax. 1997. Auxin-cytokinin interactions in higher plants: old problems and new tools. Trends in Plant Science 2(9):351–356. Cuddihy, L. W., and C. P. Stone. 1990. Alteration of Native Hawaiian Vegetation: Effects of Humans, Their Activities and Introductions. Honolulu: Cooperative National Park Resources Studies Unit, Hawaii. Dodds, J. H. 1991. Introduction: conservation of plant genetic resources—the need for tissue culture. Pages 1–9 in J. H. Dodds (ed.), In Vitro Methods for Conservation of Plant Genetic Resources. New York: Chapman & Hall. Engelmann, F. 2000. Importance of cryopreservation for the conservation of plant genetic resources. Pages 8–20 in F. Engelmann and H. Takagi (eds.), Cryopreservation of Tropical Plant Germplasm: Current Research Progress and Application. JIRCAS International Agriculture Series no. 8. Tsukuba: Japan International Research Center for Agricultural Sciences. Falk, D. A., and K. E. Holsinger. 1991. Genetics and Conservation of Rare Plants. Oxford, UK: Oxford University Press. George, E. 1993. Plant Propagation by Tissue Culture, Part 1, The Technology. Eversley, UK: Exegetics Ltd. Gon, S., and D. Matsuwaki. 1999. Hawaii Natural Heritage Program. Native Ecosystems before Human Settlement and Remaining Native Ecosystem Today. GIS Ecosystem Data Layers of the State of Hawaii. Honolulu, Hawaii: U.S. Fish and Wildlife Service. Grierson, M., and P. S. Green. 1996. A Hawaiian Florilegium. Honolulu: University of Hawaii Press. Guarino, L., V. Ramanatha Rao, and R. Reid. 1995. Collecting Plant Genetic Diversity: Technical Guidelines. Wallingford, UK: CAB International. Hartmann, H., D. Kester, and F. Davies. 1997. Plant Propagation Principles and Practices. 6th edition. Englewood Cliffs, NJ: Prentice Hall. HRPRG (Hawaii Rare Plant Restoration Group). 1999a. Hawaii Rare Plant Genetic Safety Net Initiative. Unpublished report prepared May 1999, Honolulu. HRPRG (Hawaii Rare Plant Restoration Group). 1999b. Protocol for Monitoring and Collecting from Rare Plant Populations. Unpublished report, Honolulu. IBPGR (International Board for Plant Genetic Resources). 1984. The Potential for
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Using In Vitro Techniques for Germplasm Collection. Rome: International Board for Plant Genetic Resources. Jackson, G. 1999. Taro leaf blight. Pest Advisory Leaflet no. 3. Suwa, Fiji: Plant Protection Service, Secretariat of the Pacific Community. Knudson, L. 1946. A new nutrient solution for the germination of orchid seeds. American Orchid Society Bulletin 15:214–217. Kyte, L., and J. Kleyn. 1996. Plants from Test Tubes: An Introduction to Micropropagation. 3rd edition. Portland, OR: Timber Press. Lloyd, G., and B. McCown. 1980. Commercially feasible micropropagation of mountain laurel, Kalmia latifolia, by use of shoot-tip culture. Proceedings of the International Plant Propagators’ Society 30:421–427. Murashige, T., and F. Skoog. 1962. A revised medium for rapid growth and bioassays with tobacco tissue cultures. Physiologia Plantarum 15:473–497. Ng, S. Y. C., and N. Q. Ng. 1991. Reduced-growth storage of germplasm. Pages 11–40 in J. H. Dodds (ed.), In Vitro Methods for Conservation of Plant Genetic Resources. New York: Chapman & Hall. Pierik, R. L. M. 1987. In Vitro Culture of Higher Plants. Dordrecht, the Netherlands: Martinus Nijhoff. Preece, J., and E. Sutter. 1991. Acclimatization of micropropagated plants to the greenhouse and field. Pages 71–98 in P. C. Debergh and R. H. Zimmerman (eds.), Micropropagation: Technology and Application. Dordrecht, the Netherlands: Kluwer. Ronchi, V. 1995. Mitosis and meiosis in cultured plant cells and their relationship to variant cell types arising in culture. Pages 65–129 in K.W. Jeon and J. Jarvik (eds.), International Review of Cytology - A Survey of Cell Biology, Vol. 158. London: Academic Press. Sakai, A. 2000. Development of cryopreservation techniques. Pages 1–7 in F. Engelmann and H. Takagi (eds.), Cryopreservation of Tropical Plant Germplasm: Current Research Progress and Application. JIRCAS International Agriculture Series no. 8. Tsukuba: Japan International Research Center for Agricultural Sciences. Scowcroft, W. 1984. Genetic Variability in Tissue Culture: Impact on Germplasm Conservation and Utilization. Rome: International Board for Plant Genetic Resources. Skirvin, R., K. McPheeters, and M. Norton. 1994. Sources and frequency of somaclonal variation. HortScience 29(11):1232–1237. Sohmer, S. H., and R. Gustafson. 1996. Plants and Flowers of Hawaii. Honolulu: University of Hawaii Press. Sugii, N., and C. Lamoureux. 1998. Micropropagation: an important tool in the conservation of endangered Hawaiian plants. Pages 43–48 in R. Rose and D. Haase (eds.), Native Plants Propagation and Planting: Symposium Proceedings. Corvallis: Oregon State University. Toll, J. 1999. Field gene bank management: problems and potential solutions. Pages 63–69 in F. Engelmann (ed.), Management of Field and In Vitro Germplasm Collections. Rome: International Plant Genetic Resources Institute. U.S. Fish and Wildlife Service. 1998. Recovery Plan for Kokia cookei. Portland, OR: U.S. Fish and Wildlife Service. U.S. Fish and Wildlife Service. 1999. Endangered and threatened wildlife and plants as of December 31, 1999. Federal Register 50 CFR:17.11–17.12. Villordon, A., and D. Labonte. 1996. Genetic variation among sweet potatoes propa-
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gated through nodal and adventitious sprouts. Journal of the American Society for Horticultural Science 121:170–174. Vuylsteke, D. 1989. Shoot-tip culture for the propagation, conservation and exchange of Musa germplasm. Pages 1–55 in Practical Manuals for Handling Crop Germplasm In Vitro 2. Rome: International Board for Plant Genetic Resources. Wagner, W. L., D. R. Herbst, and S. H. Sohmer. 1999. Manual of the Flowering Plants of Hawaii. Revised edition. Bishop Museum Special Publication 97. Honolulu: University of Hawaii Press. Wester, L. 1992. Origin and distribution of the adventive alien flowering plants in Hawaii. Pages 99–154 in C. P. Stone, C. W. Smith, and J. T. Tunison (eds.), Alien Plant Invasions in Native Ecosystems of Hawaii: Management and Research. Honolulu: University of Hawaii Cooperative National Park Resources Studies Unit. Withers, L. A. 1985. Long-term storage of in vitro cultures. Pages 137–148 in A. Schafer-Menuhr (ed.), In Vitro Techniques: Propagation and Long Term Storage. Dordrecht, the Netherlands: Martinus Nijhoff.
Chapter 11
Ex Situ Conservation Methods for Bryophytes and Pteridophytes Valerie C. Pence
Like seed-bearing plants, bryophytes and pteridophytes are threatened by habitat loss, pollution, overcollection, invasive species, and other factors. The International Union for the Conservation of Nature (IUCN) Red List (Walter and Gillett 1998) lists 770 species of pteridophytes (ferns and fern allies) that are threatened worldwide (Table 11.1). Bryophytes (mosses, liverworts, and hornworts) are less well documented, but 92 species have been listed by the IUCN in their 2000 Red List of Bryophytes (http://www.art data.slu.se/guest/SSCBryo/WorldBryo). Several species of bryophytes (e.g., Neomacounia nitida, Orthotrichum truncato-dentatum, and Dactylolejeunea acanthifolia) have not been found in the wild for many years, and because their original habitat is gone, they are thought to be extinct. Similarly, several ferns are thought to be extinct or are extinct in the wild (e.g., Tectaria amesiana, Thelypteris altissima, and Diplazium laffanianum). Whereas in situ conservation should be the primary focus for conserving these threatened species (Hallingbäck and Tan 1996; Wagner 1995), ex situ growth and germplasm storage can also be important complementary aspects of plant conservation (Given 1987; Laliberté 1997). Species preserved ex situ could serve as an important resource for research on reproduction and growth, on the habitat needs of the species, and for recovery programs. There has been less focus on bryophytes and pteridophytes than on seed plants by botanical gardens and other conservation organizations, perhaps because they are often less charismatic or less well known than seed-bearing taxa, and their identification requires specialized expertise. Several efforts are under way to target the needs of these groups, including the IUCN 206
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table 11.1 Threatened taxa of seed and nonseed plants. Taxa
Bryophytes Fern allies True ferns Total pteridophytes Total seed plants
Total No. of Speciesa
No. Listedb
8,000–12,000 1,318 9,053 10,371 231,461
92 87 683 770 32,498
a Bryophytes from Bold et al. (1980); pteridophytes and seed plants from Walter and Gillett (1998). b Bryophytes from http://www.artdata.slu.se/guest/SSCBryo/WorldBryo; pteridophytes and seed plants from Walter and Gillett (1998).
Bryophyte Specialist Group, the IUCN Pteridophyte Specialist Group, the European Committee for Conservation of Bryophytes (Söderström 1998), and a 3-year pilot project on ex situ conservation of bryophytes at the Micropropagation Unit, Royal Botanic Gardens, Kew. Several approaches can be taken for ex situ conservation of bryophytes and pteridophytes, and because of their different life forms and adaptations, these taxa offer more options for ex situ conservation than seed plants (Table 11.2).
Horticultural Collections Traditionally, the ex situ conservation of threatened species has taken place in botanical gardens. Ex situ cultivation has meant survival for a number of species of seed plants that have become extinct in the wild, such as Sophora toromiro (Maunder et al. 2000) and Franklinia alatamaha (Lucas and Synge 1978). Of the nonseed taxa, ferns have been the most well represented in horticultural collections, and in some cases they have been the subject of dedicated programs or exhibits (Pattison 1992; Page et al. 1992; Theuerkauf 1993). Bryophytes are also often found among other plants in public displays, but they have not been widely used as exhibit subjects, although there are a few notable exceptions (e.g., the Cryptogamic Garden at the Royal Botanic Gardens, Edinburgh). In addition, there are a few noteworthy private living collections (Longton, in press). Some bryophyte species grow easily in greenhouse culture, but others need specialized conditions (Schenk 1997). In addition, because some species can be invasive, maintaining pure species lines in the greenhouse takes a particular vigilance beyond that needed for most vascular plants.
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table 11.2 Summary of ex situ conservation methods used with bryophytes and pteridophytes. Method
Group and Generation
Horticultural collections Spore banking Vegetative tissues In vitro collections In vitro propagation Vegetative tissue banking Field collecting In vitro collecting Ex vitro collecting
B(g), P(s) B, P B(g), P(s), P(g) B(g), P(s), P(g) B(g), P(s), P(g) P(s) B(g)
B, bryophytes; g, gametophytes; P, pteridophytes; s, sporophytes.
Spore Banking Another traditional method of germplasm preservation has been seed and spore banking. Whereas seed banking has been increasingly practiced in both botanical gardens and agricultural institutions (Laliberté 1997; Committee on Managing Global Genetic Resources 1991), there has been only limited institutional storage of spores. A few botanical gardens store fern spores, and some spore storage takes place in spore exchanges, but there has been less systematic medium- or long-term banking for germplasm preservation than with seeds. Like seeds, however, spores of many species lend themselves well to long-term germplasm storage. In mosses, spores are born in capsules on conspicuous sporophytes for several weeks each year. These can be removed from the plants and stored intact. The spore capsules easily withstand surface sterilization, and if they are collected before they open, they provide a convenient method for obtaining sterile spores for subsequent aseptic culture. In the case of pteridophytes, placing a frond in a paper envelope and letting it air dry often releases large numbers of sporangia, which can be easily collected from the envelope and placed into storage. Additional spores can be collected by scraping sporangia from the frond, and this method can also be used to collect very fresh spores before drying. The methods used for spore banking mirror those of seed banking. Seeds that can be dried generally remain dormant for long periods of time, and storage at low temperatures increases longevity (Ellis and Roberts 1980). Similarly, the spores of many pteridophytes can also remain dor-
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mant for up to several years when they are dried, and longevity can be improved by reducing storage temperature (Lloyd and Klekowski 1970; Windham et al. 1986; Kiss and Kiss 1998). Like seeds, however, they eventually lose viability in storage (Beri and Bir 1993; Camloh 1999). The spores of pteridophytes often are classified as chlorophyllous spores, which generally lose viability in a matter of weeks, or nonchlorophyllous spores, which may remain viable for several years. This characteristic has allowed the recovery of plants from nonchlorophyllous spores obtained from dried herbarium specimens (Windham et al. 1986). However, some nonchlorophyllous spores, such as those from the family Cyatheaceae, may maintain viability for only a few weeks (Page 1979). Lower temperatures have been shown to improve survival, as with spores of the tree fern Cyathea delgadii (Simabukuro et al. 1998). Spores of the endangered tree ferns Cyathea spinulosa and Dicksonia sellowiana have also been successfully grown after cryopreservation, or storage in liquid nitrogen (LN; Agrawal et al. 1993; Rogge et al. 2000). Dried nonchlorophyllous spores of several fern species from other families have germinated well after 4.5 or 6 years of storage in LN (Pence 2000b). On the other hand, a number of unrelated genera of ferns, as well as the Equisetaceae, have chlorophyllous spores. Most of these species inhabit wet, mesophytic sites, and the spores germinate soon after being shed. If they do not germinate, they generally do not survive more than a few weeks (Stokey 1951; Lloyd and Klekowski 1970; Pérez-García et al. 1994; Lebkuecher 1997), although there appear to be exceptions to this (Dyer and Lindsay 1996). As with nonchlorophyllous spores, however, storage at lower temperatures (4°C and –70°C) has increased longevity (Lloyd and Klekowski 1970; Whittier 1996). Chlorophyllous spores from Onoclea sensibilis and Osmunda regalis have survived drying and exposure to LN, and spores of the latter have remained viable for at least 18 months under these conditions (Pence 2000b). These spores survived LN exposure using either the open drying technique or the encapsulation dehydration procedure (described later in this chapter), but preservation methods were successful only if the spores were freshly harvested. These results suggest that chlorophyllous spores may resemble suborthodox seeds, which are generally short-lived but can be dried and frozen successfully if fresh seeds are used (Bonner 1986). Bryophyte spore banking has not been widely practiced, possibly because of the relative ease with which gametophyte fragments of mosses and liverworts can be transported and grown. However, bryophyte spores
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are similar to fern spores in that, depending on the species, they may have chlorophyllous or nonchlorophyllous spores (Fulford 1956; Morgensen 1978; Schuster 1983). In this laboratory, storage of nonchlorophyllous spores of several bryophyte species in LN has been successful. Spores were surface sterilized after thawing and successfully germinated after sowing on a half-strength Murashige and Skoog (1962; MS) medium with 1.5 percent sucrose. Further research is needed to determine whether these techniques can be transferred to other bryophyte species, including those with chlorophyllous spores. Spore banking, though not yet widespread, holds the potential for longterm ex situ germplasm storage. Even more compact than seeds, spores can be used to maintain a broad range of genetic material in a small space with minimal input of time and labor. Storage at reduced temperatures is needed to maximize longevity, and both chlorophyllous and nonchlorophyllous spores of pteridophytes and at least nonchlorophyllous spores of bryophytes appear to be adaptable to low temperature, including LN storage. Further research is needed to confirm the applicability of the technique to a wider range of species, but the results thus far suggest that it will be an important tool for maintaining threatened and other important germplasm of both pteridophytes and bryophytes.
Vegetative Tissues When spores of pteridophytes or bryophytes are not available, vegetative tissues can be maintained or preserved for ex situ conservation using tissue culture or cryopreservation techniques. In the case of pteridophytes, where the sporophyte and the gametophyte are free-living, separate approaches have been developed for each. In the case of bryophytes, where the sporophyte is dependent on the gametophyte, work has focused on the growth and preservation of gametophytic tissues.
In Vitro Propagation With both ferns and seeds plants, in vitro techniques have been used for plant propagation, particularly for commercial horticultural or agricultural purposes (Fernández and Revilla 2003). In vitro propagation also has a role in the ex situ management of threatened species. Many threatened species are not abundant enough to warrant the translocation of whole plants from the wild, so some propagation method must be used to establish the ex situ
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population. When traditional methods of seed or spore germination or clonal propagation by cuttings or offshoots are not adequate, in vitro methods may provide an alternative method for propagation. A number of threatened species of seed plants have been propagated through tissue culture to provide materials for ex situ populations, research, and reintroduction (Fay 1992; Pence 1999). Because in vitro propagation can provide only clonal material, care must be taken when dealing with threatened taxa to maintain as much diversity as possible through the propagation process. Each genetic line that is initiated should be propagated and maintained separately to avoid any in vitro selection for hardier or faster-growing material. Similarly, the possibility of somaclonal variation should be kept in mind when dealing with the propagation of threatened plants. Genetic change in cultured materials is generally rare, although the rate is determined by the species, culture conditions, time in culture, and other factors (Karp 1994). However, it has been suggested that for species with an extremely limited genetic base, somaclonal variation may be of benefit as a source of genetic variability (Jacobsen and Dohmen 1990). Techniques for propagating a number of species of ferns and fern allies have been reported (Table 11.3), and these techniques should be adaptable to related threatened species of pteridophytes. When only a few spores are available, in vitro germination, followed by sporophyte production and micropropagation of the sporophyte, could be used to increase the numbers of individuals for research or reintroduction. To increase numbers further, in vitro germination of spores could be followed by maceration of the gametophytes, which in some species has led to development of more gametophyte tissue from the fragments (Knauss 1976). When spores are not available, carefully excised buds might be used to initiate shoot cultures, again to increase numbers (Dykeman and Cumming 1985). With some ferns, also, shoots can be regenerated from runner tips or pieces of sporophyte leaves (Hirsch 1975; Pais and Casal 1987; Borgan and Naess 1987). The production of new sporophytes from fragments of macerated sporophytic tissue has also been reported (Cooke 1979; Fernández et al. 1991). There is less information on the propagation of fern allies in vitro, although in this lab, Selaginella species have multiplied well on a half-strength MS medium (Pence 2001). Methods for the in vitro culture of bryophyte gametophytes are well established. These methods can be used to propagate threatened species for research and possible reintroduction into suitable habitats (Longton, in press).
table 11.3 Some examples of pteridophytes that have been grown and propagated in vitro. Species
Filicopsida Aspleniaceae Asplenium nidus L.; Asplenium nidusavis L. Phyllitis scolopendrium (L.) Newman Blechnaceae Blechnum brasiliense Desv.; B. gibbum (Labil.) Mett.; B. punctulatum Sw.; B. spicant (L.) With. Woodwardia virginica (L.) Sm. Cyatheaceae Cyathea gigentea (Wall. ex Hook.) Holttum; C. australis (R. Br.) Domin Davalliaceae Davallia fejeensis Hook. Dennstaedtiaceae Pteridium aquilinum (L.) Kuhn Dryopteridaceae Dryopteris filixmas (L.) Schott.; D. affinis (Lowe) Fraser-Jenk. Matteuccia struthiopteris (L.) Todaro Phanerophlebia falcatum Polystichum falcata (L.f.) Diels Rumohra adiantiformis (G. Forst.) Ching Gleicheniaceae Dicranopteris linearis (Burm. f.) Underw. Hymenophyllaceae Trichomanes speciosum Willd. Marsiliaceae Marsilea quadrifolia L. Pilularia globulifera L. Nephrolepidaceae Nephrolepis cordifolia (L.) Presl.; N. exaltata (L.) Schott.; N. exaltata “Bostoniensis”; N. exaltata “Scottii”; N. falcata (Cav.) C. Chr.; N. multiflora (Roxb.) Jarret ex Morton; Nephrolepis Schott sp.
References
Higuchi and Amaki 1989; Amaki and Higuchi 1991; Ferna´ndez et al. 1991, 1993 Zenkteler 1993 Janssens and Sepelie 1989; Ferna´ndez et al. 1996 Ferna´ndez et al. 1999 Padhya 1987; Goller and Rybczynski 1995 Cooke 1979 Sheffield et al. 1997 Breznovits and Mohay 1987; Ferna´ndez et al. 1999 Dykeman and Cumming 1985; Hicks and von Aderkas 1986; Thakur et al. 1998; Materi et al. 1995 De Garcia and Furelli 1987 Amaki and Higuchi 1991; Stamps 1992 Henson 1979 Raine and Sheffield 1997 Breznovits and Mohay 1987 Breznovits and Mohay 1987 Sulklyan and Mehra 1977; Loescher and Albrecht 1978; Henson 1979; Petersen 1979; Soede 1981; Leffring and Soede 1982; Beck and Caponetti 1983; Paek et al. 1984; Higuchi et al. 1987; Breznovits and Mohay 1987; Padhya 1987; Borgan and Naess 1987; Amaki and Higuchi 1991; Sara et al. 1998
table 11.3 (continued) Some examples of pteridophytes that have been grown and propagated in vitro. Species
Osmundaceae Osmunda cinnamonea L.; O. japonica Thunb.; O. regalis L. Todea barbara (L.) T. Moore Polypodiaceae Microgramma vacciniifolia (Langsd. and Fisch.) Copel. Platycerium bifurcatum (Cav.) C. Chr.; P. coronarium (Mull.) Desv.; P. stemmaria (P. Beauv.) Desv.
Polypodium cambricum L.; P. vulgare L. Pteridaceae Adiantum capillus-veneris L.; A. cuneatum G. Forst; A. pedatum L.; A. raddianum C. Pred.; A. trapeziforme L. Ceratopteris thalictroides (L.) Brongn. Cheilanthes alabamensis (Buckl.) Kunze; C. tomentosa Link. Notholaena R. Br. “Sun-tuff” Pellaea rotundifolia (Forst.f.) Hook. Pteris cretica L.; P. ensiformis Burm.; P. henryi H. Christ; P. vittata L. Schizaeaceae Anemia phyllitidis (L.) Sw. Thelypteridaceae Ampelopteris prolifera (Retz.) Copel. Cyclosorus contiguous; C. dentatus (Forssk.) Ching Lycopsida Lycopodiaceae Lycopodiella inundata (L.) Holub Lycopodium cernuum L. Selaginellaceae Selaginella muscosa Spring; S. uncinata (Desv. ex Poir.) Spring; S. willdenovii (Desv. ex Poir) Bak. Salviniaceae Salvinia natans (L.) All.
References
Whittier and Steeves 1960; Breznovits and Mohay 1987; Ferna´ndez et al. 1999; Kawakami et al. 1999; Morini 2000 DeMaggio and Wetmore 1961 Hirsch 1975 Hennen and Sheehan 1978; Cooke 1979; Henson 1979; Camloh and Gogala 1991; Camloh et al. 1994; Kwa et al. 1995, 1997; Kim et al. 1996; Ambrozic-Dolinsek and Camloh 1997; Teng and Teng 1997; Camloh et al. 1999 Bertrand et al. 1999; Zenkteler 1991 Wetmore 1954; Whittier and Steeves 1960; Murashige 1974; Pais and Casal 1987; Amaki and Higuchi 1991; Padhya 1995 Cheema and Sharma 1993, 1994 Whittier 1965 Rogers and Banister 1992 Janssens and Sepelie 1989 Kshirsagar and Mehta 1978; Padhya 1987; Breznovits and Mohay 1987; Amaki and Higuchi 1991; Ferna´ndez et al. 1996, 1999 Sheffield et al. 1997 Mehra and Sulklyan 1969 Mehra and Palta 1971; Breznovits and Mohay 1987 Atmane 1999 Wetmore 1954 Wetmore 1954; Pence 2001
Nakamura and Maeda 1994
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Whether seed or nonseed plants, propagation of a threatened species in vitro is limited by the availability of workable techniques for that species. Some species are “recalcitrant” in culture and are not easily grown. In addition, experimentation on threatened species is limited by the amount of tissue available. With seed plants, one approach has been to develop techniques with similar or related nonthreatened species before attempting propagation on the target species. Such an approach can also be taken with pteridophytes and bryophytes, allowing work with nonthreatened species to benefit work with threatened taxa.
In Vitro Ex Situ Collections Because of the difficulty in maintaining horticultural collections, it is not surprising that tissue culture methods have been used for maintaining bryophyte species (Lal 1984; Sargent 1988). Once established, in vitro lines of gametophytes can be kept in a minimum of space and can avoid problems of weediness and cross-contamination of species that can occur in horticultural collections. Sporophytes of bryophytes have also been grown in vitro, but this has been primarily for experimental purposes (Raudzens and Matzke 1968). Many bryophyte gametophytes grow well in vitro on a simple medium, although species with specialized habitat requirements may need particular modifications (Duckett et al., in press). Cultures usually are initiated from spores obtained from surface-sterilized spore capsules. It is also possible to surface sterilize gametophytic tissue directly, although the sensitivity of bryophyte tissues generally necessitates precise timing to avoid killing the tissue during sterilization. Fern gametophytes have also been grown and maintained in vitro, particularly for experimental purposes (Miller 1968; Dyer 1979). They can be initiated either by germinating spores aseptically or by surface sterilizing a fern gametophyte directly and placing it on a tissue culture medium (Ford and Fay 1990; Ford 1992). Most cultures have been grown on a simple agar solidified medium, although methods for liquid culture have also been described (Sheffield et al. 1997). Because the gametophyte can reproduce vegetatively, these lines can be maintained indefinitely. Some fern taxa are known only from the gametophyte in certain geographic areas (Farrar 1967), and these techniques could be used to maintain ex situ collections of fern germplasm of this type. For fern sporophytes, in vitro prop-
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agation techniques have been developed for a number of species, and these methods could be adapted for maintaining collections of fern sporophytes in vitro and possibly for slow growth storage. For example, cultures of Nephrolepis can be stored for at least 36 months when kept at 9°C in the dark (Hvoslef-Eide 1990).
Vegetative Tissue Banking Maintaining tissues in vitro entails a high input of labor and facilities, and there is the possibility that they may undergo genetic change with increased time in culture. In addition, tissue culture lines that are used to propagate threatened taxa should be preserved for future use. To address these needs, methods for stable vegetative tissue banking are being developed for both bryophytes and pteridophytes. These techniques generally involve storage in LN to maintain the viability of the tissues.
Bryophyte Gametophytes Several techniques have been developed for cryopreserving bryophytes, generally directed at preserving in vitro grown tissues. In early studies, some survival was observed in moss cultures that were stored on agar at –15°C or –20°C for from several days to several years (Longton and Holdgate 1967; Longton 1981). Protoplasts of the liverwort Marchantia polymorpha and protonemal cultures of the moss Physcomitrella patens have both been successfully recovered after exposure to LN using a slow freezing method with cryoprotectants (Takeuchi et al. 1980; Grimsley and Withers 1983). More recent studies have explored techniques that avoid the use of programmable freezers. One of these methods has been called open drying, or drying without physical or chemical protectants. This procedure is particularly useful with tissues that show a degree of natural desiccation tolerance, as do many moss and some liverwort gametophytes (Oliver 1996). In some cases, pretreatment with the plant stress hormone abscisic acid (ABA) is needed to induce desiccation tolerance (Hellwege et al. 1996). If excised growing tips of in vitro–grown Riccia fluitans are dried for an hour under the airflow of a laminar flow hood, the tissue is killed. However, if R. fluitans is grown for as little as a day on medium supplemented with 10 µM ABA and then similarly dried, there is 100 percent survival when the tissues are placed back onto growth medium (Table 11.4; Pence 1998; Plair
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1998). This behavior is not uniformly distributed among bryophytes; for instance, Marchantia polymorpha had poor recovery growth with or without ABA preculture (Pence 1998). Another method, which has been developed for the cryopreservation of in vitro–grown gametophytes, is a slow freezing method that includes an ABA and proline pretreatment (Christianson 1998). The tissues are precultured for 3 or 4 days on 10 µM ABA and 100 mM proline. They are then exposed to a cryoprotectant solution for 1 hour and transferred to a passive slow freezer in which the samples are frozen slowly to –80°C. They are then transferred to LN. When needed, the samples are thawed rapidly and the tissues placed back into culture for growth. The encapsulation dehydration technique of Fabre and Dereuddre (1990) has also been used to successfully cryopreserve several bryophyte species grown in vitro. With this technique, shoot tips are suspended in a 3 percent solution of alginic acid. The solution with the tissue is pulled into a pipette and then dropped into a solution of calcium chloride. The calcium causes the drops of alginic acid to gel, forming beads, each of which contains one or more pieces of tissue. The encapsulated tissues are incubated for 18 hours in a solution containing 0.75 M sucrose on a rotary shaker overnight. The next day, the beads are removed from the solution and placed under the airflow of a laminar flow hood to dry for 4 hours. They are then placed into a cryovial and immersed directly into LN. When the beads are removed from LN, they are thawed on the benchtop for 20 minutes, removed from the cryovial, and placed on growth medium, where the tissues rehydrate and resume growth. This procedure resulted in good survival with the same species that were tested for survival through open drying (Table 11.4; Pence 1998). With R. fluitans, there was 100 percent survival with the encapsulation dehydration procedure, with or without ABA preculture. There was also good survival with other species, although M. polymorpha survived only if precultured on ABA. Most mosses and some liverworts possess a degree of desiccation tolerance, and these species may also be adaptable to cryopreservation ex vitro, that is, without first establishing in vitro cultures. Because of the damaging effects of ice crystals on living tissues, cryopreservation protocols depend largely on the removal of tissue water, either by direct desiccation or by removal or replacement of water by chemical means. Thirteen species of temperate mosses collected in the Cincinnati area were washed of soil,
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table 11.4 Survival of gametophytes of several bryophytes and ferns through liquid nitrogen exposure, using open drying and encapsulation dehydration procedures with and without preculture on 10-M abscisic acid (ABA). Percentage Survival Open Drying Species
Bryophytes Riccia fluitans Marchantia polymorpha Plagiochila sp. Helicondontium cappelare Pteridophytes Cibotium glaucum Adiantum tenerum Drynaria quercifolia Davallia fejeensis Polypodium aureum Adiantum trapeziforme
Encapsulation Dehydration
No ABA
With ABA
No ABA
With ABA
0 0 0 25
100 10 0 80
100 10 100 81
100 100 100 83
10 60 62 40 0 8
88 100 71 76 16 54
100 94 100 100 84 93
100 100 100 100 100 100
Sources: Pence (1998, 2000a).
blotted dry, dried under the airflow of the laminar flow hood for 3 hours, and then exposed to LN storage for at least 1 hour. They were thawed at room temperature and then placed back on soil, where all resumed growth. ABA pretreatment was tested but found not to be necessary (Leverone and Pence 1993, cited in Pence and Christianson, in press). The additional ability of bryophytes to regenerate from fragments is also valuable for cryopreservation protocols. If a portion of the tissue is damaged in the freezing process, undamaged tissues have the ability to regenerate the specimen. Thus, cryopreservation appears to have significant potential for germplasm storage of bryophyte species both in vitro and ex vitro (Burch and Wilkinson 2002). At least one extensive collection of bryophyte species that has been maintained in vitro for a number of years is being systematically cryopreserved for long-term banking (Christianson in Pence and Christianson, in press).
Pteridophyte Gametophytes Although the full extent of desiccation tolerance in fern gametophytes is not known, tolerance in some species has been reported (Mottier 1914;
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Page 1979). For those species, open drying may be useful, as with some bryophyte species. Open drying has resulted in some success in recovering in vitro–grown fern gametophytes after freezing (Table 11.4; Pence 2000a), and this appears to be improved with preculture on ABA. However, the encapsulation dehydration procedure has been even more successful, providing up to 100 percent survival in the gametophytes of the species tested, with or without ABA preculture. Fern gametophyte tissues of the species listed in Table 11.4 and prepared by encapsulation dehydration have also been stored successfully for at least 3.5 years in LN.
Pteridophyte Sporophytes The sporophytes of some ferns and fern allies, those that are “resurrection” species, are adapted for surviving desiccation (see Proctor and Pence 2002 for species list). As with desiccation-tolerant gametophytes, dried tissues from such plants should also be adaptable to cryostorage. In addition, several methods have been developed for cryopreserving shoot tips from in vitro cultures of seed plants and should be applicable to the shoot tips of in vitro–grown pteridophytes as well. Techniques include a two-step slow freezing (Withers 1985), vitrification (Sakai et al. 1990), encapsulation dehydration (Fabre and Dereuddre 1990), and encapsulation vitrification (Tannoury et al. 1991), and they have been successful with a wide range of seed-bearing taxa, both temperate and tropical. Shoot tips of in vitro–grown Selaginella uncinata can survive LN storage if they are subjected to preculture on ABA followed by the encapsulation dehydration protocol. They have also been recovered and regrown after 18 months of LN storage (Pence 2001). Preliminary work in this laboratory has also demonstrated survival of shoot tips of the fern Adiantum tenerum through freezing. It is likely that shoot tips of in vitro cultures of sporophytes of other ferns and fern allies should be adaptable to cryostorage using one of the several methods available.
Field Collecting An important part of ex situ conservation is collecting the materials to be conserved. With pteridophytes, spores or whole sporophytes generally are collected, whereas with bryophytes, spores or gametophytes can be used to initiate growth ex situ. When spores are not available, some nontraditional
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techniques are being developed to maximize the success and efficiency of collecting vegetative tissues.
In Vitro Collecting In vitro collecting (IVC) is the initiation of tissue cultures in the field. It has been used to collect germplasm from selected economically important species (Withers 1995) and from several threatened taxa (Pence 1999; Clark and Pence 1999; Pence et al. 2002). It can be useful when seeds or spores are not available because they are not being produced by the plant adequately or they are not available at the time a collection is made. For IVC, the tissue is surface sterilized in the field with alcohol or other sterilants and placed into a small container with tissue culture medium. This allows for the initiation of the culture using very fresh tissue and for the efficient transport of many samples in a small amount of space. Many field-collected materials contain endogenous contaminating organisms, necessitating the use of antibiotics or fungicides in the medium (Pence and Sandoval 2002). These techniques have been used successfully with a number of species, both temperate and tropical (Pence 1996), including the pteridophyte Selaginella sylvestris from the Costa Rican rainforest. By applying media and methods that have been successful with the tissue culture propagation of ferns and fern allies, IVC could be a useful tool for collecting germplasm from pteridophyte sporophytes for ex situ conservation. Preliminary results suggest that it may also be possible to collect gametophytes of ferns and bryophytes by IVC as well.
Field Collecting Bryophytes Because many bryophytes, particularly mosses, have some desiccation tolerance and can survive drying and LN exposure, these methods could also be used in collecting bryophyte germplasm. Bryophytes of several species have been dried and frozen in LN in the field in Trinidad and Costa Rica. They were transported to Cincinnati as frozen samples, stored in LN, and then thawed and placed onto soil in baby food jars. In a number of cases, the bryophytes survived—or at least a portion of the tissue survived—and resumed growth on soil. Further research is needed in this area, but the results suggest that many bryophytes should be adaptable to ex vitro field drying and freezing.
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Conclusions A variety of techniques are available for the ex situ conservation of nonseed plants. Traditional methods of horticultural cultivation and spore banking involve little equipment and could be used more extensively to maintain collections of bryophyte and pteridophyte germplasm. These could be supplemented by methods for in vitro growth and propagation and by desiccated or cryoprotected freezing of both spores and vegetative tissues. These techniques have been widely used with seed plants, and protocols are available that have been adapted to nonseed species. Finally, when spores are not available, techniques for collecting and preserving vegetative tissues in the field can be used to extend ex situ conservation methods to a wider range of taxa. Taken together, these methods comprise an effective collection of tools for the propagation and preservation of pteridophytes and bryophytes. Although there is less systematic conservation of these species than of seedbearing plants, there is a need to preserve the many endangered bryophytes and pteridophyte taxa that have been documented. Fortunately, efforts are beginning in this direction (e.g., Söderström 1998; Jackson, in press). The demonstration of the effectiveness of the many techniques available should facilitate the development of ex situ conservation programs for threatened species of both bryophytes and pteridophytes. References Agrawal, D. C., S. S. Pawar, and A. F. Mascarenhas. 1993. Cryopreservation of spores of Cyathea spinulosa Wall. ex Hook. f., an endangered tree fern. Journal of Plant Physiology 142:124–126. Amaki, W., and H. Higuchi. 1991. A possible propagation system of Nephrolepis, Asplenium, Pteris, Adiantum and Rumohra (Arachniodes) through tissue culture. Acta Horticulturae 300:237–243. Ambrozic-Dolinsek, J. A., and M. Camloh. 1997. Gametophytic and sporophytic regeneration from bud scales of the fern Platycerium bifurcatum (Cav.) C. Chr. in vitro. Annals of Botany 80:23–28. Atmane, N. 1999. Multiplication d’une Lycopodiale medicinale menacée de disparition [Lycopodiella inundata (L.) Holub] par les techniques de culture in vitro et interets pour ses alcaloides endogènes. Ph.D. thesis, University of Sciences and Technology of Lille, France. Beck, M. J., and J. D. Caponetti. 1983. The effects of kinetin and naphthaleneacetic acid on in vitro shoot multiplication and rooting in the fishtail fern. American Journal of Botany 70:1–7. Beri, A., and S. S. Bir. 1993. Germination of stored spores of Pteris vittata L. American Fern Journal 83:73–78.
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Pence, V. C., and J. Sandoval. 2002. Controlling contamination during in vitro collecting. Pages 30–40 in V. C. Pence, J. Sandoval, V. Villalobos, and F. Engelmann (eds.), In Vitro Collecting Techniques for Germplasm Conservation. IPGRI Technical Bulletin No. 7. Rome: International Plant Genetic Resources Institute. Pérez-García, B., A. Orozco-Segovia, and R. Riba. 1994. The effects of white fluorescent light, far-red light, darkness, and moisture on the spore germination of Lygodium heterodoxum (Schizaceae). American Journal of Botany 8:1367–1369. Petersen, M. A. 1979. Some aspects of nursery production in Queensland. Combined Proceedings of the International Propagators Society 29:108–110. Plair, B. L. 1998. The Effect of Abscisic Acid on Desiccation Tolerance and Soluble Sugars in the Liverwort Riccia fluitans L. Master’s thesis, University of Cincinnati, Ohio. Proctor, M. C. F., and V. C. Pence. 2002. Vegetative tissues: bryophytes, vascular “resurrection plants” and vegetative propagules. Pages 207–237 in M. Black and H. Pritchard (eds.), Desiccation and Plant Survival. Oxford, UK: CAB International. Raine, C. A., and E. Sheffield. 1997. Establishment and maintenance of aseptic culture of Trichomanes speciosum gametophytes from gemmae. American Fern Journal 87:87–92. Raudzens, L., and E. B. Matzke. 1968. Induced apospory in the liverwort Blasia pusilla. American Journal of Botany 55:1190–1196. Rogers, S. M. D., and S. Banister. 1992. Micropropagation of Notholaena Sun-Tuff fern. Hortscience 27:1224–1225. Rogge, G. D., A. M. Viana, and A. M. Randi. 2000. Cryopreservation of spores of Dicksonia sellowiana: an endangered tree fern indigenous to South and Central America. CryoLetters 21:223–230. Sakai, A., S. Kobayashi, and I. Oiyama. 1990. Cryopreservation of nucellar cells of navel orange (Citrus sinensis Osb. var. brasiliensis Tanaka) by vitrification. Plant Cell Reports 9:30–33. Sara, S. C., V. S. Manickam, and R. Antonisamy. 1998. Regeneration in kinetin-treated gametophytes of Nephrolepis multiflora (Roxb.) Jarret in Morton. Current Science 75:503–508. Sargent, M. L. 1988. A guide to the axenic culturing of a spectrum of bryophytes. Pages 17–24 in J. M. Glime (ed.), Methods in Bryology. Nichinan, Japan: Hattori Botanical Laboratory. Schenk, G. 1997. Moss Gardening. Portland, OR: Timber Press. Schuster, R. M. (ed.). 1983. New Manual of Bryology, Vol. 1. Nichinan, Japan: Hattori Botanical Laboratory. Sheffield, E., G. E. Douglas, and D. J. Cove. 1997. Growth and development of fern gametophytes in an airlift fermenter. Plant Cell Reports 16:561–564. Simabukuro, E. A., A. F. Dyer, and G. M. Felippe. 1998. The effect of sterilization and storage conditions on the viability of the spores of Cyathea delgadii. American Fern Journal 88:72–80. Söderström, L. 1998. Conservation of bryophytes in Europe: the work of the European Committee for Conservation of Bryophytes (ECCB). Pages 140–144 in H. Synge and J. Akeroyd (eds.), Planta Europa. Uppsala, Sweden: Proceedings of the Second European Conference on the Conservation of Wild Plants. Soede, A. C. 1981. Vermeerdering van Nephrolepis in weefselkweek nu betrouwbaavder. Vakblad voor de Bloemisterij 51:48–51.
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Stamps, R. H. 1992. Commercial leatherleaf fern culture in the United States of America. Pages 243–249 in J. M. Ide, A. C. Jermy, and A. M. Paul (eds.), Fern Horticulture: Past, Present and Future Perspectives. London: International Symposium on the Cultivation and Propagation of Pteridophytes. The British Pteridological Society (Intercept, Andover). Stokey, A. G. 1951. Duration of viability of spores of the Osmundaceae. American Fern Journal 41:111–115. Sulklyan, D. S., and P. N. Mehra. 1977. In vitro morphogenetic studies in Nephrolepis cordifolia. Phytomorphology 27:396–407. Takeuchi, M., H. Matsushima, and Y. Sugawara. 1980. Long-term freeze preservation of protoplasts of carrot and Marchantia. CryoLetters 1:519–524. Tannoury, M., J. Ralambosoa, M. Kaminski, and J. Dereuddre. 1991. Cryopreservation by vitrification of coated shoot-tips of carnation (Dianthus caryophyllus L.) cultured in vitro. Comptes Rendus de l’Academie des Sciences, Paris, Ser. III 313:633–638. Teng, W. L., and M. C. Teng. 1997. In vitro regeneration patterns of Platycerium bifurcatum leaf cell suspension culture. Plant Cell Reports 16:820–824. Thakur, R. C., Y. Hosoi, and K. Ishii. 1998. Rapid in vitro propagation of Matteuccia struthiopteris (L.) Todaro: an edible fern. Plant Cell Reports 18:203–208. Theuerkauf, W. D. 1993. South Indian pteridophytes: ex situ conservation. Indian Fern Journal 10:219–225. Wagner, W. H. Jr. 1995. Evolution of Hawaiian ferns and fern allies in relation to their conservation status. Pacific Science 49:31–41. Walter, K. S., and H. J. Gillett (eds.). 1998. 1997 IUCN Red List of Threatened Plants. Compiled by the World Conservation Monitoring Centre. Gland, Switzerland: IUCN. Wetmore, R. H. 1954. The use of in vitro cultures in the investigation of growth and differentiation in vascular plants. Brookhaven Symposium of Biology 6:22–38. Whittier, D. P. 1965. Obligate apogamy in Cheilanthes tomentosa and C. alabamensis. Botanical Gazette 133:336–339. Whittier, D. P. 1996. Extending the viability of Equisetum hyemale spores. American Fern Journal 86:114–118. Whittier, D. P., and T. A. Steeves. 1960. The induction of apogamy in the bracken fern. Canadian Journal of Botany 38:925–930. Windham, M. D., P. G. Wolf, and T. A. Ranker. 1986. Factors affecting prolonged spore viability in herbaceous collection of three species of Pellaea. American Fern Journal 76:141–148. Withers, L. A. 1985. Cryopreservation of cultured plant cells and protoplasts. Pages 243–267 in K. K. Kartha (ed.), Cryopreservation of Plant Cells and Organs. Boca Raton, FL: CRC Press. Whittier, D. P. 1995. Collecting in vitro for genetic resources conservation. Pages 511–526 in L. Guarino, V. R. Rao, and R. Reid (eds.), Collecting Plant Genetic Diversity. Technical Guidelines. Oxford, UK: CAB International. Zenkteler, E. K. 1991. Micropropagation of Polypodium vulgare L. by rhizome explants. Bulletin of the Polish Academy of Sciences Biological Sciences 43:77–84. Zenkteler, E. K. 1993. Homo- and heterophasis in the in vitro reproduction of Phyllitisscolopendrium (L.) Newm. Bulletin of the Polish Academy of Sciences Biological Sciences 41:257–261.
part three
The Ecological and Evolutionary Context of Ex Situ Plant Conservation Ex situ conservation is more than the retention of a Latin binomial in cultivation. It is the collection, management, and use of dynamic evolutionary units in artificial conditions. The primary conservation value of this whole enterprise is to support species survival in the wild and, if needed, to restore the material to the wild in a condition where it can survive and prosper. This process, which can be measured in months, decades, or even centuries, subjects the collected samples to a series of novel ecologies and selection pressures that can shape and distort the original composition of the sample, rendering it less able to survive in a wild environment. Part III reflects the increasing understanding of ex situ ecological and evolutionary dynamics that can influence a sample during its period of ex situ management, a period of transition from wild collection through storage and reintroduction. Samples of growing plants, tissue, or seed are subject to genetic modification, and this section reviews these changes and provides practical steps that can be taken to mitigate the potentially damaging impacts of ex situ life. The process of ex situ conservation implies the maintenance of a diverse and viable set of samples from point of collection through use. Husband and Campbell (Chapter 12) and Schaal and Leverich (Chapter 13) show that ex situ populations can be subject to profound modifications as a result of artificial selection pressures, an inevitable influence of the horticultural or storage environment, in effect a process of domestication. The advance of modern molecular tools has allowed an increasingly sophisticated understanding of ex situ genetic dynamics, but the application of these tools must be better understood. For as powerful as they are, molecular genetic tools do not address directly genetically based differences in adaptive traits. Vitt
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and Havens (Chapter 14) review quantitative genetic techniques and assess their application to the management of ex situ populations. The potentially detrimental impact on sampled populations of seed collection has long been of serious concern to many, and until now it has not been addressed quantitatively. Menges et al. (Chapter 15) use computer simulation to estimate the relative impact on extinction probability of various intensities and frequencies of seed collection for populations of different sizes. Without minimizing the uncertainties involved, they provide a series of general harvest recommendations designed to avoid impacting the viability and status of wild populations. But simply collecting a genetically representative sample is not enough to ensure that sample can be used to reestablish a comparable population. Once collected, the samples are subject to numerical and genetic erosion as a result of mortality patterns during storage and reintroduction. Guerrant and Fiedler (Chapter 17) review these often cryptic processes and assess their impact on ex situ management. These findings are applied in the genetic sampling guidelines (Appendix 1). Hybridization is both a widespread evolutionary influence on wild populations and a valued horticultural tool, but it can wreak havoc in an ex situ conservation setting. Traditional plant collections can be characterized by high levels of artificial sympatry. When naturally isolated taxa are grown together, conditions are ideal for hybridization. Maunder et al. (Chapter 16) review the impact of hybridization on ex situ management. Ex situ conservation has a proven ability to retain high levels of alpha or taxonomic diversity; indeed, Franklinia (Theaceae) graces gardens after extinction in the wild because of prolonged cultivation. However, the ability to maintain evolutionary lineages will depend on an understanding of the ecology, genetics, and demography of collection, cultivation, and storage.
Chapter 12
Population Responses to Novel Environments: Implications for Ex Situ Plant Conservation Brian C. Husband and Lesley G. Campbell
The primary goal of ex situ plant conservation is to establish and maintain seed or growing collections of wild species outside their natural habitat for the direct or indirect purposes of species recovery in situ. Such programs typically involve three stages: collection from natural populations, establishment and maintenance of seed or growing material off site and, when appropriate, use of ex situ plant material for in situ reintroduction efforts. Viewed in this way, from initial collection to final end use, the success of an ex situ conservation program depends on its ability to adequately represent the species of interest in the ex situ population and to preserve the utility of the population for future recovery efforts. The challenge for conservationists, then, is to determine the genetic and demographic factors that affect the implementation and long-term utility of an ex situ conservation program. In this chapter, we identify the critical genetic and demographic factors influencing ex situ conservation by examining relevant evolutionary principles. We begin by considering the historical role of evolutionary biology in plant conservation and the need for its greater use in ex situ conservation. We argue that the challenges facing ex situ conservation programs can be viewed within the conceptual framework of populations in abruptly changing environments. Following on this theme, we characterize the selective environment that a transplanted population will experience and the demographic consequences that such an environmental shift may impose. We then explore the genetic and demographic factors that may influence the success of such a transplantation or colonization event. Finally, we discuss the implications of our findings for ex situ conservation programs and suggest steps that increase the chances for success. 231
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Evolutionary Biology in Ex Situ Conservation Evolutionary biology has contributed much to the development of conservation, particularly as it relates to the protection of extant populations in situ. For example, the theoretical and empirical literature regarding the ecology and genetics of small populations (Franklin 1980; Lande 1988, 1993; Barrett and Kohn 1991; Rieseberg 1991; Schemske et al. 1994; Lynch et al. 1995; Newman and Pilson 1997; Fischer and Matthies 1998a, 1998b; Bataillon and Kirkpatrick 2000) has helped to quantify the risks of extinction facing many threatened species. In addition, knowledge of the organization of genetic variation (Frankel and Brown 1984; Brown and Briggs 1991) has led to the refinement of collection and management strategies used in national programs of plant conservation (Falk 1991; Center for Plant Conservation 1991). Unfortunately, the evolutionary biology of ex situ conservation for wild species has been less well explored and therefore has played a smaller role in conservation practice. As a result, many ex situ collections have involved a small number of threatened taxa whose representation in collections is narrow (Brown and Briggs 1991; Maunder et al. 1999). The genetic base on which many reintroduction programs are founded is likely to be equally narrow. To determine the potential role of evolutionary biology for guiding offsite collections and restorative plantings, we conducted a survey of recent ex situ plantings and in situ reintroductions for several North American species at risk, available through online searches (Table 12.1). In total, 50 cases were identified; although the list is far from exhaustive, it shows that 79 percent of all plantings were based on propagules from only a single source population, and 50 percent of these plantings were based on fewer than 10 source individuals (Table 12.1). Clearly, collections of threatened species, for either ex situ or in situ conservation, are necessarily constrained by the limited availability of source material (Brown and Briggs 1991). Indeed, many of the species listed in Table 12.1 are not known from more than a single population. It is especially important for this reason that conservation programs consider the genetic (e.g., genetic variance, population differentiation, inbreeding) and ecological (e.g., number of individuals, habitat characteristics) attributes of organisms to ensure that plantings, both on and off site, are viable in the long term and can meet specified recovery targets. Interestingly however, in our limited survey, 85 percent
table 12.1 Summary of recent plantings and reintroductions conducted as part of recovery programs for rare or endangered species. We indicate the habitat (ex situ vs. in situ, including the ecological type), the number of source populations, and number of transplanted propagules when available. Also, we note whether ecological (cultural methods, choice of habitat) or genetic criteria (diversity, source of samples, facilitating evolutionary response in new environment) are given consideration and the outcome of the transplants when given.
Species
Habitat
Source Material: Number of Populations/Individuals
Ex Situ Grevillea scapigera Conradina glabra Carex paupercula Carex capillaris Viola rupestris Galium boreale Draba incana Tofieldia pusilla Antennaria dioica Carex ericetorum Juncus alpinoarticulatus Thalictrum alpinum Gentiana verna Primula farinosa Ruiza cordata Dombeya rodriguesiana Penstemon barrettiae Penstemon barrettiae
Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex Ex
na/10 ind. 1 pop./na Local/na Local/na Local/na Local/na Local/na Local/na Local/na Local/na Local/na Local/na Local/na Local/na 1 pop./1 ind. na/2 ind. 1 pop./na 1 pop./na
situ situ situ situ situ situ situ situ situ situ situ situ situ situ situ situ situ situ
Transplant Material: Number of Propagules
Criteria for Sample
Success Rate of Transplants
Source
1,300 plants — 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 50 plants 20 cuttings Cuttings Cuttings —
— — Ecological Ecological Ecological Ecological Ecological Ecological Ecological Ecological Ecological Ecological Ecological Ecological — — — Genetic
na Established Established Established Established Established Established Established Limited Limited Limited Limited 0% 0% 5% na 34.6% 0%
13 10 9 9 9 9 9 9 9 9 9 9 9 9 11 17 18 18
table 12.1 (continued) Summary of recent plantings and reintroductions conducted as part of recovery programs for rare or endangered species. We indicate the habitat (ex situ vs. in situ, including the ecological type), the number of source populations, and number of transplanted propagules when available. Also, we note whether ecological (cultural methods, choice of habitat) or genetic criteria (diversity, source of samples, facilitating evolutionary response in new environment) are given consideration and the outcome of the transplants when given.
Species
Habitat
Source Material: Number of Populations/Individuals
In Situ Hymenoxys herbacea Amsinkia grandifolia Cordylanthus maritimus Cordylanthus maritimus Lupinus guadalupensis Lupinus guadalupensis
Alvar — Marsh Marsh Grassland, scrub Grassland, scrub
— — 1 marsh 1 marsh 1 pop./50 ind. 3 pop./150 ind.
Argyroxiphium sandwicense Styrax texana Cirsium tuberosum Cirsium pitcheri Pediocactus knowltonii Stephanomeria malheurensis Pseudophoenix sargentii Opuntia corallicola Jacquemontia reclinata
Cliffs — Grassland Shoreline Dry woodland In situ na — —
1 pop./2 ind. na/40 ind. 1 pop./1 plant 2 pop./10 plants 1 pop./250 ind. 1 pop./na 5 ind. — —
Agropyron scribneri Agropyron trachycaulum Carex drummondii Deschampsia caespitosa
Alpine mine Alpine mine Alpine mine Alpine mine
Local/many Local/many Local/many Local/many
soil soil soil soil
Transplant Material: Number of Propagules
— — na 10,000 seeds 750 seeds, 4 sites 2,250 seeds/3 pop. Seedlings — 6 cuttings 78 plants 150 cuttings 1,000 seedlings 250 seedlings 340 propagules Experimental outplanting Many seeds Many seeds Many seeds Many seeds
Criteria for Sample
Success Rate of Transplants
Source
Genetic — — Genetic Ecological Genetic, ecological — — — Ecological — Ecological — — —
— — 0% (20 yr) Pres. (9 yr) 50% (1 yr) 100% (1 yr)
1 2 3 4 5 5
Pres. (27 yr) — 83% (4 mo) 29.9% (1 yr) 83% (3 yr) — “Good” (9 yr) — —
6 7 15 25 14 20, 26 — — —
Ecological Ecological Ecological Ecological
0% Established 0% Established
8 8 8 8
Species
Habitat
Phleum alpinum Poa alpina Trisetum spicatum Triticum aestivum Conradina glabra Deschampsia caespitosa Carex pyrenaica Kobresia myosuroides Sibbaldia procumbens Acomastylis rossii Carex rupestris Phebalium glandulosum Phebalium equestre Dictyosperumum album Craceana concinna Penstemon barrettiae Lupinus sericatus Gasteria baylissiana Amsonia kearneyana Amsonia kearneyana Helianthus schweinitzii Paphiopedilum rothschildianum
Alpine mine Alpine mine Alpine mine Alpine mine In situ Alpine Alpine Alpine Alpine Alpine Alpine In situ In situ In situ In situ In situ In situ In situ Canyon Canyon Prairie Serpentine
Schwalbea americana
Savanna
soil soil soil soil
Source Material: Number of Populations/Individuals
Transplant Material: Number of Propagules
Local/many Local/many Local/many Local/many — na/280 ind. na/70 ind. na/70 ind. na/140 ind. na/490 ind. na/140 ind. na/na na/na 1 pop./1 ind. na/na 1 pop./21 ind. na/na na/10 ind. na/8 ind. na/8 ind. 1 pop./na 1 pop./2 capsules
Many seeds Many seeds Many seeds Many seeds 1,300 280 70 70 140 490 140 cuttings — 50 plants — 70 cuttings Plants, seedlings 210 plants 76 1 yr old 105 2 yr old 80 plants 100 seedlings
1 pop./12 genets
131 seedlings
Criteria for Sample
Ecological Ecological Ecological Ecological Genetic Ecological Ecological Ecological Ecological Ecological Ecological — — — — — — — — — — Ecological, genetic —
Success Rate of Transplants
Source
Established Established 0% Established 95%/1 yr 98% 70% 83% 44% 79% 17% 75% (3 yr) 27–83% (2 yr) na na 50% (3 yr) 50% (5 mo) — 46% (3 mo) 97% na na
8 8 8 8 10 12 12 12 12 12 12 16 16 17 17 18 19 21 22 22 23 24
3.8%
27
Sources: De Mauro (1994); Pavlik et al. (1993); Helenurm and Parsons (1997); Parsons and Zedler (1997); Helenurm (1998); Robichaux et al. (1997); 7 Cox (1990); 8Brown and Johnston (1976); 9Cranston and Valentine (1983); 10Wallace (1992); 11Lesouef (1988); 12May et al. (1982); 13Rossetto and Dixon (1993); 14Olwell et al. (1990); 15Pigott (1988); 16Jusaitis (1991); 17Anonymous (1993); 18Guerrant (1990); 19Edmunson et al. (1984); 20Anonymous (1988); 21van Jaarsveld (1994); 22Reichenbacher (1990); 23Anonymous (1989); 24Grell et al. (1988); 25Bowles and Flakne (1993); 26Parenti and Guerrant (1990); 27Obee and Cartica (1997). na, data not available; ind., individual(s); pop., population(s); pres., present 1
2
3
4
5
6
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and 26 percent of these cases did not state explicit genetic or ecological criteria, respectively, to guide their programs. When these criteria were considered, they were most often applied to decisions regarding sampling or collecting the original material but rarely in decisions regarding storage or establishment of plantings off site and subsequent reintroduction efforts. Developing scientifically informed programs of ex situ conservation will involve more than simply applying the theory that is applicable to in situ conservation because their goals and procedures can be quite distinct. In contrast to in situ methods, ex situ conservation involves sampling material from existing populations. Sampling error and loss of genetic diversity through genetic drift are likely to be important in both aspects of conservation and in fact may be exaggerated in ex situ efforts by the joint effects of small source populations and finite samples. Currently, there are several excellent discussions in the literature concerned with the biology and the conservation risks associated with small populations (Soulé and Simberloff 1986; Barrett and Kohn 1991; Lande 1993). Perhaps what is most distinct about ex situ conservation, in comparison with in situ, is that target plants are being removed from their native location and introduced and maintained in a new environment whose abiotic and biotic conditions certainly are different from that of the original population. Although the specific environment of the ex situ collection will vary widely among cases, the expectation of an abrupt shift in the ex situ environment will apply equally to seed collections and actively growing plant material. Whereas in one case (in situ) the conservationist is attempting to facilitate survival and growth of an extant population, in the other he or she is potentially adding a new source of endangerment, namely maladaptation. Because the biological features of in situ and ex situ populations are inherently different, so too are the best management practices necessary for success. In particular, the conservation biologist is faced with the challenge of collecting a sample that is genetically representative and maintaining it in an environment that may be anything but ecologically representative of the native habitat. Furthermore, ex situ collections are not an end in themselves but rather a means toward the goal of long-term viability for populations in situ. Therefore, ex situ collections often must be managed to simultaneously maintain their short-term viability off site and their long-term utility in restoration or reintroduction efforts (Guerrant 1996).
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A conceptual framework and some practical guidelines have been developed for ex situ conservation of crop plants (Brown and Marshall 1995; Schoen and Brown 1995). This application of evolutionary biology may not be completely transferable to wild species because the goals and the target species for the latter are somewhat different. For crop species, ex situ collections are more permanent and serve primarily as sources of single characters or individual genes that plant breeders will transfer into locally adapted stocks. For wild organisms, ex situ programs operate on a short- to long-term timescale and often originate from a smaller number of plants. In addition, the primary goal of the ex situ collection is to facilitate demographic viability of the species in the wild, not genetic preservation, so the ex situ collection must provide a source of propagules that can be assembled into whole, functioning populations. Moreover, conserving wild species off site is more complex because of their diverse and complex life histories, variable mating systems, and lower storage tolerance (Brown and Briggs 1991).
Population Response to New Environments From an evolutionary perspective, an ex situ conservation program comprises all the basic elements of a colonizing event (Templeton 1991) and therefore is affected by many of the same evolutionary forces. As in most founder events, establishing an off-site collection involves sampling a finite number of individuals or propagules from one or many source populations. This sample forms the basis of a new population, maintained as growing or dormant, which occupies a location and environment different from what had previously been occupied. The evolutionary factors acting on a population in changing environments have been examined in theoretical terms by Lande (1993), Lynch and co-workers (Lynch and Lande 1993; Lynch et al. 1991; Bürger and Lynch 1995), and Gomulkiewicz and Holt (1995). Here we briefly describe the theory and empirical support for two aspects of a colonizing population: maladaptation in the new environment and the response to the new environment.
Maladaptation When a population experiences an abrupt change in environment, the phenotypic distribution for a given quantitative trait (z) may initially be
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Figure 12.1 Graphic representation of the initial state of a population (t = 0) that has been subjected to an abrupt change in environment. The solid line depicts the fitness function for a quantitative trait, z. The dotted line represents the frequency distribution of phenotypes in the initial population. The difference between the optimal phenotype (fitness = maximum) and the average phenotype represents the magnitude of maladaptation experienced by the population. (Modified from Gomulkiewicz and Holt 1995.)
displaced from the optimal phenotypic distribution for that environment, depicted as its fitness function (w[z]). As depicted by Gomulkiewicz and Holt (1995), the magnitude of the difference between the optimal phenotype (maximum fitness) and the population mean phenotype represents the degree of maladaptation experienced by the population with respect to that trait (Figure 12.1). In quantitative genetic terms, maladaptation is roughly equivalent to the selection differential (Falconer 1989). For the field practitioner, the degree of maladaptation is directly related to the difference between the source and ex situ environments (assuming the population was adapted to its native habitat). Moreover, this description applies to a single phenotypic character, which is reasonable only if plant fitness in the ex situ or transplant environment is limited by a single phenotypic attribute. In most cases, viability and fertility are regulated by a more complex interaction between the environment and a suite of interrelated characters (zi, i = 1 . . . n). In this case, the magnitude of maladaptation experienced by a population is the product of maladaptation with respect to each phenotypic character, phenotypic covariances between traits, and their individual contributions to fitness (Lande 1982). Where increased fit-
12. Population Responses to Novel Environments
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ness in a new environment depends on a combination of traits, maladaptation is best viewed in multivariate space (Lande and Arnold 1983). Although colonization is an integral feature of all plant species, in natural or artificially established populations the initial degree of maladaptation (deviation of mean phenotype from optimal phenotype) associated with a population founding event or transplant has rarely been estimated, even for a single character. The lack of estimates may be explained in part by the difficulty in calculating the parameter, given the many assumptions that are unlikely to be met. For example, measuring the degree of maladaptation entails knowing which characters are of adaptive value. These traits are likely to provide the most meaningful measure of maladaptation; however, without preliminary research on the physiological, ecological, or morphological limits imposed by a particular environment, the critical characters may not be readily identifiable. Moreover, estimates of maladaptation also depend on knowledge of the optimal phenotype for a given habitat at a given time, which is rarely known for natural populations. In practice, an approximation of the optimal phenotype for a particular environment could be obtained in two ways. The first method involves using the mean phenotype of well-established individuals in a population as an estimate of the optimal phenotype for that environment. This method requires that the plant of interest is already established in the target environment and assumes that established plants are evolutionarily stable, a premise likely to be violated in temporally or spatially fluctuating environments. A second approach is to use phenotypic selection methods (Lande and Arnold 1983) to estimate the relationship between multivariate phenotypes and fitness for a collection of plants. In this case, if the shape of the selection surface is identified, the phenotypes with the highest relative fitness can be assumed to be optimal for that particular environment. This approach could conceivably be conducted on established plants or newly transplanted individuals if sufficient phenotypic variation is available. We used the first approach to estimate the magnitude of maladaptation in plants. Specifically, we surveyed the literature for data from transplant and common garden studies, which contained phenotypic measures for resident and alien (foreign) populations transplanted into the same location. Maladaptation was calculated as the percentage difference between the phenotype of the alien population and that of the resident population. Values of maladaptation were highly variable, ranging from 0 to 900 percent, depending on the phenotype measured and the species involved. In general
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maladaptation was high, reflecting the fact that local differentiation of plant populations is widespread (Table 12.2). An interesting pattern emerged with respect to the ecological similarity between source and recipient environments. In cases where plants have been transplanted into ecologically similar habitats that are in close proximity to the source population, the degree of maladaptation averaged 51 percent, which was similar to examples in which transplants were moved farther away (47 percent). This confirms that the degree of maladaptation experienced by transplants is determined largely by the ecological similarity, not geographic proximity, of habitats. Interestingly, plants transplanted from their native populations to sites altered indirectly by human activity experienced a much higher degree of maladaptation, averaging 254 percent. Although crude, the results from these transplants confirm that the magnitude of maladaptation probably depends on the similarity between source and recipient habitats. One clear example in which maladaptation has been quantified involves the colonization of contaminated mine soils by the grass Anthoxanthum odoratum (Antonovics 1976; Grant 1974). In this case, the source of the colonists on the mine tailings (adjacent pasture population) and the relevant ecological trait (tolerance to heavy metals) can be identified with little argument. Physiological tests showed that the mean index of tolerance in the mine site (75.4) was 8.6 times as large as the mean for the pasture (8.8). Assuming that the value for the mine population represents the optimum degree of tolerance in the “new” environment, the percentage difference (757 percent) can be taken as the degree of maladaptation experienced by colonists from the pasture. This example provides a rough indication of the strength of selection experienced by populations that are transplanted into ecological settings beyond that found in their natural habitats, particularly human-influenced environments (Jain and Bradshaw 1966). Despite the high degree of maladaption estimated for some species in our survey, it should be noted that species having no detectable maladaptation were observed in two of the three transplant categories (Table 12.2). An additional example of a species with small differences between populations was observed in a common garden study of ecological differentiation in Morus rubra, an endangered tree in Canada (Burgess and Husband, unpublished data 2003). Only about 200 plants remain in Canada, the majority of which are located in six core populations in southern Ontario. Burgess and Husband found only minor differences in growth
table 12.2 Estimates of maladaptation and relative fitness of plants transplanted into new habitats that are occupied by conspecifics. Data are taken from selected transplant studies involving natural and human-influenced environments. Maladaptation is estimated as the percentage phenotypic difference between alien and resident plants for ecologically important traits. Relative fitness is the fitness of alien transplants relative to resident controls. Species
Environmental Gradient
Adaptive Character
Natural Habitats: Local Scale Zostera marina
Water depth
Pmax Respiration rate Chlorophyll content Shoot compactness Aphid resistance Leaf morphology Peduncle height
Prunella spp. Polemonium viscosum Dryas octopetala Plantago major Natural Habitats: Regional Scale Poa annua
Altitude Altitude Snowbank Mowing Moisture
Time to flowering Number of inflorescences Reproductive biomass
Phenotypic Deviation (%)
7.0a 0.0 10.1 163.0 39.0 80.9 57.0a 14.1 81.0 28.0
Relative Fitness of Alien
0.47 (composite) 0.47 (composite) 0.47 (composite) 0.69a 0.41 (survival) 0.78 0.18a (reproduction)
Source
1 1 1 7 12 8 17 11 11 11
table 12.2 (continued) Estimates of maladaptation and relative fitness of plants transplanted into new habitats that are occupied by conspecifics. Data are taken from selected transplant studies involving natural and human-influenced environments. Maladaptation is estimated as the percentage phenotypic difference between alien and resident plants for ecologically important traits. Relative fitness is the fitness of alien transplants relative to resident controls. Species
Environmental Gradient
Adaptive Character
Quercus rubra Phlox drummondii Phlox drummondii Human-Influenced Habitats Ceratodon purpureus Agrostis tenuis Plantago major Festuca ovina Lotus corniculatus Lotus purshianus Lupinus bicolor Plantago lanceolata Cynodon dactylon Typha latifolia Anthoxanthum odoratum
Aspect
Herbivore resistance Life history (survival) Life history (fecundity)
Metals Metals Ozone Metals Metals Metal Metals Metals Metals High metals High metals
Metal tolerance Cu tolerance Ozone resistance Pb tolerance Pb tolerance Cu tolerance Cu tolerance Pb tolerance Pb tolerance Metal tolerance Metal tolerance
Phenotypic Deviation (%)
Relative Fitness of Alien
40.5a,b 25.5 47.0
0.57 (lambda) 0.57 (lambda)
900.0b 212.0 31.9 60.0 0.0 527.0 174.0 133.0 0.0 0 757
0.0001
Source
5 9 9 13 14 10 15 15 16 16 4 4 2, 3 6
Sources: 1Dennison and Alberte (1986); 2Taylor and Crowder (1984); 3McNaughton et al. (1974); 4Wu and Antonovics (1976); 5Sork et al. (1993); 6Antonovics (1976); 7Fritsche and Kaltz (2000); 8McGraw and Antonovics (1983); 9Schmidt and Levin (1985); 10Reiling and Davison (1992); 11Till-Bottraud et al. (1990); 12 Galen et al. (1991); 13Jules and Shaw (1994); 14McNeilly and Bradshaw (1968); 15Shaw (1984); 16Wu and Kruckeberg (1985); 17Warwick and Briggs (1980). a Maladaptation value average from several habitats. b Data estimated from figures in publication.
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and morphology between plants from six different populations and from two different habitat types (cliff faces and sand spits). If one assumes that the mean phenotype approximates the optimal phenotype for a particular environment, then these results suggest that transplants between populations would experience little maladaptation because these characters do not differ between populations in this region. Although one cannot rule out differences for other characters or different environments that have not yet been observed, such observations raise the larger question of whether there are specific ecological, physiological, and life history attributes that are associated with low maladaptation. Although our survey is too limited to detect such patterns, additional research would be beneficial for identifying the species most likely to exhibit adaptive differentiation between populations and therefore most likely to experience maladaptation.
Response to Novel Environments A population’s response to maladaptation can be divided into two parts: short-term changes in survival and reproduction caused by the physiological impacts of the new environment and long-term population growth and persistence, determined by the adaptive response to the new environment. The short-term response represents the immediate demographic impact of maladaptation. The larger the degree of maladaptation, the more likely a population’s vital rates (birth, survival) and mean fitness will decline. Especially relevant when selection in a new environment is hard, this reduction in growth is called the demographic cost of selection, or demographic load (Haldane 1957; Gomulkiewicz and Holt 1995). It should also be noted that when plants are being introduced into new habitats, the short-term response may also include the effects of transplant shock and damage. This additional effect has a separate cause in that it is imposed by the act of introducing growing plants, but it can be confounding because its impact may be greater in maladapted plants. Introducing plants in the seed stage may ameliorate this problem. Surrogate estimates for the demographic cost of maladaptation can be gleaned from the literature on ecological differentiation in plants, in which foreign plants are compared with residents with respect to their fitness components (summarized in Bradshaw 1984; Levin 1984; Huenneke 1991; Table 12.1). We consider these minimum values simply because the recipient habitats are already occupied by the transplanted species, so the transplant
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environment probably is less extreme than when transplants are introduced to previously unoccupied environments. Bradshaw (1984) reviewed several of these transplant experiments and found that the fitness of transplants moved to environments similar to that of their origin often was about half that of native residents, a pattern supported by other surveys (Davies and Snaydon 1976; Levin 1984), including results summarized in Table 12.2 on transplants into natural habitats. This result suggests that fitness costs associated with maladaptation are widespread among plant species, even when plants are transplanted between apparently similar environments. Bradshaw (1984) also found that plants moved to noticeably different environments often were less than one-tenth as fit as residents. This value stems primarily from examples of species that have colonized mine tailings (Jain and Bradshaw 1966; Table 12.2). In the case of A. odoratum (Antonovics 1976; Grant 1974), the distribution of metal tolerance in the pasture plants overlaps with that of the mine tailing (a measure of the optimal fitness function) by less than 1 percent, suggesting that only an extremely small fraction of the initial colonists would ever survive (Table 12.2). This study and many others (Turkington and Harper 1979; Silander 1985) show clearly that the demographic response to maladaptation, and hence the decline in population size, can be extremely large even for plants colonizing neighboring habitats. The long-term population response to a novel environment depends on the magnitude of maladaptation and the potential for an adaptive change in the new environment (Figure 12.2). In the absence of an evolutionary response to selection, the mean phenotype of the population and the degree of maladaptation remain unchanged. As a result, the population vital rates are depressed, and the population size therefore is expected to decline (Figure 12.2A). If the population size falls below some critical size, the impact of demographic stochasticity, inbreeding, and genetic stochasticity predominate and may increase the population’s vulnerability to extinction (Goodman 1987; Lande 1993). However, if the population can respond to the selective pressures, the mean population phenotype shifts toward the optimum and the degree of maladaptation declines. This adaptive response reduces maladaptation and hence the magnitude of demographic load. As a result, the population growth rate should increase, thereby rescuing the population from extinction (Figure 12.2B). Gomulkiewicz and Holt (1995) also point out that the ability to respond to selection doesn’t eliminate the risk of extinction. In particular, if the evolutionary response is slow, the popu-
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Figure 12.2 Graphic representation of the evolutionary response to an abrupt shift in the environment under two conditions in which (A) the population is devoid of potentially adaptive genetic variation or (B) the population contains adaptive genetic variation. In the absence of genetic variation, no evolutionary shift in mean phenotype toward the optimum occurs; therefore, the population declines below the critical population size and ultimately goes extinct. In the presence of adaptive genetic variation, the population declines initially because of demographic load; however, as the population adapts to the new environment and its phenotype shifts toward the optimum, the average fitness increases and the population begins to increase. (Modified from Gomulkiewicz and Holt 1995.)
lation may still be at risk because the reduction in maladaptation won’t be sufficient to prevent the population from falling below some critical population size (Figure 12.2B). Although the theory shows that an evolutionary response can increase the time to extinction for a population in a novel environment, it remains unclear to what extent maladaptation influences the longevity of real populations and whether an evolutionary response actually reduces the risk of
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extinction. The population response to a novel environment is complex and probably depends on many genetic and demographic factors, which must be understood before any useful predictions can be formulated.
Constraints on the Adaptive Response The long-term persistence of a population, which experiences an abrupt change in environment, depends, in part, on its ability to adapt to its new environment. What determines whether this evolutionary response will be sufficient to prevent extinction and lead to successful establishment? Here we divide these factors into those that are genetic and those that are ecological.
Genetic Constraints Many of the genetic constraints acting on newly transplanted populations arise as a result of genetic drift, which increases in finite populations (Kimura and Crow 1963). Limits to population size are inevitable in most founder events and arise during the population bottleneck associated with the initial founding event and during the establishment (or maintenance in the case of ex situ collections) phase of a new population. The theoretical effects of genetic drift in small populations are widely understood (Nei et al. 1975; Maruyama and Fuerst 1985a, 1985b; Lacy 1987; Barrett and Kohn 1991; Lande 1994). In most cases, drift is expected to cause random fluctuations in allele frequencies across generations, which can lead to the loss of potentially adaptive variation, inbreeding depression, and the fixation of mildly deleterious mutations (Lande 1994). Here we briefly define these processes and examine how they may affect a population’s response to a new environment.
Genetic Variance Genetic drift in small populations increases the temporal variance in allele frequencies and the loss of potentially adaptive genetic variation, particularly rare alleles. The magnitude of this effect is inversely proportional to effective population size, Ne. In a typical colonizing event, the loss of variability is influenced by the size of the initial founding population and the size of the population in subsequent years (Nei et al. 1975). The predicted effects
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of drift on genetic variability are theoretically well established and have been confirmed most convincingly through a large number of studies that have found a positive relationship between the log of population size and isozyme diversity (Barrett and Kohn 1991; Barrett and Husband 1997; Frankham 1996). However, most studies on drift have focused on diversity at neutral genes (although for genes under selection, see Barrett et al. 1989; Husband and Barrett 1992a, 1992b; Barrett and Husband 1997), and it is less clear whether a decline in such diversity has any detectable negative impacts on population viability or evolutionary potential. Most of the characters of interest to ecologists and conservation biologists are metric or quantitative and are influenced by many genes and the environment. Because genetic variation is virtually ubiquitous for quantitative traits (Bürger and Lynch 1995), most populations have some capacity to respond to selective challenges. However, in finite populations the genetic variance can indeed drift (Bürger et al. 1989; Keightley and Hill 1988; Zeng and Cockerham 1991), and if maladaptation is large, the demographic costs of selection may be high (Bürger and Lynch 1995). Reduced genetic variance has two related consequences for the response to novel environments. First, it diminishes or prevents an evolutionary response to selection from occurring. The importance of heritable variation to evolutionary change is shown most simply by the breeder’s equation, R = S h2, which indicates that the response to selection (R), measured as the between-generational change in mean phenotype in a population, is the product of S, the magnitude of selection (maladaptation) acting on a particular trait, and h2, the narrow-sense heritability of that trait. Although the response to selection also depends on the genetic covariance between traits and selection gradients acting on them (Lande 1979, 1982), in general, a loss of heritable (genetic) variation determines the degree and rate of response to a new selective environment. Put another way, a reduction in heritable variation reduces the degree of maladaptation from which a colonizing population can be rescued by an evolutionary response. Although an evolutionary response can still occur in a population with reduced variability, the time frame over which the response occurs may be long enough to increase the length of time a population is small and thereby increase its vulnerability to extinction (Bürger and Lynch 1995; Gomulkiewicz and Holt 1995). Similar evolutionary dynamics probably operate at the mar-
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gins of species’ ranges and may be an important determinant of the position and shape of species boundaries (Antonovics 1976; Hoffmann and Blows 1994; Kirkpatrick and Barton 1997). Despite its evolutionary significance, no comprehensive evaluation of the importance of genetic variance to population viability and persistence exists (Lande and Shannon 1996). Therefore, to assess the importance of genetic variance for the successful establishment of new populations, one needs some measure of the frequency of transplant failures and the genetic context in which they fail. Because of the nature of scientific discourse, however, it is easier to locate transplant or colonization studies in which the founding plants survive and reproduce than studies of transplants that fail. Nevertheless, there are still plenty of cases in which the founding plants were not completely successful (Table 12.1), and in some of those cases, a lack of genetic variance may be a contributing factor. For example, in outplantings of the endangered Lupinus guadalupensis, Helenurm (1998) observed a very high degree of mortality among transplants, an outcome that is echoed for many plant reintroduction attempts to date (Table 12.1; Pavlik et al. 1993 and references therein). These results suggest that either an evolutionary response did not occur in these newly established populations or the response was not sufficiently quick to prevent the population from decreasing below some minimum size. Helenurm (1998) also found that transplant success was higher on average for certain source populations, the largest in this case, as well as for certain recipient locations. For this reason, he recommended that large populations are better sources of seed, either because they are likely to retain more variation or perhaps they produce the most vigorous or stress-tolerant propagules. This study also highlights the fact that, for many studies in conservation, it is difficult to identify the precise contributions of the adaptive response (or lack thereof) to such transplant failure (compared with transplant shock, for example) because fitness measures of transplants from individuals native to the new environment are rarely available for comparison. In one of the earliest studies to examine the adaptive role of genetic variation in plant populations, Martins and Jain (1979) planted 135 colonies of Trifolium hirtum with low, medium, or high levels of allozyme variability. Colonization success of the colonies was not related to the initial degree of genetic polymorphism. This may not be surprising, given that the treatments represented differences in allozyme diversity, which are often considered to be selectively near neutral, rather than for traits of ecological
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significance. In contrast, Newman and Pilson (1997) observed a greater probability of extinction in experimental populations of Clarkia pulchella with lower genetic variability. They constructed populations of low and high effective size by manipulating the relatedness of their members. However, in this example it is not possible to distinguish the effects of low variability from the effects of inbreeding. Research on the evolution of metal tolerance and herbicide resistance may provide the most convincing evidence of the importance of appropriate adaptive variation for long-term success after colonization (Bradshaw 1991; Warwick 1991). For example, Bradshaw (1991) indicated that only a small subset of the species exposed to metal-contaminated soils or herbicides have successfully colonized, a pattern that he attributes to the lack of suitable genetic variation and evolutionary response in these extreme environments. Conversely, populations of species such as Anthoxanthum, Agrostis, and Mimulus, which have evolved tolerances to such extreme conditions, appear to have responded sufficiently rapidly to remain demographically viable. Interestingly, in these species the genetic basis of tolerance often is simple and the heritability high (McNair 1993), two attributes that increase the likelihood of persistence in a new environment. However, beyond these extraordinary cases, our understanding of the genetic basis of ecotypic differences remains minimal, and further research is necessary to predict the likelihood of an adaptive response for different species and in a range of different environments.
Deleterious Mutations Mutations are the ultimate source of adaptive genetic variability. However, a large proportion of mutations are detrimental (Drake et al. 1998) and can reduce a population’s ability to persist in a novel selective environment, especially when the population is small (Lynch and Gabriel 1990; Lynch et al. 1995; Lande 1995; Lande and Shannon 1996). Deleterious mutations can constrain the response to selection through two mechanisms: inbreeding depression and mutational meltdown. If the initial founders in a colonization event harbor any recessive or partially recessive deleterious mutations, their effects may be expressed during the population bottleneck because small populations may experience a higher probability of inbreeding and homozygosity. Specifically, increases in the inbreeding coefficient, which is a measure of the history
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of inbreeding, will occur at a rate that is inversely proportional to the effective population size (Kimura and Crow 1963). Inbreeding depression, measured as the reduction in fitness of inbred offspring relative to randomly outcrossed offspring, may occur as homozygosity in small populations rises and deleterious mutations are expressed. The fitness effects of inbreeding can be severe, especially in conifers and perennial species that are outcrossing or partially outcrossing (Husband and Schemske 1996). For example, a single generation of self-fertilization in Chamerion angustifolium can cause a decrease in fitness of up to 95 percent (Husband and Schemske 1995, 1997) and manifest itself during embryo development, seed maturation, germination, and flowering. A recent survey by Husband and Schemske (1996) suggests that, in general, species that are outcrossed experience more inbreeding depression after a single bout of self-fertilization than species that are historically inbred. This pattern is consistent with theoretical models that assume that most deleterious mutations are caused by lethal recessive mutations (Lande and Schemske 1985; Charlesworth and Charlesworth 1987) and that with increased inbreeding many of these mutations are selectively purged from the population. However, it should be noted that selfing species often exhibit some inbreeding depression, albeit in later life stages than in outcrossed offspring (Husband and Schemske 1996). This observation argues for the presence of partially recessive mutations of mild effect, which are more difficult to purge, even with repeated bouts of inbreeding (Byers and Waller 1999). If inbreeding depression is substantial and sustained over time, inbreeding could counteract the benefits of any adaptive response and reduce population survival in a new environment. Although the impacts of inbreeding depression are well documented in controlled experiments, a lack of data on changes in inbreeding during bottlenecks and associated declines in fitness precludes any clear demonstration of its relative importance in natural populations. Theoretical studies have indicated that inbreeding and a loss of heterozygosity may be negligible during a population bottleneck unless the founding population remains small for an extended period of time (Nei et al. 1975; Maruyama and Fuerst 1985a, 1985b). This is consistent with a comparison of stable and bottlenecked populations of the aquatic Eichhornia paniculata in Brazil (Barrett and Husband 1997). Despite the fact that population bottlenecks resulted in a 90 percent decrease in population size, bottlenecked populations were no less het-
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erozygous than stable ones. In this case, a contributing factor may also be that the bottlenecks were simply not severe enough to cause a reduction in genetic diversity. In addition to inbreeding, population viability may decline as a result of the accumulation of mildly deleterious mutations after a population has colonized the new environment. Muller (1964) first pointed out that the accumulation of deleterious mutations may reduce individual fitness in asexual populations. However, because of the feedback between the effects of mutational load and genetic drift, it is theoretically plausible that mutations will also accumulate in finite sexual populations. Over hundreds of generations, these mutations can contribute to a decline in population fitness, effectively increasing maladaptation through a process called mutational meltdown. Lynch et al. (1995) indicates that populations with Ne less than 100 are highly vulnerable to extinction because of mutational meltdown, and Lande (1993) suggests that Ne should be more than 5,000 to avoid its effects on fitness and simultaneously maintain sufficient adaptive variation. In theory, then, small populations may diminish a population’s response to an environmental change through either the lack of adaptive genetic variation or reduced fitness caused by exposure of partially recessive mutations through inbreeding and the fixation of mildly deleterious mutations. All three of these evolutionary processes reduce fitness in small populations, although their relative importance depends on factors such as population size, life history, and mating system. Despite their potential importance for population viability, there are few studies of the effects of population size on fitness. Ouborg and Van Treuren (1995) examined the fitness differences between populations of Salvia pratensis in a common environment. Although they did observe large population differences in seed size, germination, and plant growth, no effects could be attributed to differences in population size. In contrast, Fischer and Matthies (1998a) found a positive correlation between population size and seed production per plant. They suggest that the fitness differences are the result of genetic causes. Unfortunately, it is not possible to separate the effects of inbreeding, low diversity, and mutational meltdown. Clearly, more work is needed to establish whether and when low genetic diversity is likely to depress population growth and persistence and to distinguish the effects of reduced genetic variance from those of inbreeding depression and mutation accumulation.
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Ecological Constraints Next we turn to ecological constraints and examine how they may affect a population’s response to a new environment.
Magnitude of Maladaptation The potential for an evolutionary response in a novel environment depends on not only the magnitude of genetic variation but also the degree of maladaptation experienced by a population. Gomulkiewicz and Holt (1995) and Bürger and Lynch (1995) both showed that the capacity for an evolutionary response can significantly increase the time to extinction, but only at moderate degrees of maladaptation. Beyond a certain level of maladaptation, the rate of adaptive response is too slow relative to the demographic cost, and extinction is inevitable. Therefore, regardless of the existing pool of genetic variation, an evolutionary response can avoid extinction only if the new environment is not too extreme relative to that of the original source population. Furthermore, the likelihood of extinction also rises as the temporal variance in the environment increases (Bürger and Lynch 1995). This reinforces the important point made previously by Lande (1988, 1993) that in general, environmental stochasticity is a more important source of vulnerability than genetic diversity. These theoretical predictions may explain the patterns of extinction in colonizing species such as Eichhornia paniculata, an aquatic plant in northeastern Brazil. Barrett and Husband (1997) followed the fate of 22 populations over a 3-year period. In 1 year, nearly one-half of all populations became absent; however, the likelihood of extinction was independent of the initial genetic variation in the populations. Barrett and Husband (1997) suggested that the environmental variation was sufficiently large to cause extinctions, regardless of the genetic variance. In other words, the shift in environment was so extreme and rapid that it precluded any evolutionary response by the population, and no plants survived. Examples of the effects of herbicide applications and heavy metal contamination provide a more optimistic view. In both cases, populations of plants have persisted in extreme environments with high measures of maladaptation (Table 12.2) and in some cases have evolved resistance or tolerance in a very short period of time. Although there are many examples of species that could not elicit such a response, these studies suggest that
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under some circumstances evolutionary responses are possible even in the face of extreme maladaptation. The effect of the initial maladaptation on a population’s ability to adapt is made more complicated by the fact that the degree of maladaptation undoubtedly will change over time as the abiotic and biotic environment changes. Environmental change may arise through modifications to the conditions external to the population (e.g., global change) or through changes induced by the founding of a new population itself and its subsequent establishment. For example, Olivieri et al. (1990) suggested that selection on traits such as autofertility and dispersal ability may change with the age of a population since colonization. They describe how in populations of Carduus, which exhibit a seed dispersal polymorphism, older populations tend to have a higher proportion of nondispersed seeds than young populations. This pattern emerges because selection during colonization favors plants that produce dispersed seeds, whereas selection later in establishment favors individuals whose seeds remain within the population. At the metapopulation scale, the frequency of these two kinds of seeds is determined by the frequency of extinction and colonization. Regardless of the specific character, changes in selection pressures are likely to increase the lag between the optimal and actual phenotypes in population and therefore will decrease the time to extinction (Lynch and Lande 1993; Bürger and Lynch 1995).
Population Size and Life History The initial size of a founding population can also determine the likelihood of an evolutionary response to a new environment. Gomulkiewicz and Holt (1995) argued that the larger the founding population, the higher the chance of persistence. Assuming there will be a demographic cost to selection in the new environment, large populations can decline longer before reaching the critical population size below which extinction risks are increased. In this sense, starting with a large initial population size effectively buys time for the evolutionary response to occur, which is necessary to rescue the population from extinction. A corollary of this observation is that organisms with high intrinsic capacity for population growth should be less likely to become extinct (Chapin et al. 1993; Bürger and Lynch 1995) because they will experience a shorter demographic lag before the population can be rescued by an evolutionary response. This theoretical argu-
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ment may explain why so many annual or short-lived species are so successful as colonizers.
Physiological Acclimation Plant populations may meet the challenges posed by novel selective environments by acclimating physiologically or morphologically. These responses can be rapid and may be particularly critical for enduring the initial stages of colonization. In species with very plastic phenotypes that are able to tolerate a wide range of environments with little or no fitness cost, this response may be sufficient for ensuring the viability of a population. Plasticity allows a plant to respond rapidly to maladaptation and thereby obviate an evolutionary response. In contrast, species that are specialists or are suited to a very specific environment may be more sensitive to change. Several authors (Bradshaw and Hardwick 1989; Via 1994) have proposed that generalist genotypes, able to cope with a wide range of environments, are more likely to evolve in populations that experience environmental changes, especially of a predictable nature. Ultimately, a better understanding of the distribution of “all-purpose genotypes” among taxa and the ecological circumstances favoring their evolution is needed to reliably identify species for which plasticity and broad ecological tolerances can be expected.
Implications for Ex Situ Conservation Conservation of wild plants has been focused for the most part on in situ programs, which are designed to ensure the long-term viability of existing populations. Much of this work has relied on scientific research to identify the genetic and ecological factors that place rare species at risk (Soulé and Simberloff 1986; Barrett and Kohn 1991) and to select strategies to reduce their probability of extinction (Falk and Holsinger 1991). Although interest in habitat preservation and protection of threatened populations is paramount in conservation, the rising frequency of species in immediate peril and the need for restorative actions necessitate that conservation off site increasingly be used to complement or directly aid in situ efforts (Falk 1987). However, programs of ex situ conservation have been slow to incorporate the principles of evolutionary biology to guide strategies for collecting, maintaining, and using plantings for recovery efforts (Guerrant 1992, 1996).
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There is little argument that populations of plants transferred to and maintained in ex situ collections will experience a shift in external environment. Over time, the new environment will impose selection on the collection, favoring adapted genotypes over maladapted ones. The result will be genetic change in the collection over time and in subsequent generations. Although it is intuitive to imagine this shift operating in collections of growing plants (e.g., in a botanical garden), the same processes could also operate on plants stored dormant in seed banks. Seeds differ in the strength of their dormancy and tolerance to storage conditions. Therefore, it is likely that genetic changes will occur in a seed bank over time. Because many seed collections must be regenerated on a regular basis, these stored populations will pass through as many or more generations as plants that are grown out. As a result, the risks of adaptive changes in seed physiology are high. Whether selection operating on seeds during dormancy has any influence on adaptive changes in growth and physiology of actively growing plants is less clear. The strategies for managing abrupt shifts in environment when establishing ex situ collections and in situ reintroductions can conflict. Both conservation goals are similar in that they seek to ameliorate differences between native and foreign environments in order to minimize the degree of maladaptation and hence selection pressures on reintroduced populations. However, the goals for managing the existing maladaptation can also be at odds with each other. Whereas ex situ conservationists are concerned with minimizing an adaptive response to the off-site environment to retain potentially adaptive variation, reintroduction efforts should favor actions that promote rapid evolutionary responses to overcome maladaptation and thus ensure high population survival (Guerrant 1996). Even in ex situ conservation the goals can conflict. Attempts to minimize adaptation to the ex situ environment can help to maintain genetic variability, but this same action may impose a continual state of maladaptation and thereby jeopardize the long-term viability of the collection. How, then, can knowledge of the genetics and demography of populations in novel environments help to better manage ex situ collections and balance the paradoxical goals of ex situ and in situ conservation? Arguably, the most obvious genetic constraint on the value of an ex situ conservation program is a lack of adaptive variation. This constraint arises initially from genetic drift, an inevitable consequence of collecting a finite sample from natural populations and also from drift and unavoidable adap-
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tation during storage and regeneration of the ex situ collection. The only way to offset these processes is to maximize the effective population size (Ne) during all phases of the ex situ program. This would not only minimize the loss of genetic diversity but also ameliorate the effects of inbreeding depression and mutational meltdown, which are also associated with small effective population size. In reality, however, the size of the collection probably will be constrained by the limited source material available for most endangered plants and the limited physical resources available to devote to each species at risk. In sampling material for ex situ collections, the ultimate goal is to capture potential adaptive variation that is representative of the species. But what is representative and how can this be achieved without further endangering the populations sampled? Hawkes (1987) argued that the aim should be to collect as much genetic diversity in a species as possible. However, Brown and Briggs (1991) suggest that, with limited resources for the maintenance of such collections, a more pragmatic approach is needed. They advocated that minimum targets should be established to guide sampling; additional sampling should be conducted only when resources and material will allow it. Our survey of past attempts at transplanting species at risk is consistent with Brown and Briggs’s (1991) observation that in most cases the sources of endangered plants are extremely limited and may be placed at further risk by overcollecting. Furthermore, we caution that the viability of an ex situ collection and its utility for later restoration depend primarily on ecologically significant variation. Therefore, emphasis should be placed on collecting material from distinct individuals at different times and from different habitats. For these reasons, research designed to quantify the spatial organization of quantitative genetic variability and tests for ecological differentiation in endangered species would provide useful guidelines for sampling. Until speciesspecific information is available, sampling procedures should be guided by the near consensus that ecological differentiation is widespread in plants and is of substantial magnitude, particularly in human-influenced environments. In the ex situ collection itself, effective population size can be maximized by ensuring equal reproductive contributions from all accessions when regenerating the collection (Frankel et al. 1995). The approach has been widely discussed in recent years, but is not without its costs. First, it requires that individuals taken from the original populations be maintained separately in the collection rather than as a bulk sample so as to keep track
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of offspring production. Furthermore, the need for equal contributions from all individuals necessitates a controlled breeding program, which is labor intensive and expensive. Finally, as Schoen et al. (1998) have pointed out, any breeding program that maximizes reproductive contributions minimizes not only the magnitude of genetic drift but also the intensity of selection operating within the collection. This may be undesirable to the extent that individuals with deleterious mutations, acquired from the original populations or later on during the maintenance of the collection, are just as likely to contribute offspring to the next generation. Although the general impact of mutation accumulation has not been thoroughly demonstrated empirically for plant populations, Schoen et al. (1998) show that significant numbers of mutations can accumulate within 25–50 regeneration cycles in an ex situ collection. This prospect highlights the need to consider a mild selective regime, perhaps ensuring contributions from all accessions but weighted by the proportion of seeds produced, for long-term viability (Lynch et al. 1995; Schoen et al. 1998). A major concern for managers of ex situ collections and captive breeding programs is the negative impact of inbreeding (i.e., inbreeding depression). Inbreeding also is a function of population size and so can become exaggerated in small collections. Two strategies have been considered to ameliorate its effects: controlled breeding programs that minimize inbreeding and purging of deleterious mutations through an intensive program of inbreeding (Templeton and Read 1984; Templeton 1991). The latter method rests on the assumption that inbreeding depression can be successfully purged from populations with continuous inbreeding. However, research on inbreeding depression in plants and theoretical studies of mutation accumulation suggest that such a program will be ineffective in the long term for two reasons. First, cross-sectional and longitudinal studies of plant taxa that differ in history of inbreeding indicate that purging is inconsistent at best and completely undetectable in a meta-analysis (Byers and Waller 1999). This may occur because most individual deleterious mutations are of mild effect and therefore difficult to purge through a controlled breeding program. Second, even if the mutations in the initial collection could be selectively purged, these would soon be replaced by new mutations that arise within the collection. Those that become fixed through sampling error would be impossible to remove, resulting in a steady decline in the fitness of the collection.
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Even with equal reproductive contributions enforced in the collection, some adaptive responses to the ex situ environment are inevitable through differential mortality. For this reason it may be critical to manage not only the response to novel environments but also the novel environment itself. In general, the more extreme the ex situ environment relative to the native environment, the greater the maladaptation and hence selective pressures acting on the collection. This circumstance would arise in collections that are well outside the native habitat of the species at risk. Extreme environments may impose stronger selection pressures on the ex situ collection and thereby increase the likelihood of losing potentially valuable adaptive variation and allowing the collection to become locally adapted or, worse, creating a degree of maladaptation so severe that no evolutionary response will be rapid enough to rescue the population. In general, ex situ environments are considered benign because of control over the availability of water, nutrients, pests, and possibly temperature. In such conditions, most individuals survive, thereby reducing the intensity of selection on adaptive variation and deleterious mutations. The importance of the environment on patterns of selection of inherently deleterious alleles has been shown in experimental studies of inbreeding depression. Previous studies have shown that selection against inbred individuals (i.e., inbreeding depression) can be significantly higher in harsh environments (Dudash 1990; Holtsford and Ellstrand 1990). In less harsh conditions, the differences between inbred and outbred individuals are not revealed, and mutations are not selectively removed. Maintaining an ex situ collection in a way that maintains potentially valuable variation while eliminating uniformly deleterious mutations is an extremely difficult balance to achieve, particularly when some deleterious variation may be beneficial under different circumstances. It is perhaps for this reason that the idea of inter-situ collections has become more popular. If an off-site collection is maintained within the natural habitat of the endangered species, selective conditions favor the elimination of deleterious variants and favor mutations that are likely to increase population viability in the typical range of environments. However, inter-situ approaches are largely untested, and controlled experiments and careful monitoring of adaptive variation in inter-situ collections will be needed to confirm their value.
Conclusions We’ve discussed some of the challenges involved in retaining sufficient adaptive potential in the ex situ collection, but how should this material be
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used in the restoration effort? Several approaches may increase the utility of ex situ material. First, the ex situ collection must be capable of generating a large number of plant propagules simultaneously. Theoretical studies indicate that the initial size of a transplanted population must be large to accommodate the high demographic cost imposed by maladaptation in the new environment (Gomulkiewicz and Holt 1995). This is especially important for plants with low intrinsic capacities for population growth, such as many trees and shrubs and plants of competitive environments. In other words, conservationists engaged in off-site planting or reintroductions should expect and plan for mortality rather than try to prevent it altogether. In fact, attempting to prevent mortality only slows the adaptive response, which in turn jeopardizes the long-term success of the restoration effort. Our survey of maladaptation is clear: local genetic differentiation between populations is widespread in plants. As a result, plants introduced into a new location undoubtedly will be suboptimal for the environment and will suffer reduced survival and reproduction as a result. Therefore, every effort should be made to reduce the initial maladaptation experienced by the transplants. This can be achieved by using material that is as variable as possible and selecting planting locations that most closely resemble the habitats of origin. The latter step requires careful records on the sources of plant material in the ex situ collection and descriptions of habitat and microclimate needs of the source material. Finally, if populations of any given species are likely to be genetically differentiated, then to achieve the fastest adaptive response to a new environment it is best to transplant material from the collection once at the outset rather than introducing new material to the reintroduced population over time. Theoretically, migration into a population will slow the adaptive response and reduce the time to extinction by introducing additional maladapted genotypes. For long-term viability, ex situ programs should be designed and conducted with scientifically sound information. Although the genetic and demographic factors influencing a population’s response to novel environments are well established, their relative importance in the context of conservation remains unclear. As a result, there are few general guidelines for conservation of endangered plants, and efforts usually are conducted in isolation from the experiences of others. To accelerate the application of the principles of population biology, we suggest the following two actions. First, strategies adopted in ex situ conservation should be tightly linked to serve the end use intended for the collection. Obviously, if the ex situ collection is to be used as source material for restoration, the practices will be
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different and more complex than if they are established as demonstration or educational material. Like others, we caution against the collection of endangered plants for their own sake, without a clear plan of action. Second, whenever possible, ex situ conservation should be practiced within an experimental framework (e.g., Edmunson et al. 1984; Parenti and Guerrant 1990; Drayton and Primack 2000). Often, ex situ programs are developed and maintained in isolation from one another and often using a prescriptive approach, in which a particular strategy is adopted with few provisions built in to evaluate the program along the way. An adaptive management strategy that uses an experimental framework provides a mechanism for evaluating management options, however tenuous, and, more importantly, guarantees a documented outcome that can be applied to other organisms of concern to the conservation community.
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Chapter 13
Population Genetic Issues in Ex Situ Plant Conservation Barbara Schaal and Wesley J. Leverich
One of the most difficult tasks conservation biologists face is evaluating the many scientific issues that surround the establishment and management of ex situ conservation programs. Numerous disparate factors affect the long-term survival of ex situ populations, including ecological, evolutionary, anthropogenic, and genetic issues. Population genetics plays a key role in the ultimate success of ex situ conservation efforts (Soulé 1986). The genetics of both the source population from which genotypes are obtained and the ex situ population itself must be considered when one devises a conservation scheme. Such factors as the levels and distribution of genetic variation within and between populations, plant mating system, and population size are all of critical importance in determining the success of ex situ conservation (Fenster and Dudash 1994; Montalvo et al. 1997; Lande 1999). In this chapter we evaluate some of the genetic issues surrounding the establishment of ex situ populations. First we consider the effects of geographic provenance and small population size on the genetic quality of source populations. Then we examine the effects of inbreeding and outbreeding depression on ex situ populations, and finally we suggest Wallace’s concept of hard and soft selection as a framework for evaluating potentially dysgenic effects.
The Genetics of Source Populations Most plant species that are candidates for ex situ conservation, by definition, are either threatened or endangered. Therefore, the source populations for establishing ex situ programs may have been altered in the 267
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recent past, and these populations could currently have a genetic structure or mating system different from the structure of the species over much of its evolutionary history. Because most at-risk plant species become threatened by habitat loss or fragmentation, we expect that some of the source populations for ex situ programs will have an altered genetic structure resulting from small population size caused by habitat loss (Young et al. 1996). On a theoretical basis, species that have experienced severe habitat loss are expected to undergo several changes (Jain 1994; Ellstrand and Elam 1993; Vucetich and Waite 1999). First, metapopulation structure is altered (Gliddon and Goudet 1994). Because of habitat loss, subpopulations that were previously in contact with other subpopulations often become isolated. Isolation is a consequence of reduced gene flow between populations and can result from lack of pollination or seed dispersal (Ellstrand 1992). Although one usually expects increased genetic isolation as a consequence of habitat fragmentation, it is important to note that in a few cases, gene flow may actually increase as barriers to migration are removed. Genetic isolation often is coupled with a decline in subpopulation size. In the case of very small populations, species that previously had an outcrossing mating system may forced to inbreed (DeMauro 1993) in some cases, accompanied by a breakdown of self-incompatibility systems (Reinartz and Les 1994). Together, these processes of increasing isolation and declining population size are expected to enhance the role of genetic drift (Ewens 1979). Drift decreases genetic variation within populations; populations should lose alleles, and the genotypic distributions should shift towards homozygosity. At the same time, random fixation of alleles by genetic drift increases the genetic differences between populations. Although one expects these changes on a theoretical basis, how often are they observed? Interestingly, many plants that have small population sizes do not seem to experience these expected genetic changes (Wolf et al. 2000; Podolsky 2001; Cruzan 2001; but see Morgan 1999). Some of these plant species may be intrinsically rare, with small population sizes (Rabinowitz et al. 1984), and may be adapted to their rare status (Lammi et al. 1999). Other species may have only recently experienced population declines through habitat loss or fragmentation. Annual plants should quickly accumulate the genetic changes associated with small population size because of their short generation times. But many of the plant species that are at greatest risk
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are plants of stable communities such as prairies or climax forests. These species may be extremely long lived. Because of these long generation times, we might expect a time lag before the genetic effects of small population size could be detected. In long-lived plants, most standing individuals in a site were established long ago, under environmental conditions different from those observed in the present population. One may expect only future generations, represented by seeds and juveniles, to show some of the effects of habitat loss and altered mating patterns. This phenomenon explains, in part, the common observation that relictual plant populations often show high levels of heterozygosity and genotypic variation, despite being quite small (Hamrick and Godt 1990). We expect to see the genetic effects of drift, inbreeding, and increased variation between populations only among recent progeny and then only weakly in recently threatened species. The effects of genetic drift accumulate over generations; plants with longer generation times show fewer immediate effects of drift. If there is little recruitment of seedlings into a population, the mature individuals at a site may all have been established before any population decline, so few or no generations have been produced since the species decline. Demographic consequences of habitat loss appear rapidly; the genetic consequences may accumulate much more slowly. Thus, one is left with the disturbing conclusion that current measurements of genetic diversity in some threatened species may be poor predictors of ultimate genetic decline. Nonetheless, the pool of adults in a population may be strongly affected by habitat loss, and the number of individuals in populations often declines. Even with little or no seedling recruitment (as observed in many stable communities), the genetic structure can be affected by the loss of individuals. Such species experience the population genetics of subtraction. Without substantial recruitment, gene and genotypic frequencies are altered only by individuals leaving the population or possibly by the accumulation of mutations (Lynch et al. 1995; Lande 1995). For example, many threatened prairie plants have no detectable seedling establishment (Bowles et al. 1998; Van der Valk 1978). Such populations can only lose genotypes, and this subtraction ultimately alters the gene and genotypic frequencies within the population. In some sense this process mimics drift. Populations may have reduced within-population variation (particularly so for vegetatively reproducing plants, as discussed later in this chapter), whereas genetic variance increases between populations through chance
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changes in allele and genotypic frequencies. These changes, in turn, can affect ex situ conservation efforts because they influence the distribution of alleles and genotypes available for ex situ conservation.
Small Population Size in Asclepias meadii A. meadii, a threatened species of Midwestern tallgrass prairies, is a good example of such genetics of subtraction. A. meadii is a long-lived perennial and reproduces by both obligate outcrossing and clonal spread via rhizomes. Like so many other U.S. native prairie plants, A. meadii has undergone severe habitat loss with the conversion of native prairies to agricultural fields. Samuel Mead, who first described the species, noted that by the late 1800s the species was already becoming increasingly difficult to find. A. meadii no longer occurs throughout much of its former range, and remaining populations are few and isolated (Bowles et al. 1998). Mead’s milkweed is a good example of a species whose genetic variation has been profoundly influenced by human activities. The genetic consequences of isolation were explored by a random amplified polymorphic DNA (RAPD) analysis of genetic variation and clonal structure. RAPD profiles were used to characterize individual genotypes and to determine the size and number of genotypes within populations (Hayworth et al. 2001). Populations of Mead’s milkweed are managed by mowing or burning. Sites are burned in winter or early spring, which allows plants to complete their cycle of sexual reproduction. Reproduction in these populations occurs by vegetative spread of established genotypes or by sexual reproduction via seed. In contrast, mowing occurs in June and removes the flowers or developing seeds of plants. Reproduction in these populations is primarily clonal (McGregor 1977). These contrasting management regimes, applied for decades at some sites, have led to differences in the demography and reproduction of populations (Betz 1989; Bowles et al. 1998). The burned sites are assumed to be more representative of pre-European conditions, whereas the mowed sites represent an alteration of habitat brought about by agriculture. Table 13.1 presents estimates of the number of plants (ramets) within a population, the number of variable markers within a population, and an estimate of the number of genotypes (clones). Two interesting aspects of the clonal structure of these populations are apparent. First, there is little relationship between the total number of plants (ramets) and the number of clones. In some mowed populations the ramet number can be high and
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table 13.1 Variation and clonal structure in Asclepias meadii. Population Location
Number of Variable Markers (random amplified polymorphic DNA)
Number of Clones
53 31 38 6 34 3 83
11 9 0 10 11 11 10 6.7
3 3 3 2 10 2 5 3.7
24 90 16 58
12 15 9 10 11.5
10 21 5 11 11.75
Number of Plants (ramets)
Mowed Populations Douglas, Kansas Barton, Missouri Jefferson, Kansas Franklin, Kansas Anderson, Kansas Bourbon, Kansas Miami, Kansas Mean values Burned Populations Dade, Missouri Jefferson, Kansas Saline, Illinois Iron, Missouri Mean values Source: Hayworth et al. (2001).
the number of clones low; the Douglas, Kansas, population has 53 plants representing only three genotypes. Second, the number of genotypes is higher in burned populations than mowed populations; mowed populations have a mean clone number of 3.7, whereas burned populations have an average of 11.75 clones. These genetic data are consistent with the hypothesis that few or no new genotypes are being recruited into mowed populations and that the remaining genotypes spread by vegetative reproduction (Tecic et al. 1998). Moreover, the data suggest that genotypes are being lost from the mown populations. The loss of genetic diversity can be devastating for A. meadii populations. Mead’s milkweed is self-incompatible, and in several of the populations with low numbers of genotypes, seed set is no longer observed even when plants from the population are crossed manually. However, seed set occurs in these populations if pollen is brought in from other populations (Bowles, pers. comm., 1996). Presumably the lack of seed set is caused by loss of allelic diversity at the self-incompatibility locus. In contrast, populations with larger numbers of genotypes routinely set seed. These diversity data for A. meadii suggest that in mowed populations genetic variation has declined because of both loss of genotypes and the lack of new genotypes
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being recruited into populations. Moreover, this decline in genetic diversity apparently has led to a direct negative consequence: the loss of the ability to set seed. Next we can examine the A. meadii data to determine whether there is evidence of a driftlike process. The effects of genetic drift are to increase the genetic variance between populations while decreasing the genotypic variance within populations. In A. meadii the loss of genotypes in mowed populations should result in a distribution of variation analogous to that expected from drift. A driftlike process would occur if mortality is influenced predominantly by environmental rather than genetic factors. Burned populations should be much less affected because they can add new genotypes via sexual reproduction. Indeed, genetic variances within populations, estimated by number of variable RAPD markers, are higher in burned and less in mown populations (11.5 vs. 6.7, p < .05). As expected, between-population variation is higher for mown populations than for burned (19.6 vs. 5.25, p < .05) What do these results mean for ex situ conservation efforts? Sampling seed of A. meadii for ex situ conservation is not straightforward. Seeds collected from only one population may have very low diversity, particularly if the site is managed by mowing. Even the slightest subsequent decline in diversity at the ex situ site may lead to a loss of seed set. An alternative strategy would be to collect seeds from several populations and mix genotypes from different geographic regions to provide seeds for restoration. A mixture would ensure adequate genetic variation to compensate for the lack of any within-population variation in self-incompatibility alleles. However, such mixing of genotypes from different geographic populations leads to additional population genetic questions.
The Geographic Provenance of Seed Sources: Ecotypic Differentiation One of the major issues for ex situ conservation and restoration is the selection of specific source populations (Primack 1996). How geographically close to the ex situ site should the source population be (Montalvo et al. 1997)? Should more than one source population be used, or is it harmful to mix genotypes from separate populations? Determining the nature of geographic constraints on the location of plant or seed sources has been a vexing problem for restoration and conservation (Frankel et al. 1995). For
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example, restoration managers are concerned that seed collected from a source population be well adapted to the new site (Montalvo et al. 1997). Effective conservation must consider the geographic provenance and adaptation of source plants. Ecotypic adaptation may make some locations unsuitable as a seed source. For example, prairie grasses collected from Canada may fare poorly in south Texas (McMillan 1964). In some cases, the concern for local adaptation has led to stringent criteria that have nearly precluded restoration efforts (e.g., rules that seeds can be collected only from within 1, 5, or 50 miles of the site of restoration). Likewise, the mixing of genotypes from different populations has often been avoided, based on concern for “contaminating the gene pool” or the potentially dysgenic effects of mixing genotypes adapted to different environments. Are these practices justified, or could the criteria for collection of source material be relaxed? Any easing of such criteria would offer many more options for ex situ conservation efforts and for restoration. How sound is the basis of our concern about the geographic provenance of source populations? Many ideas on precise geographic adaptation come from common garden reciprocal transplant experiments that have been used to study ecotypic differentiation. The classic work of Clausen, Keck, and Hiesey in the 1940s demonstrated geographic and altitudinal adaptation of populations within some plant species (Clausen et al. 1940, 1948; Clausen and Hiesey 1958). Moreover, when plants were moved to an environment different from the native environment, individuals often showed lower viability or seed set, and in some cases greater mortality. Subsequent work done in the 1950s and 1960s by Calvin McMillan has shown phenotypic and life history adaptation associated with latitudinal and longitudinal variation in prairie grass species. McMillan collected grasses along a latitudinal transect from Canada to Mexico and then grew the grasses at several sites (McMillan 1959, 1964, 1969). These common garden transplant experiments indicated broad-scale latitudinal differentiation for key life history traits such as flowering time and survival. Genetic variation of some 60 days in initial flowering time and dramatic differences in survivorship were observed (Tables 13.2 and 13.3). McMillan’s studies convincingly demonstrate geographic adaptation and the harmful effects of growing genotypes in environments for which they are not suited. From these studies alone we would conclude that great care must be taken in choosing a source population because of local adaptation. But transplant studies such as these deserve greater scrutiny.
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table 13.2 Survival of clones transplanted to Austin, Texas. Number Surviving (Number Transplanted) Transplant Source Location
A. scoparius
A. gerardii
P. virgatum
S. nutans
North Dakota Nebraska Texas Arizona Mexico
0 (12) 2 (12) 12 (12) 0 (10) —
0 (12) 0 (12) 12 (12) 6 (6) 0 (5)
0 (6) 11 (12) 12 (12) 0 (12) 0 (9)
— 0 (12) 12 (12) 11 (12) 0 (11)
Source: McMillan (1969).
table 13.3 Flowering date for Bouteloua gracilis in a common garden. Transplant Source Location
Manitoba North Dakota Northern Nebraska Southern Nebraska Colorado Kansas Oklahoma Texas
Flowering Date (from June 1)
12 12 8 5 11 33 44 54
Source: McMillan (1964).
McMillan used clonal transplants, adult individuals from the field. Presumably these genotypes have been through a selective sieve of seed development, germination, seedling establishment, and growth to adulthood. Thus, the field plants are the end product of competition and selection in the native habitat. Geographic adaptation clearly has occurred in these grasses, but how precise is it? Despite adaptation, there was significant variation in survivorship (Table 13.2) and time of flowering (Table 13.3) among the genotypes collected from a given location and then planted in Texas. Moreover, we would expect that far greater variation should occur among the genotypes represented by seeds than among the adults. Seed genotypes are the product of meiotic recombination and independent assortment, and they have not yet been winnowed by selection in the native environment. Offspring from single plants should show a much wider range of flowering times and survivorship than do their parent plants in the field.
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table 13.4 Survivorship and reproduction of an identical set of Helianthus tuberosus clones. Transplant Site
1 2 3 4
Survivorship to Day 200
Reproduction (no. of ramets)
54% 44% 97% 58%
3.27 1.39 25.42 1.12
Source: Zampini (1991).
One of the most interesting conclusions from the extensive work on common garden studies in plants is the variable role of phenotypic plasticity. Phenotypic plasticity, variation in phenotype of a single genotype in response to environmental variability, may be a major way of adapting some plant species to different environments (Schlichting 1986; Scheiner 1993). The degree of phenotypic plasticity of a species will also affect the ultimate success of an ex situ conservation program by adaptively modifying the phenotype of a specific genotype. Population differentiation and phenotypic plasticity were studied in Helianthus tuberosus transplants (Zampini 1991). A large transplant experiment looked at quantitative fitness traits such as vegetative and sexual reproduction and survivorship (Zampini 1991). Identical clones from different source populations were planted in four distinct environments in the greater St. Louis area. Mean survivorship and reproduction varied widely. Although populations were genetically differentiated from each other for these traits, most of the variation in phenotype resulted from phenotypic plasticity in response to the variable environments. The effect of genotype often was insignificant or small. Table 13.4 lists a subset of these data. Vegetative reproduction of an identical set of clones planted in different environments varied from 1.12 ramets per plant on average to 25.42; likewise, survivorship varied from 44 to 97 percent. These results are hardly surprising to anyone who has grown plants in the field. Phenotypically plastic responses to environmental variation can easily overshadow any underlying genetic differences. Genetic differences between source populations may in some cases be tempered by phenotypic plasticity. We can also consider indirect genetic information on the degree of geographic adaptation. What evidence is there for geographic differentiation when we look at neutral marker genes? If there is strong genome-wide
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selection for local adaptation, one would assume that marker genes would reflect divergent selection (geographic differentiation can also result from limited gene flow or geographic sorting of lineages). There are numerous examples of geographic differentiation within plants for marker loci, such as allozymes or DNA haplotypes (Tecic et al. 1998; Matos and Schaal 2000). But when one considers total genetic diversity, most of the variation within plants results from difference between individuals within a population, even in species that exhibit strong geographic differentiation (Lacerda et al. 2001). Typically, 5 to 50 percent of the variability within a plant species is the result of differentiation between populations. Thus, for a majority of plant species, most of the genetic diversity within a species results from differences between the genotypes within a population rather than from genetic differences between populations (Hamrick et al. 1991). Such surveys of genetic variability across species ranges and the analysis of patterns of genetic diversity are potentially useful in addressing the question of seed sources. Although marker loci such as allozymes or RAPDs cannot reliably predict differentiation at selected loci or indicate quantitative genetic variation (Frankham 1999; Reed and Frankham 2001), they do provide an estimation of population differentiation. If a species shows very little geographic differentiation between populations, there is probably small risk in mixing seed sources. On the other hand, if genetic analyses indicate a large between-population component of total genetic diversity, then one should be more cautious in mixing seed sources. In general, a conservative approach to seed mixing is prudent. One can always add to a restoration seeds from another source if necessary, whereas it is nearly impossible to remove genotypes once introduced (Holsinger, pers. comm., 2000). An increasing number of threatened and endangered plant species have been analyzed for patterns of allozyme or DNA variation, and these data will be useful in practical decisions about ex situ conservation. This argument points to the value of broad allozyme or DNA marker surveys for plant species that are of conservation interest.
Genetics of Ex Situ Populations Until now we have been concerned with the genetics of source populations. Once seeds or plants are collected and established at an ex situ site, the genetics of the new ex situ population must be addressed. Many of the issues that affected the source population’s genetics, such as inbreeding
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and genetic drift, can also be important in the new population (Frankham 1995). Moreover, if the source population is from a different geographic provenance than the conservation site, adaptation and selection play an important role as well. Finally, if seeds from several sources are used to establish the ex situ population, one needs to worry about the effects of crossing disparate genomes, or outbreeding depression (Beyers 1998). In many ex situ programs, population size may be small, which leads to concerns about inbreeding (Frankham and Ralls 1998; Hedrick and Kalinowski 2000). Inbreeding depression has variable effects on ex situ conservation programs. Inbreeding depression increases homozygosity and has potential negative effects on progeny fitness. At an ex situ site, inbreeding will result from small numbers of genotypes, both from the initial source population and in the ex situ population. Inbreeding depression is a documented phenomenon in many plants but does not affect all species equally. If the species is naturally an inbreeder, if a species has a mixed mating system (the species produces seed by both selfing and outcrossing), or if genetic load has been purged, inbreeding will be more easily tolerated with little negative effect on fitness (Beyers and Waller 1999; Frankham 1995). Likewise, if the ex situ conservation effort is very large, with both large numbers of plants or seeds for establishment and large population sizes maintained after establishment, inbreeding will be less of an issue. But in the case of outbreeding plants or small ex situ populations, inbreeding may well affect the viability of progeny.
Inbreeding Depression: Lupinus texensis L. texensis is an endemic annual species of Texas. It typically forms very large populations and has a mixed mating system (Helenurm and Schaal 1996). The effect of inbreeding depression was explored by germination of field-collected seed from 27 populations representing the range of the species and by selfing and outcrossing of 30 plants per population (Schaal 1989). L. texensis typically experiences inbreeding depression in the greenhouse. For example, in one population the mean size of plants at 6 weeks of age is smaller for plants produced by selfing (20.8 leaves) than for plants produced by outbreeding (23.4 leaves), an 11.2 percent reduction. Likewise, plants show inbreeding depression in the number of flowers. Progeny produced by outcrossing have on average 108.6 flowers, whereas selfers have only 88.4 flowers, representing an 18.2 percent reduction in fitness.
ecological/evolutionary context of ex situ conservation
Number of Plants
278
35
Outcrossed
30
Selfed
25 20 15 10 5 0 5
10
15 20 25 30 35 40 45 Number of Leaves per Plant
50
55
Figure 13.1 Plant size at 6 weeks in progeny from outcross and self-pollination.
The number of leaves is an indication of plant size. The totals represent offspring grown from seed resulting from 30 outcrosses and 30 self-pollinations from a single Lupinus texensis population grown in the greenhouse.
These results are typical for Lupinus texensis populations, and they reflect the general outcrossed breeding system of the species. Inbreeding depression of this magnitude is common in outbreeding plants. Although the mean reduction in fitness is of interest, the more relevant information for conservation applications is the distribution of individual fitnesses. That is, where inbreeding depression occurs, how many plants show below-average fitness as a result of inbreeding? How does this mean loss of fitness occur: is the inbred population skewed toward the low fitness end of the distribution, or is the entire distribution shifted? Perhaps the most relevant question is, How many inbred plants have high average fitness? Figures 13.1 and 13.2 show the distribution of leaf numbers and flower numbers for the aforementioned experimental population. For both leaf and flower numbers, inbred progeny show a wide range of fitnesses, and the distributions of inbred and outbred progeny appear quite similar. In fact, 45 percent of selfed progeny have a leaf number greater than the mean leaf number of outbred plants. Likewise, 47 percent of selfed progeny have a flower number greater than the mean for outbred plants. Thus, the loss of fitness associated with inbreeding is not seen uniformly across all
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30 Outcrossed
Number of Plants
25
Selfed
20 15 10 5 0 50 100 150 200 250 300 350 400 450 500 550 600 650 700 750 Number of Flowers per Plant
Figure 13.2 Number of flowers produced by progeny from outcross and self-
pollination. The totals represent plants grown from seed resulting from 30 outcrosses and 30 self-pollinations from Lupinus texensis.
plants. Some plants have strong fitness reductions, whereas other inbred plants have very high fitnesses and seem unaffected. The production of large numbers of plants with high fitness, despite inbreeding, has direct implications for conservation programs.
Ex Situ Conservation: Hard or Soft Selection? The various genetic processes described in this chapter—inbreeding depression, genetic drift, local adaptation, and outbreeding depression— present the conservation manager with a daunting list of genetic concerns. In the design of a conservation management plan, how does one take these issues into consideration and still develop a realistic plan for ex situ conservation? The inbreeding data discussed in this chapter point to some ways to evaluate these issues. Bruce Wallace (1968) attempted to integrate genetics into the ecological regulation of population size. He considered selection to interact with population size in two fundamentally different ways. Selection is “hard” when an individual experiences an absolute loss of fitness because of maladapted or lethal genes. Such selection operates independently of species population density. Selection is “soft” when popula-
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tions are at high densities. The concept of soft selection is based on the carrying capacity of the environment, an ecological idea. An environment’s carrying capacity is limited, so any excess reproduction leads to individuals being eliminated from the population (soft selection). That is, some individuals cannot become established because there is no room in the population for them. This concept seems reasonable for many plants, where large numbers of seeds and seedlings are produced but few become established as adults because of limited space. In essence, soft selection is the result of the excess reproductive capacity of a species, and this concept seems particularly appropriate for plants, which often have large reproductive excesses. Consider the inbreeding data presented earlier for L. texensis. If hard selection is at work, inbreeding depression will cause a loss of individual fitness, will result in a low mean population fitness, and will ultimately result in an unsuccessful conservation effort. If soft selection operates, ex situ populations may be able to tolerate inbreeding depression and, by analogy, outbreeding depression as well. Simply put, not all seeds can mature into adult plants because of the finite carrying capacity of the environment. If the fraction of the population that does not survive comprises mostly individuals of low fitness, inbreeding could be tolerated with little or no reduction in mean population fitness. As an example, assume that a population has a carrying capacity of 40 plants. If 100 seeds are sown, we expect that 60 will not mature. Among the plants that don’t survive would be plants with lethal, developmental abnormalities, and those that experience reduced fitness because of inbreeding depression. This excess reproduction (and, by analogy, excess seed sown) accommodates these genetic losses. The distribution of fitnesses in L. texensis shows clearly that enough highly fit individuals are produced even by inbreeding to easily fill a population. In fact, an analogous process occurs in all natural sexually reproducing plant populations. Sexual reproduction produces new combinations of alleles that yield individuals with a wide range of fitness. The production of some individuals with low fitness does not necessarily affect mean population fitness in most plant species because there usually is an excess of seeds and juveniles. We expect that individual plants that establish, survive, and reproduce should come from the high end of the fitness distribution. Remaining is the question of genetic change in such populations by soft selection. If survivorship is predominantly environmentally deter-
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mined, with only a small genetic component influencing fitness, the genetic structure of soft-selected populations should not change. However, if there is a strong genetic component in determining which individuals survive and reproduce or if there are significant gene-by-environment interactions in determining fitness, then a population’s genetic structure may indeed be altered by soft selection. It should be pointed out that alterations in genetic architecture may occur in any restoration, regardless of how source seeds are selected or how the site is managed. Other “dysgenic” processes could be evaluated in a similar manner. The production of individuals of low fitness by inbreeding, outbreeding depression, or a poorly adapted genotype might be accommodated by excess reproductive capacity. Of course, there are limits to this approach. If a large proportion of individuals have low fitness, then the demographic stability of a population is threatened. But in many cases plant species should be able to tolerate a moderate amount of these dysgenic effects without harming population fitness or restoration efforts because plants within a population exhibit a range of fitnesses (Leverich and Levin 1979). To more precisely evaluate the likelihood of a species tolerating such negative fitness effects, we need to examine not just the mean changes in fitness associated with a particular process but also the distribution of individual fitness. When such additional data are available, we can evaluate whether processes such as outbreeding depression would significantly alter population fitness and whether such processes should be of concern in the design of ex situ conservation strategies.
Conclusions Managers of ex situ conservation efforts have a difficult task. They are faced with a series of environmental, ecological, and genetic issues. In many cases, genetic concerns are based on theoretical considerations, often with little supporting data. In other cases when empirical data are available, they may be contradictory. For example in a recent study of the rare plant Lychnis, there was no correlation between genetic diversity and fitness components (Lammi et al. 1999), whereas a strong association between genetic diversity and fitness was found in eelgrass (Williams 2001). It is not reasonable to expect that every species involved in an ex situ conservation effort will be studied at the same level of detail as these species; ex situ conservation often must proceed with a minimum of data. As a first principle,
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a conservative approach is the most appropriate. When possible, a single large source population from a similar habitat and geographic provenance as the ex situ site is the best possible choice. Likewise, inbreeding and outbreeding depression should be avoided if possible. But if the ideal conservation situation is not present, it is appropriate to proceed with ex situ restoration and not be overly concerned with theoretical population genetic issues that are being debated in the literature. The discussion of hard and soft selection suggests that there may be some flexibility for conservation biologists and that some of the dysgenic effects of small population size may be accommodated in the reproductive excess of most plants. Finally, the overall importance of genetics in relation to the demographic declines associated with habitat loss is a matter of debate (Holsinger et al. 1999). References Betz, R. 1989. Ecology of Mead’s milkweed (Asclepias meadii Torrey). Pages 187–191 in T. Bragg and J. Stubeendieck (eds.), Proceedings of the Eleventh North American Prairie Conference. Lincoln: University of Nebraska. Beyers, D. 1998. Effect of cross proximity on progeny fitness in a rare and a common species of Eupatorium (Asteraceae). American Journal of Botany 85:644–653. Beyers, D., and D. Waller. 1999. Do plant populations purge their genetic load? Effects of population size and mating history on inbreeding depression. Annual Review of Ecology and Systematics 30:479–513. Bowles, M., J. McBride, and R. Betz. 1998. Management and restoration ecology of the federal threatened Mead’s milkweed, Asclepias meadii (Asclepiadeaceae). Annals of the Missouri Botanical Garden 85:110–125. Clausen, J., and W. Hiesey. 1958. Experimental Studies on the Nature of Species. IV. Genetic Structure of Ecological Races. Publication 615. Washington, DC: Carnegie Institute. Clausen, J., D. Keck, and W. Hiesey. 1940. Experimental Studies on the Nature of Species. I. Effect of Varied Environments on Western North American Plants. Publication 520. Washington, DC: Carnegie Institute. Clausen, J., D. Keck, and W. Hiesey. 1948. Experimental Studies on the Nature of Species. III. Environmental Responses of Climatic Races of Achillea. Publication 581. Washington, DC: Carnegie Institute. Cruzan, M. 2001. Population size and fragmentation thresholds for the maintenance of genetic diversity in the herbaceous endemic Scutellaria montana (Lamiaceae). Evolution 55:1569–1580. DeMauro, M. 1993. Relationship of breeding system to rarity in the Lakeside daisy (Hymenoxys acaulis var. glabra). Conservation Biology 7:542–550. Ellstrand, N. 1992. Gene flow by pollen: implications for plant conservation genetics. Oikos 63:77–86.
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Tecic, D., J. McBride, M. Bowles, and D. Nickrent. 1998. Genetic variability in the federal threatened Mead’s milkweed, Asclepias meadii Torrey, as determined by allozyme electrophoresis. Annals of the Missouri Botanical Garden 85:97–109. Van der Valk, A. 1978. The role of seed banks in the vegetation dynamics of prairie glacial marshes. Ecology 59:322–335. Vucetich, J., and T. Waite. 1999. Erosion of heterozygosity in fluctuating populations. Conservation Biology 13:860–868. Wallace, B. 1968. Topics in Population Genetics. New York: Norton. Williams, S. 2001. Reduced genetic diversity in eelgrass transplantations affects both population growth and individual fitness. Ecological Applications 11:1472–1488. Wolf, A., S. Harrison, and J. Hamrick. 2000. Influence of habitat patchiness on genetic diversity and spatial structure of a serpentine endemic plant. Conservation Biology 14:454–463. Young, A., T. Boyle, and T. Brown. 1996. The population genetic consequences of habitat fragmentation for plants. Trends in Ecology and Evolution 1:413–418. Zampini, C. 1991. Genetic and Environmental Sources of Variation in Natural Populations of Helianthus tuberosus L. Doctoral dissertation, Washington University, St. Louis, MO.
Chapter 14
Integrating Quantitative Genetics into Ex Situ Conservation and Restoration Practices Pati Vitt and Kayri Havens
Preserving the genetic diversity in natural populations has long been a focus of conservation biologists. In 1991, Genetics and Conservation of Rare Plants (Falk and Holsinger) proved to be a seminal volume addressing the conservation of genetic diversity in rare plants. Since that time, the use of molecular markers in both plant and animal conservation biology has exploded, largely because of the advent of polymerase chain reaction–based methods such as randomly amplified polymorphic DNA (RAPD). Allozyme markers also remain common, and given their ease and low expense, they remain a useful tool in assessing genetic diversity on a large spatial scale. Using markers to detect diversity within populations has revealed much about the genetic structure of imperiled plant species. Hamrick and Godt (1989) have correlated levels of genetic diversity and life history characteristics across taxa, and have made generalized management recommendations based on these correlations (Hamrick et al. 1991). Nevertheless, in the end we are left questioning whether small populations have enough genetic diversity to withstand catastrophic events, demographic stochasticity, and, indeed, the forces of evolution itself. The fundamental reason we care about the genetic structure of small populations is the desire to ensure that they retain enough diversity to withstand the forces of selection over evolutionary time (Hamrick et al. 1991). Although molecular markers have been used extensively to detect diversity because of both ease and practicality, we argue that investigations into the genetic diversity of rare plant populations should include quantitative 286
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traits. As Frankham (1999) has pointed out, most of the major genetic threats to rare species (i.e., inbreeding depression, loss of evolutionary potential, genetic adaptation to captivity, and outbreeding depression) involve quantitative traits. We also believe that this type of data is important because quantitative traits are directly acted on by natural selection, unlike marker traits, which are generally selectively neutral. In addition, we believe that many species with low levels of neutral trait diversity may be found to have much higher levels of quantitative diversity. If so, new models of population viability may be needed. Ultimately, many species thought to be genetically depauperate may in fact be more resilient in the face of a changing environment.
Quantitative Traits Quantitative traits are those characterized by continuous variation, that is, variation that does not fall into discrete categories. These diffuse traits may include aspects of life history such as time to flowering in annuals, size at reproductive onset in perennial species, and phenological variation in anthesis among individuals in a population. They may also include other types of characters such as variation in length and width of leaves, flower number, or petal size. Variation in such characters may or may not be reflected in allozyme or molecular variation (Holsinger and Vitt 1997; Holsinger 1991). However, variation in quantitative traits is important in interpreting the genetic structure of threatened species. As Darwin noted, variation in these characters thus “afford materials for natural selection to accumulate” (1859: 45). If you consider plants in their natural setting, no two individuals look alike, although they are members of the same species. Some individuals are taller, bushier, or spindlier than others are. Unlike monogenic variation, quantitative variation is the result of a multitude of genes acting in concert to give rise to the particular form you observe, which is the phenotype. In such polygenic characters, the individual genes acting on the phenotype have only a small effect. In contrast, monogenic characters are quite discrete and are controlled by one gene, usually with two or sometimes three loci that have very large effects. For example, the peas that Mendel used to help him define ratios of inheritance had either smooth or wrinkled seeds. This character was inherited in a 3:1 ratio, so only one gene with two forms,
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or loci, was involved. The gene involved in this example had a very large effect on the phenotype, the result of which is easily classified as smooth or wrinkled. The genes involved in quantitative traits are also inherited in the classic Mendelian ratios. However, their effects on the phenotype are quite small, resulting in traits with smooth and continuous distributions across easily defined ranges. Quantitative traits also result from the action of many genes, and there may be two or three forms of each gene represented in a population. Given the number of combinations possible, phenotypic variation of quantitative traits can cover a broad range for many characters. A final difference between most monogenic traits and polygenic traits is that the cumulative expression in the phenotype for quantitative traits is highly dependent on the environment. Quantitative variation must be measured and analyzed in such a way that the purely genetic effects on the phenotype can be distinguished from the environmental effects. It is this last complication that may have limited the application of quantitative traits in surveying the genetic diversity between and within populations of rare plants. Quantitative trait diversity may be measured at the individual (plasticity), family, population, or geographic scale. Traits that may be important to the survival or reproductive success at the individual level, and thus be under selection, have proven the most appropriate for study. Table 14.1 presents a summary of traits that have been measured in many studies, particularly with regard to species of conservation interest.
Heritability of Quantitative Traits Quantitative traits arise from the independent effects of multiple genes acting on a trait. These are called additive genetic effects. Additivity implies that the joint effect on the phenotype reflects the sum of the action of each gene involved in the expression of a particular trait. Additive genetic effects can also occur if the alleles of a gene are additive. This occurs when the heterozygote is intermediate between the two homozygotes. In some instances they arise from the additive effects of both alleles and from multiple genes. One excellent example of this is provided by Nilsson-Ehle (1909), who showed that the intensity of red pigment in the glume of Triticum vulgare is acted on by three unlinked genes, each with two alleles. Individuals that are homozygous recessive across all of the genes have no red pigmentation, whereas homozygous dominant individuals express
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table 14.1 List of traits measured in many quantitative studies. These are traits that have some fitness component, perhaps unmeasured, and represent studies taken from the literature, with a preference for species of conservation concern. This list is not intended as an exhaustive account of traits used in examining plant quantitative genetics. However, it does indicate traits typically used in such studies and illustrates the potential fitness effects of quantitative traits. Stem and Root Characters
Reproductive Characters
Plant growth rate Stem diameter 3,11,26 Stem pubescence19 Number of nodes on stem1 Length between internodes6 Weight of dried stems4,18 Root to shoot ratio21 Total plant mass21,22 Plant height1,6,8,10–13,21,24–27 Time of budburst11
% Flowering plants17 Flower number 2,6–8,16,18,21 Pollen number 10 Pollen tube growth rate9 Stamen length2 Pistil length7 Anther length7 Anther width7 Calyx diameter 16 Corolla diameter 7,12,21 Petal width19,21 Petal length2,19 Florets per head25 Pappus length25 Number of flowering heads17,24 Corolla color12,19 Fertilization rate5 Seed weight1,4,13,18,23,25 Seed diameter 18 Seed shape20,25 Embryo weight23 Cotyledon area3,21 Total seed number 3,13,15 Fruit number 4,12,21 Seed number per fruit5,6,12,21 Seedling survival rates18 % Germination12,15,18 Pine cone production10
4,8,14,22
Sources: 1Andersson (1991); 2Ashman (1999); 3Bennington and McGraw (1996); 4Bonnin et al. (1997); 5Colas et al. (1997); 6Donohue et al. (2001); 7Elle (1998); 8Farris (1988); 9Havens (1994); 10 Hedrick and Savolainen (1996); 11Jaramillo-Correa et al. (2001); 12Kercher and Sytsma (2000); 13 Knapp and Rice (1998); 14Kuittinen et al. (1997); 15Lammi et al. (1999); 16Meagher (1994); 17 Meagher et al. (1978); 18Ouborg and Van Treuren (1995); 19Podolsky (2001); 20Prentice (1992); 21 Schwaegerle and Levin (1991); 22Schwaegerle et al. (2000); 23Thiede (1998); 24Waldmann and Andersson (1998); 25Widen and Andersson (1993); 26Yang et al. (1996); 27Zhong and Qualset (1995).
the greatest pigmentation. Heterozygous individuals express intermediate levels of pigmentation. This model shows simple additivity: the addition of a locus at each of the three genes adds a dimension to the color, independent of any interaction between the genes. Additive genetic effects result
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from the independent action of multiple genes. Epistatic effects arise when there is an interaction between genes. Although it is not clear how many genes or alleles per gene are operating on most quantitative traits, it is possible to measure the transmission of quantitative traits from parent to offspring. In particular, it is possible to determine the additive genetic effects transmitted to offspring. The additive component of genetic variation is called the narrow-sense heritability, denoted as h2. More formally, h2 is the ratio of additive genetic variance to phenotypic variance and expresses the extent to which phenotypes are determined only by the genes transmitted by the parent, excluding environmental effects. In a sense, because this can be transmitted from parent to offspring, the ratio of additive variance to phenotypic variance is a measure of how much variation is available in the population. Therefore, h2 may be regarded as a measure of how much of the variation is heritable and ultimately how much variation exists in a population on which selection can act. It should be noted that estimates of h2 apply only to that population, in the environment and time in which h2 was actually measured. As a result, because investigations into the nature of h2 in individual populations are conducted in a glasshouse or common garden, the estimate of h2 provided reflects the genetic variability as expressed in the experimental environment. Nonetheless, narrow-sense heritability can determine how much of the variation that exists in a population has the potential to respond to selection. As a result, it is an ideal measure for conservation purposes. Estimates of heritability (h2) may suggest evolutionary potential of study population. Heritability is expressed within a range from zero to one, and can be thought of as value expressing the efficiency with which a population responds to natural selection. The ability to respond to novel selective challenges is proportional to the additive genetic variance for the selected trait. (Lynch 1996) Another estimate of quantitative variation is called broad-sense heritability (H2). H2 is defined as the portion of phenotypic variation that is genetic, or the degree of genetic determination, or more formally the ratio of genotypic variance to phenotypic variance in a population. Broad-sense heritability does not distinguish between additive genetic effects and dominance effects on the phenotype and cannot be used to estimate how much variation exists in a population that has the capacity to respond to selection.
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Broad-sense heritabilities often are straightforward to obtain, for instance through clonal repeatability experiments (Havens 1994). Because persistent environmental effects can contribute to the similarity of clonal individuals (Schwaegerle et al. 2000), H2 should be considered the upper limit for the heritability value. Characters with the lowest heritabilities, that is, traits that are conservative and have only a small amount of measurable fluctuation, are those most closely connected with reproductive fitness. In contrast, traits with the highest heritabilities are less likely to be important determinants of fitness (Falconer and Mackay 1996). This is because heritability is a measure of phenotypic variation, on which selection acts if there are differences in fitness associated with phenotypic variation for any given trait. Therefore, a trait strongly associated with fitness is expected to have lower variability if there are concomitant differences in fitness related to that variability, relative to a trait that is only loosely associated with fitness. However, even traits closely related to fitness are subject to environmental influences and may prove to be worthy components of variation to study for conservation purposes. Schaal et al. (1991) point out the utility of quantitative traits in the conservation of rare species. In particular, they discuss the direct measurement of phenotypic variation that is often assumed to indicate underlying genotypic variation. Such morphological variation is easy and inexpensive to measure and probably reflects local differentiation or even ecotypic variation. They also advocate controlled quantitative genetic studies for their utility in providing an estimate of potential response to selection. They point out that understanding the potential of a population to respond to a changing environment might be particularly useful if the species is to be introduced to a new site or habitat.
Redefining Genetic Diversity and Conservation It is almost axiomatic among conservation professionals that knowledge of underlying patterns of genetic variation is important to preserving species. Equally important is the method by which genetic variation is measured. Molecular markers can be exceptionally useful in determining historical and phylogeographic patterns that define population genetic structure, perhaps even at a regional scale. For instance, molecular evidence has been used to determine that the maxipinon, Pinus maximartinezii, which is con-
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fined to a single population of approximately 2,000 to 2,500 mature trees, is derived from an extreme bottleneck four to five generations ago, which is less than 1,000 years in this species (Ledig et al. 1999). They can be useful in determining reproductive patterns within and between populations and can help us detect underlying causes of reproductive failure when selfincompatibility alleles fall below viable thresholds, as in the Lakeside daisy (Hymenoxys acaulis var. glabra) (DeMauro 1993). They can also be used to define taxonomic units of conservation concern. Cole et al. (2001) used both isozyme and RAPD data to corroborate the recent treatment of Aconitum noveboracense and A. columbianum as a single species. Although Steen et al. (2000) use molecular marker data to conclude that two endemic species in the Saxafragaceae, Saxifraga opdalensis and S. svalbardensis, probably were derived from the same hybrid combination, they are morphologically and genetically distinct and should be referred to separate species. However, several authors have recently challenged the use of molecular marker data to estimate genetic variation as a conservation tool (Hedrick 2001; Reed and Frankham 2001; Butlin and Tregenza 1998; Storfer 1996). Reed and Frankham (2001) review the underlying assumption that the various measures of genetic diversity are positively and strongly correlated. They question, in particular, that a linear relationship is expected between mean heterozygosity and the variance for a polygenic trait (provided that all gene action is additive). They also discuss in some detail several factors that will disrupt the expected relationship between mean heterozygosity and quantitative variation, including differential selection and nonadditive genetic variation. Because isozyme characters and other molecular markers are selectively neutral, these characters are subject to fixation through genetic drift. Under drift, neutral markers are expected to go to fixation at a rate relative to their original frequencies and the effective population size. Fixation occurs more rapidly in small populations than in large ones because of sampling effects. Thus, many patterns of genetic diversity visualized using molecular markers reflect only the random effects of drift. Therefore, it is not surprising that many studies of small populations have revealed little or no genetic variation in studies carried out with molecular markers alone. An increasing body of evidence is revealing that the forces operating on neutral markers are profoundly different from those operating on quantitative traits that are often under strong selection. Quantitative traits, even
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those that are affected by a few genes of large effect, are unlikely to be fixed because of drift in small populations. Additionally, because quantitative traits generally are acted on by multiple genes of small effect, even genes that may have become fixed are not likely to have profound effects on quantitative variation (Holsinger and Vitt 1997). For example, a comparison of quantitative variation and isozyme variation in Pinus contorta ssp. latifolia (Pinaceae) revealed that quantitative traits associated with fitness were differentiated between populations in a nonrandom fashion (Yang et al. 1996). The authors conclude that differences found between populations for specific gravity of the wood, stem height and diameter, and branch length probably were attributed to some form of selection. JaramilloCorrea et al. (2001) found similar results in Picea glauca (Pinaceae). Therefore, it should not be assumed that quantitative traits related to fitness are under similar pressures of genetic erosion as are neutral traits measured by molecular markers. Additional evidence is provided by studies comparing genetic variation in small and large populations. Although marker variation often is limited in small populations because of drift, several recent studies have suggested no difference in levels of quantitative trait variation between large and small populations (e.g., Primula scotica, Primulaceae [Ennos et al. 1997], and Clarkia dudleyana, Onagraceae [Podolsky 2001]). Widen and Andersson (1993) found that a small population of Senecio integrifolius (Asteraceae) displayed significant additive genetic variation for a greater number of characters than a large, continuous one. In addition, they found slightly higher heritabilities in the characters measured in the smaller population than in the larger one. No differences were found in heritability estimates between populations of Scabiosa canescens (Dipsacaceae), a locally rare species, and its widespread congener S. columbaria (Waldmann and Andersson 1998). Ouborg and Van Treuren (1995) found no significant differences in fitness-related characters correlated with population size in Salvia pratensis (Lamiaceae), a threatened perennial in the Netherlands. Although molecular techniques often have been used to discern patterns of genetic variation in rare species, Hamrick et al. (1991) pointed out that less than half of the observed variation at polymorphic loci could be correlated with life history traits. Other authors have also shown that correlations between molecular marker data and quantitative genetic diversity are highly variable. For example, Schemske et al. (1995) point out that
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there are no empirical data directly linking molecular genetic composition of plant populations with growth rate or survival. Although Hedrick and Savolainen (1996) showed that levels of molecular variation and quantitative trait variability were similar within populations of Scots pine (Pinus sylvestris, Pinaceae), differentiation between populations was greater for the quantitative traits. Reed and Frankham (2001), in particular, have put this assumption to the test. They performed a meta-analysis of 71 published data sets and found a weak mean correlation between molecular and quantitative measures of genetic variation. They also found no significant relationship between the heterozygosity measured by molecular markers and heritabilities of life history traits within natural populations. They conclude that molecular measures of genetic diversity have only a very limited ability to predict quantitative genetic variability (Reed and Frankham 2001). In short, then, measures of genetic diversity using molecular markers generally don’t answer the question, Does enough genetic diversity exist in natural populations of rare, endangered, or threatened species to respond to selection? Most practitioners of conservation biology have focused on neutral, allelic genetic polymorphism and, with a few exceptions, have largely ignored quantitative variation. It is as a result of this bias that proponents of ex situ conservation have received the strongest criticism. Hamilton (1994) argues that propagule sampling that encompasses only neutral allelic variation is short-sighted because quantitative traits are those under selection. He advocates more research on the effects of using different methods of genetic diversity to assess the success or failure of ex situ conservation methods. However, practitioners who engage in ex situ conservation of rare native taxa do so because they are severely imperiled, and preservation of genetic material is seen as a backup preservation method to in situ methods or as a supplemental method to reintroduce or reestablish extirpated populations. The revised Center for Plant Conservation (CPC) guidelines are sensitive to the balance between collecting a genetically representative sample that includes enough quantitative genetic variation to respond to even future selective pressures and the need to limit propagule collection to what the population may safely bear (Appendix 1, this volume). Indeed, a broad array of conservation strategies are becoming informed by the need to ensure that quantitative genetic variation is addressed in reintroduction and restoration attempts from the species level to the community level.
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Quantitative Genetics, Ecosystem Restoration, and Species Reintroduction Degraded and altered habitats have become a major portion of the mosaic of our landscapes. As we become more aware of the problems loss of native habitats and ecosystem services engender, restored natural areas have become a viable alternative to blighted and abandoned properties. In many areas, particularly the Midwest, California, Florida, and more recently along the eastern seaboard, restoration efforts are speeding up the recovery of native habitats. The science of restoration ecology is still in its infancy, and many question whether restored landscapes can function as well as, and provide similar ecosystem services to, unaltered natural areas. There are also questions about how restorations are conducted and, most importantly, whether seed sources are appropriate (for a review, see Knapp and Dyer 1997). When large-scale ecosystem restoration efforts are under way, often the biggest challenge is obtaining seeds. Restoration projects often use huge amounts of seed, of multiple species that are native to a particular area. In many cases, demand for native seed is met with seed from remnant populations, which are often small. The copious amount of seed needed is then produced under agricultural conditions. In the first steps, wild collected seed often is subjected to sorting and storage conditions that are far removed from the conditions faced in the wild. Seeds of some of the less common species may even be grown out under greenhouse or nursery conditions before being placed in commercial beds. Even seeds that are directly sown in agricultural fields are subjected to highly unnatural conditions as they are grown in large-scale monocultures where they are not exposed to selective regimes that mimic the conditions of native, or even restored, landscapes. What are some of the potential consequences of using questionable seed sources? First, use of nonlocal seed could result in plantings of poorly adapted individuals that ultimately fail. This is a waste of time and financial resources and may result in restorations that are problematic, without a clear reason why they have failed, because they are likely to fail over a fairly long period of time. Seedlings and juveniles may become established under such circumstances but never form a viable adult population from which further recruitment occurs. An example of this phenomenon is cited by Knapp and Dyer (1997), who describe the use of a single accession of
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Purshia tridentata promoted for use across its range, where it is easily established but declines before maturity, ultimately resulting in nonsustainable populations. Second, when nonlocal seed sources are used, existing native populations may also be negatively affected by large-scale introductions of maladapted genes. This can arise even when the plantings ultimately fail if flowering occurs and propagules are transferred from maladapted individuals into local populations. For a discussion and review of this phenomenon, see Knapp and Dyer (1997). Knapp and Rice (1996) found a strong relationship between quantitative traits and local climate and suggest that climatic data may be used to form rough guidelines for adaptive zones. They conclude that isozyme variation may not be the best method of describing the genetic architecture of a species when the spatial scales of adaptation are the primary concern, and for restorations they suggest requiring plant material that originates from a region with a similar climate to that of the site being planted. As conservation efforts evolve from preserving extant populations to restoring, supplementing, and establishing new populations, maintaining a quantitative perspective becomes imperative. Individual species reintroductions have many of the same pitfalls as multispecies restoration efforts; the difference is largely one of scale and the rarity of species concerned. In addition, establishing viable populations during reintroductions entails consideration of local adaptation and the ability of small founder populations to respond to local selection forces. This means that species reintroduction must be conducted in such a way that quantitative genetic variation is maximized. Reintroductions of rare taxa often involve the most critically endangered species, and it is unusual for the natural history or biology of the species to be well understood. In many cases, little is known about the genetic structure of extirpated or even extant populations, and even less is known about the quantitative variation in natural populations. Often, when species are the subjects of reintroduction, the sources of founding propagules are extremely limited, almost guaranteeing low levels of genetic variability, regardless of how diversity may be measured. Given such inherent constraints, species reintroductions often are carried out with an attempt to account for what is known about the biology of the species in question. Often this includes an unacknowledged attempt to account for quantitative variation and possible local adaptation. When sites are chosen, there
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is usually an attempt to match the new environment with that of a site known to be occupied, currently or in the recent past, in regard to soil type, slope, aspect, and community structure (Guerrant and Pavlik 1998). The underlying assumption and hope is that the propagules contain the genetic resources to respond to the chosen environment in an appropriate fashion.
Ex Situ Conservation Given that most ex situ conservation programs have the ultimate aim of reestablishing or reinvigorating natural populations, it is imperative that propagule collection protocols be designed to retain genetic diversity for quantitative traits that may be adaptive. Often, collectors may be unaware of traits that are adaptive in the current environment, and they may collect without regard to what may be adaptive in a novel environment to which such propagules will be reintroduced. Because traits that are neutral, or possibly even maladaptive, in one environment may prove to be adaptive in another, it is imperative to collect as broad a sample as possible. In some ways, this may go against the collectors’ natural inclinations. By and large, humans are apt to select the largest, most robust individuals, perhaps rightly assuming that such individuals have the highest fitness and therefore will produce the best offspring. However, it must be pointed out that this may be true only in the current environment. Given a different set of environmental characters, these same individuals may not fare as well and may even be eliminated from the gene pool. Every attempt must be made to include as broad a spectrum of individuals as possible, even though some of the smaller plants may yield only a small amount of fruit or seed. Such low production may reflect not necessarily an overall lack of fitness but rather a lack of fitness in the current environment.
CPC Guidelines for Seed Collection: Do They Capture Quantitative Variation? Although the original CPC guidelines do not explicitly address the question of propagule collection to optimize or maintain high levels of quantitative variation, they were designed with this in mind (Holsinger, pers. comm., 2002). At the larger end of the given ranges in the guidelines (i.e., 50 individuals per population and up to 20 propagules per individual sampled), a fair amount of the quantitative genetic variation within each pop-
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ulation will be sampled. It is especially noteworthy that the guidelines suggest that collections err on the side of being larger when there are observed microsite differences within the population, when the breeding neighborhood is small, and when the population itself is quite large. All of these factors increase potential quantitative variation between individuals within a collected sample. The previous CPC guidelines recommended collecting the number of propagules (seeds) from a given individual to ensure that just a single representative of the original genotype survived to reproductive maturity. Additionally, to maximize quantitative variation, the sample collected must be truly random, within a stratified random context. The technical protocols for Seeds of Success, a collaborative program between the Bureau of Land Management and the Royal Botanic Gardens, Kew, instruct seed collectors to “sample equally and randomly across the extent of the population, maintaining a record of the number of individuals sampled.” The protocols further elaborate, “where the population exhibits a pattern of local variation, use a stratified random sampling method to ensure sampling from each microsite” (Bureau of Land Management 2003). Although the concept of a stratified random sample is easy to comprehend, actually collecting a sample free from human bias may be more difficult. For example, even though you may endeavor to collect seeds from random individuals, based on the order in which you encounter them, sometimes even just finding them involves human bias because our attention is directed toward those that are larger and tend to have a greater display. And although it may seem that collecting from the largest individuals ensures a sample from the most robust and best fit, collectors need to recall that they are then sampling fitness only within a particular environmental context. Under different environmental circumstances, other individuals may well have higher fitness. It may also be difficult to determine how the distribution of a population should influence the sampling strategy. The “ideal” population may indeed be distributed uniformly across a homogeneous environment, but in reality few, if any, populations are distributed in such a manner. Many populations are distributed in clumped fashion, constant to a particular environmental gradient. Or there may be obvious local variation in the habitat and in the populations being sampled. In this case, the best scenario is to first walk the site to delimit the extent of the population and determine possible environmental heterogeneity. Random sampling within
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each apparent microsite is imperative, in theory. In reality, a truly random sample within each stratum may be replaced by a haphazard sample conducted in such a manner as to eliminate as much human bias from the sample as possible. A simple random determinative device such as a coin or gaming die may be used to assist in making a sample more random. In the likely event that microsite differences occur, collecting a proportion of the seed sample, relative to the total population size, from each microsite should suffice. Additionally, because environment plays such an important role in determining relative fitness, it is likely that individuals will perform differently in different years. A genotype that performed poorly in one year may perform exceptionally well in another. As a result, a simple means of potentially increasing quantitative genetic variance is to sample across multiple years. Such a sampling scheme has additional benefits. Sampling less intensively each year and increasing the number of years of collection decreases the demographic impact of propagule removal (Chapter 15, this volume). Sampling across multiple years, even while sampling seeds from the same maternal plants, provides an opportunity to increase genetic diversity among the offspring because of the potential for increasing the number of sires among the progeny. One method of ensuring the greatest number of sires involves withinpopulation cross-pollination to ensure the broadest genetic representation by removing the potential for inbred progeny and ensuring the identity of each of multiple fathers. Another tactic to ensure the broadest genotypic diversity in a sample is to ensure that there are equal numbers of propagules from each maternal individual. Although this may seem to contradict the previous advice to collect seed even from poorly producing individuals, the effects of each strategy are somewhat different but may act in concert by increasing the number of dams when individuals sampled have fewer seeds. Zhong and Qualset (1995) use the weedy species Dasypyrum villosum as a model to understand how propagule sampling might be used to collect the greatest amount of quantitative genetic variation. They found that they could expect to capture 95 percent of the quantitative variation in the species by collecting seed from as few as five populations, each population with five or more half-sib seeds taken from five plants, on a regional basis. Even though they estimate capturing a high amount of existing quantitative variation in a small sample, they conclude that to account for differential seed germination (including seed mortality in long-term storage) a
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much greater number of seeds, perhaps more than 1,000, would be needed. They also conclude that interregional sampling is necessary to ensure capturing the broadest array of adaptive variation. Some populations that are locally adapted, perhaps to unique ecological circumstances, might be singled out for more intensive sampling to ensure that unique alleles, and combinations of alleles, are captured.
Conclusions As recent work has shown, our understanding of the differences between quantitative genetic variation and molecular marker variation has improved dramatically. Indeed, we have seen through meta-analysis by Reed and Frankham (2001) that diversity measured in quantitative traits and diversity measured through isozymes or other molecular marker methods are decoupled, probably as a result of different selective forces acting on quantitative traits, and marker data are affected primarily through random drift. Many of the studies cited share the same good news for conservationists: the news about genetic diversity is not as dire as previously thought. Populations of most species studied, even small, marginal populations, maintain higher levels of quantitative variation than of other forms of genetic variation. And as Holsinger and Vitt (1997) point out, adaptation to changing environments is not likely to involve the alleles most likely to be lost through genetic drift. It is more likely that they retain enough variation to respond to changing environments, provided that the changes are gradual enough for selection to operate in an adaptive context. Conservationists are beginning to see that quantitative, and presumably adaptive, variation demands our attention when we design restoration and ex situ conservation strategies.
Acknowledgments The authors thank Kent Holsinger, Ed Guerrant, Mike Maunder, and Stuart Wagenius for their thoughtful reviews of earlier versions of the manuscript and Jen Taylor for her assistance preparing the table. References Andersson, S. 1991. Quantitative genetic variation in a population of Crepis tectorum subsp. pumila (Asteraceae). Biological Journal of the Linnean Society 44:381–393.
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Chapter 15
Effects of Seed Collection on the Extinction Risk of Perennial Plants Eric S. Menges, Edward O. Guerrant Jr., and Samara Hamzé
The ultimate purposes of ex situ plant conservation collections are to provide sufficient plant material to establish new populations, reintroduce populations, or augment existing populations. In removing material from natural populations, we also want to minimize their extinction risk. The existence of genetically representative samples stored off site both reduces the probability of catastrophic loss of the donor population’s genetic legacy and provides the genetic material necessary to reintroduce or augment the population if needed. Ex situ methods therefore are a means to an end: enhanced medium- to long-term survival of wild populations. However, the act of removing samples for off-site storage increases, however slightly, the short-term risk of extinction. We are left with a basic conflict, which we must resolve to the advantage of endangered species: how can we collect enough material to meet conservation goals without damaging populations in the process? The answer to this question depends on the kind of material collected. Seeds are widely considered to be the propagule of choice for ex situ conservation collections. Seeds usually can be collected in greater numbers than plant parts or whole plants. For plants with orthodox seeds (i.e., those that can survive frozen dry storage), it is easier and more economical to maintain a greater fraction of the genetic complement of a collection as seeds in a seed bank than it is to grow plants (Eberhart et al. 1991; Frankel et al. 1995, Chapter 18, this volume). Seeds that can be stored help conservationists avoid artificial selection for individuals that do well in a garden setting, possibly reducing adaptation to their native habitats (Frankel et al. 1995; Reinartz 1995; Appendix 3, this volume). In a garden setting, 305
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plants may acquire diseases or pests that could be transferred to wild populations after outplanting (Brown and Briggs 1991; Gordon 1994). Finally, and perhaps most important for the current discussion, removing seeds is thought to be far less damaging to the survival prospects of a sampled population than is removal of other plant parts or entire plants (Frankel et al. 1995; Menges 1998). But a less damaging method does not necessarily mean that there is no meaningful damage. Seed removal may still significantly reduce the population growth rate and increase probabilities of extinction. Our challenge is to balance the seed collection with harm to the population from which we are collecting. In this chapter we simulate various seed collection scenarios and the resulting population dynamics of various species, using empirically derived data. We address several questions. What is the relative vulnerability to seed harvest of perennial plants across a wide range of life histories? How do intensity and frequency of harvest interact to affect population persistence? How does population size affect a sampled population’s response to collection? Is it better to harvest infrequently but heavily or frequently but lightly?
Demographic Models To explore some of the potential impacts of seed collection on sampled populations, we used empirically derived stage-based transition matrices as a basis for stochastic modeling (Menges 1990, 1998; Guerrant 1999). These models summarize (as demographic parameters) data on the fate of individuals (survival and growth), their fecundity (seed production), and the fate of seeds (germination, entry into persistent seed bank, and survival). Individuals are generally divided into classes based on stage or size. Projection matrices, often specific for individual populations or years, can be used to estimate whether populations governed by these demographic parameters are likely to be increasing or decreasing at equilibrium (judged by the finite rate of increase, lambda). Because environments are variable, stochastic simulations using varying matrices or matrix elements can explore extinction risks (Menges 2000). Using stochastic simulations, we examined the impact of intensity and frequency of seed harvests and the effects of population size on extinction risk in plants. We started with published projection matrices for species with at least two projection matrices so stochastic approaches could be
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used. We then simulated various levels and patterns of seed harvest. We examined 22 taxa representing a wide range of perennial life histories (Table 15.1). From the literature, we chose species with published projection matrices representing perennial plants with a range of life histories. Almost no studies of annuals are summarized with projection matrices. These might be expected to be particularly sensitive to seed collection, especially if they lacked seed banks. Most of our included species are perennial herbs (n = 20 cases) that reproduce iteroparously, or repeatedly (17 species; Bierzychudek 1982, 1999; Fiedler 1987; Moloney 1988; Menges 1990; O’Connor 1993; Horvitz and Schemske 1995; Lesica 1995; Nantel et al. 1996; Byers and Meagher 1997; Menges and Dolan 1998; Valverde and Silvertown 1998; Guerrant, unpublished 2000; T. Kaye, unpublished 2000). One perennial herb reproduces semelparously, that is, it produces seeds once, then dies (1 species; Werner and Caswell 1977). Two species may do either (Lesica and Shelly 1995). Six of the herb species also spread clonally. We also considered five species of nonclonal woody plants: three tree species and two shrubs. In total, 22 species (25 cases) were examined. We included only species represented by two or more separate projection matrices, preferring studies with three matrices or more. The different matrices represented two or more pairs of years for one population (preferred) or two or more populations of the species. Only studies with explicit matrices were included to avoid problems with our interpretations of data in other formats. Matrices had to include seeds or seedlings produced as fecundity terms in order to perform the simulations of seed harvests. For species where a seed stage was incorporated into the matrix, the authors document seed dormancy. Thus, we are confident that the pitfall of artificial dormancy pointed out by Caswell (1989) was avoided. In only one case (Bierzychudek 1982) the matrix included a seed stage in which there was no seed dormancy, but correcting this error made little difference in the values of lambda (Bierzychudek 1999) and no difference in extinction probability (EP; this chapter). In choosing from among multiple matrices for a given species, we avoided those producing very high or very low finite rates of increase (lambda) because seed harvesting effects were overwhelmed by these extreme rates of increase. Our approach was to apply stochastic matrix modeling using available demographic data summarized in projection matrices for populations of wild plants. Stochastic modeling was preferred over deterministic
table 15.1 Characteristics of 25 cases (22 perennial species, herbaceous perennials unless indicated as shrubs or herbs) for which seed harvesting effects were modeled. Number of Matrices
Matrices
Range of Lambda
Arabis fecunda (Vipond) A. fecunda (Charley’s)
3
Years
0.73–1.79
3
Years
0.80–1.15
Ardisia escallonioides
2
Sites
0.98–1.10
Arisaema triphyllum
2
Years
0.89–0.93a
Astragalus scaphoides Astrocaryum mexicanum Calathea ovandensis
4 2
Years Sites
0.83–1.31 0.99–1.02
4
Years
0.90–1.25
Calochortus obispoensis Danthonia sericea Dipsacus sylvestris
2
Years
0.96–1.03
4 4
Sites Sites
0.60–1.33 0.53–2.22
Species
Reproductive Mode
Response
Semelparous and iteroparous Semelparous and iteroparous Shrub
Seed
Sensitive II (low)
Seed
Extinction prone
Seed
Insensitive
Iteroparous
Seed/clonal
Extinction prone
Iteroparous Tree
Seed Seed
Sensitive II (low) Insensitive
Iteroparous
Seed
Sensitive II (low)
Iteroparous
Seed
Insensitive
Iteroparous Semelparous
Seed/clonal Seed
Sensitive I (high) Sensitive II (low)
Source
Growth Form
Lesica and Shelly (1995) Lesica and Shelly (1995) Pascarella and Horvitz (1998) Bierzychudek (1982, 1999) Lesica (1995) Pin˜ ero et al. (1984) Horvitz and Schemske (1995) Fiedler (1987) Moloney (1988) Werner and Caswell (1977)
Erythronium elegans
5
Years
1.00–1.05
Eupatorium perfoliatum Eupatorium resinosum
3
Sites
0.77–1.09
3
Sites
0.75–1.17
Fumana procumbens Heteropogon contortus Horkelia congesta Neodypsis decaryi
6 4 6 3
Years Years Years Sites
0.86–1.09 0.90–1.13 0.89–1.07 1.06–1.16
Panax quinquefolium Pedicularis furbishiae (all) P. furbishiae (Hamlin) P. furbishiae (St. Francis) Primula vulgaris
4 3
Sites Years
0.88–1.18 0.77–1.27
Guerrant (unpublished) Byers and Meagher (1997) Byers and Meagher (1997) Bengtsson (1993) O’Connor (1993) T. Kaye (pers. comm.) Ratsirarson et al. (1996) Nantel et al. (1996) Menges (1990)
Iteroparous
Seed/clonal
Insensitive
Iteroparous
Seed/clonal
Extinction prone
Iteroparous
Seed/clonal
Sensitive I (high)
Shrub Iteroparous Iteroparous Tree
Seed Seed Seed Seed
Sensitive II (low) Sensitive II (low) Sensitive II (low) Insensitive
Iteroparous Iteroparous
Seed Seed
Sensitive II (low) Sensitive II (low)
2 3
Years Years
1.05–1.12 0.58–0.98
Menges (1990) Menges (1990)
Iteroparous Iteroparous
Seed Seed
Insensitive Extinction prone
2
Years
0.95–1.04
Valverde and Silvertown (1998) Menges and Dolan (1998) O’Connor (1993) Olmsted and AlvarezBuyulla (1995)
Iteroparous
Seed/clonal
Insensitive
Silene regia
3
Years
1.32–1.42
Iteroparous
Seed
Sensitive II (low)
Themeda triandra Thrinax radiata
4 2
Years Sites
0.83–1.33 1.09–1.31
Iteroparous Tree
Seed Seed
Insensitive Insensitive
Corrected values, as used in Bierzychudek (1999), for Brooktondale population.
a
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ecological/evolutionary context of ex situ conservation
approaches as a more realistic depiction of both the demography of wild populations and the process of seed harvesting. It also allowed us to estimate the probability of extinction under various seed harvest scenarios, arguably the most fundamental worry of conservationists. We introduced stochasticity by randomly alternating matrices representing different years, and harvest or no-harvest years. If there were no published matrices representing different years for a species, we alternated published matrices from different populations, assuming that this variation was comparable to temporal variation. Alternating matrices rather than varying individual matrix elements preserves correlations between different parts of life histories, such as the tendency for positive correlations between life history parameters (Horvitz and Schemske 1995; Fieberg and Ellner 2001). Because positive correlations project more variable population trajectories, matrix alternation produces a more conservative risk assessment (Greenlee and Kaye 1997; Menges 2000) while having no data requirements relating to the variance structure of life histories. Complete specification of correlation structure does alter estimates of stochastic population growth (which correlates with extinction risk; Fieberg and Ellner 2001). Seed harvest scenarios varied the level of harvest in each year harvested, and the probability of harvest in each year, in various combinations. Individual year harvest levels were 10, 50, or 100 percent of fecundity. Probabilities of harvest were 10, 50, or 90 percent. We evaluated harvest scenarios with the initial population size of 10, 50, 100, or 500. The product of three harvest levels, three harvest probabilities, and four initial population sizes yielded 36 harvest scenarios. We also examined four no-harvest scenarios, one for each initial population size. In total, there were 40 scenarios for each case. No-harvest matrices were simply the published matrices. Initial simulations suggested that the most informative cases were matrices with finite rates of increase (lambda) not far from 1. EPs for cases with high or low values for lambda were not affected by seed harvest. Therefore, when there were multiple matrices for a species, we selected those with 0.85 < < 1.2. Differential harvest matrices were produced by reducing each fecundity term (seed production or seedling production) by 10, 50, or 100 percent (see discussion of assumptions later in this chapter). For a species with three published matrices, we produced 12 matrices representing three harvest matrices and one no-harvest matrix for each of the original matrices. We generally based our simulations on a single set of matrices (usually a single population) for each species. However, for Pedicularis furbishiae we
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simulated seed harvesting on two individual populations as well (resulting in three cases for this species). Many life history studies and population viability analyses have used lambda (the finite rate of increase) to compare years, populations, or treatments (Menges 2000). However, stochastic simulations do not equilibrate to produce a meaningful estimate of lambda. In addition, lambda does not always adequately predict transient dynamics (Werner and Caswell 1977; Bierzychudek 1999), which may lead to extinctions, even in apparently growing populations. Furthermore, periodic harvesting will prevent an equilibrium from being reached. Transient dynamics, though realistic, are difficult to compare among species. For these reasons, we used EP as one measure of extinction risk. Extinction risk may be the major concern of conservationists and is appropriate in the case of periodic seed harvests, which are likely to be a stochastic process. Extinctions of local populations may cause loss of alleles and thus erosion of overall genetic variation. The use of stochastic simulations with multiple matrices allows the direct estimate of EP.
Modeling Protocol For each analysis, we conducted stochastic simulations using POPPROJ3 (modified from earlier POPPROJ programs; Menges 1998). This program allows the selection of matrices for each year of the simulation with specified probabilities of selection for each matrix. For example, if the probability of seed harvest was 10 percent in any year, there would be a 10 percent chance of choosing one of the matrices with reduced fertility, representing a seed harvest year. Because each year is independent, there can be runs of several years with harvests or no harvests. We performed separate analyses for each of the 40 harvest scenarios, for each of the 25 cases. The probabilities of any matrix being selected reflected the designated probability of seed harvest, but otherwise different years and populations were weighted evenly. We began each simulation with one of the specified total population sizes. Initial populations were at stable stage distribution for one of the original matrices (the matrix with the median finite rate of increase or else the finite rate of increase closest to 1 for cases of two original matrices). Simulations were run 1,000 times for 100 years each. The population was considered extinct when its size dropped below 1.
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Some Assumptions An implicit assumption of our approach is that seedling recruitment and seed production are correlated in species lacking seed dormancy. This occurs because the matrix elements representing fecundity are actually products of total seed production and the probability of seedling emergence and survival; they represent seedlings added to the population at the time of the annual census. Because seed harvesting may take place before that census, the proportion of seeds removed may not exactly translate to the proportional reduction of fertility as defined by seedlings. For plants with seed banks, fertility is a combination of seedlings produced by last year’s seeds and seedlings from dormant seeds, and this assumption is not necessary. In addition, we assume that the variety of matrices represents the variance structure of population dynamics through time (environmental stochasticity). This is likely to be only approximately true when only a few matrices (years) are used. When only a few years of data are used for population projections, some errors in projection can be expected (Fiedler et al. 1998; Bierzychudek 1999). The projections are also independent of density. If removal of seeds or declines in population size are compensated by (negative) density-dependent effects, then our projections will be too pessimistic. Finally, our seed harvest scenarios assume several things specifically about the seed harvests themselves. We assume that the seed harvests are scaled to the population’s annual seed production. When seed production is higher, this same percentage seed harvest is a greater absolute harvest. Comparing species using relative seed harvests is necessary given a wider range of fecundities among species. Second, seed harvests are stochastic in that in any given year there is a probability of seed harvest. This means that there can be unusual runs of years with frequent or infrequent seed harvests compared with the mean seed harvest probability. We believe that this is a more realistic scenario than regularly spaced seed harvests. Third, seed harvests are assumed to vary randomly, whereas in many situations there may be particular patterns of seed harvests over time.
Response to Harvest: An Example For many species, increased frequencies and intensities of seed harvests led to increased extinction risks, and smaller populations were most vul-
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Figure 15.1 Extinction percentages for four initial population sizes (10, 50,
100, 500 in separate panels) and for four levels of percent of years harvested (0, 10, 50, 90 in different bar styles) as a function of percentage harvested each year, for Pedicularis furbishiae (from three mean matrices): a “sensitive type II” case.
nerable. We illustrate the full range of 40 scenarios for the mean projection matrix for Pedicularis furbishiae, a nonclonal perennial herb that is narrowly endemic to the banks of the St. John River in northern Maine and adjacent New Brunswick (Menges 1990). Extinction risks are low (<25 percent) for infrequent (10 percent of years), small (10 percent of seeds) harvests for all population sizes of 50 or more (Figure 15.1). Frequent (50 percent or higher), intense (50 percent or more seeds) harvests threaten populations smaller than 500. Complete, intense harvests threaten even these large populations. Various intermediate combinations of population size, harvest frequency, and harvest intensity produce intermediate extinction risks (Figure 15.1).
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Figure 15.2 Extinction percentages, by sensitivity class, all species combined,
for initial population size of 50.
Our modeling scenarios produced 40 scenarios for 25 cases. To organize these results, we first classify species into broad categories of response to harvesting.
Classifications of Population Responses Effects of seed harvests varied widely across species. We distinguish four distinct categories of response (Table 15.1, Figures 15.2 and 15.3): insensitive, extinction-prone, sensitive type I (high initial extinction risk), and sensitive type II (low initial extinction risk). We assumed that a safe harvest caused no more than a 5 percent increase in EP over baseline EP (no harvest scenario).
Insensitive Cases For nine cases, called insensitive, seed harvest regardless of intensity level or frequency caused no extinction risk over 100 years (EP = 0) for populations of 50 or more individuals (Table 15.1). For five of these cases, no har-
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Figure 15.3 Extinction risks for four species with initial population size of 10 for four levels of percent of years harvested (0, 10, 50, 90 in different bar styles) as a function of percentage harvested each year. (A) An extinction-prone species (Arisaema triphyllum). (B) A species sensitive to seed harvest with low initial extinction risk, sensitive type II (Fumana procumbens). (C) A species sensitive to seed harvest with high initial extinction risk, sensitive type I (Eupatorium resinosum). (D) An insensitive species (Primula vulgaris).
vest effects were evident at any population size. For three cases, harvest effects were only seen for very small (10) populations. In one case, harvest effects were evident only at 100 percent harvest for 90 percent of years but for all population sizes. Because these effects are seen only at extremely small population sizes or extremely high harvesting levels, we consider these species insensitive to seed harvests. All three tree species simulated were insensitive to seed harvests. Other insensitive species included shrubs and iteroparous herbaceous perennials (Table 15.1).
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Extinction-Prone Cases For four cases, called extinction-prone, EP was 100 percent even without seed harvests. These four cases were all herbs, including two herbs combining seed production and clonal growth. Populations are predicted as heading toward extinction even without seed harvests because of their demographic rates. In such cases, seed harvests for ex situ conservation could be justified as necessary in a desperate situation. The effects of seed harvests on extinction-prone species may be secondary to fundamental issues of population dynamics, which themselves may be tied to habitat quality.
Sensitive Cases The most interesting cases are the sensitive cases in which seed harvests increase extinction risk under certain circumstances. In general, for this group, increasing harvest intensities markedly increases extinction risk (Figure 15.2). Two types of sensitive species can be distinguished. For sensitive type I (high initial extinction risk), which occurred in two iteroparous herbs with clonal growth (Danthonia sericea and Eupatorium resinosum), initial extinction risk was fairly high (EP > 40 percent without seed harvest for initial populations of 500), and it increased with seed harvest (Figure 15.3). Sensitive type II cases have initially low extinction risk (EP = 0 percent without harvest for initial populations of 500) with increased extinction risk with seed harvest frequency and intensity (10 cases, e.g., Pedicularis furbishiae, Figure 15.1; Fumana procumbens, Figure 15.3). All these species are nonclonal, and all but one are herbaceous. There is little (<5 percent) increased extinction at 10 percent harvest, 10 percent of years, for any population size, but there is variable sensitivity to higher levels of harvests among species. Sensitive cases (extinction risks increased with seed harvests) had alternate matrices with > 1 and < 1.
Variation within a Species The population response of a species can vary from one population to another, depending on the demography of each population. For Pedicularis furbishiae, a growing population at Hamlin (1.05 < < 1.12) is insensitive to seed harvests, whereas a declining population at St. Francis (0.59
15. Effects of Seed Collection on the Extinction Risk of Perennial Plants
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Figure 15.4 Effect of initial population size (10, 50, 100, 500 in different bar
styles) for three species on extinction risk under the 50–50 harvesting scenario. as, Astragalus scaphoides; ca, Calathea ovandensis; hc, Horkelia congesta.
< < 1.08) is extinction-prone at any seed harvest level. However, the mean Pedicularis furbishiae “population” (0.77 < < 1.02) is sensitive to seed harvests (Figure 15.1).
Effects of Initial Population Size and Seed Harvesting Patterns Across all four response types, small populations (10) had the highest EPs. As expected, increasing initial population size decreased EP (Figure 15.4). When we examined the effects of initial population size within the most frequent response type, sensitive type II (low initial EP; n = 10 cases), we found that harvesting can occur only at the lowest intensity (10 percent) and with no more than 50 percent probability of harvest for initial populations of 10 (Table 15.2). However, as initial population size increases to 50, low-intensity, frequent harvests (i.e., 10–90 percent) can occur without increasing extinction risk. Once initial populations have at least 100 individuals, either 10 percent harvest at 90 percent frequency or 100 percent harvest at 10 percent frequency is safe. On average, even populations of 500 cannot sustain more than 50 percent harvest at 50 percent frequency or 100 percent harvest at 10 percent frequency. However, reasonable harvesting levels do not threaten healthy populations of 500 or more.
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table 15.2 Mean percentage extinction (EP) for sensitive type II species (with a low initial extinction probability). Shaded entries indicate 5 percent increase of EP over baseline EP. Frequency Initial Population
10
50
100
500
Intensity (% harvested)
% of Years Harvested
0 10 50 90 0 10 50 90 0 10 50 90 0 10 50 90
0
10
50
100
10.63 13.10 15.97
13.28 34.91 62.49
30.74 94.69 100.00
0.97 1.64 2.60
2.01 13.75 41.70
8.35 78.03 100.00
0.27 0.56 0.92
0.63 6.89 32.91
4.06 70.41 99.78
0.02 0.03 0.04
0.02 1.10 11.50
1.22 51.41 99.79
9.77
0.94
0.31
0.00
The nine insensitive species, regardless of population sizes, have no probability of extinction in the absence of harvesting. Populations of 10 individuals are only slightly affected by 50 percent harvest at 90 percent frequency or 100 percent harvest at 50 percent frequency. For populations of all sizes, the EP rises more than 5 percent above the baseline EP of zero only at the highest intensity and frequency of harvest (e.g., Primula vulgaris, Figure 15.3).
Effects of Frequent versus Intense Harvests For all species combined, frequent low-intensity harvests rather than infrequent but high-intensity harvests produced lower extinction risk (Figure 15.2). Similarly, we found that the 10 sensitive type II species (with low initial EP) were less affected by frequent harvest than intense harvest. For individual species, frequent harvests of lower intensity usually produced lower extinction risk (Figure 15.5). For example, a comparison of 100–10 and 10–90 harvest scenarios (each removing 9–10 percent of seeds in total) for individual species, for initial population size = 10, shows sig-
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Figure 15.5 Comparison of 100/10 and 10/90 harvest scenarios for 10 species, for initial population size = 10. Wilcoxon signed rank test, p = 0.009. af, Arabis fecunda; as, Astragalus scaphoides; ca, Calathea ovandensis; ds, Dipsacus sylvestris; fp, Fumana procumbens; hc, Horkelia congesta; hp, Heteropogon contortus; pf, Pedicularis furbishiae; pq, Panax quinquefolium; sr, Silene regia.
nificantly lower extinction risk for the frequent, low-intensity scenario (Wilcoxin signed rank test, p = 0.009). Other comparisons of intense and frequent harvests at initial population size = 50 sometimes show that frequent harvests produced lower extinction risks. For 100–50 and 50–90 (each removing 45–50 percent of seeds), the differences in extinction risk are significant (p = 0.005), but for 50–10 and 10–50 (each removing 5 percent of seeds), the differences are not significant (p = 0.47).
Conclusions For most woody species and rapidly expanding perennial herb populations, short-term seed harvests of any intensity and frequency are unlikely to cause short-term extinctions. Of course, this conclusion depends on the quality of data. Because demographic parameters vary widely in such herb populations, short-term studies may not capture the extent of demographic variation. Longer-term changes driven by succession or other environmental
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change may be important to such species but may not be anticipated by short-term demographic data (Bierzychudek 1999). In addition, sites supporting growing populations may be insensitive to seed harvest, whereas unfavorable sites may have vulnerable populations. To the extent that data are collected in only one type of site, extrapolation of results to other types of habitats may be misleading. We did not simulate seed harvest effects on annual plants (there are few studies with projection matrices for annuals). Annuals without persistent seed banks depend on each year’s seed production for recruitment and persistence and would be especially sensitive to seed harvests. Long-lived tree species may appear unaffected by seed harvest in our scenarios, which use 100-year simulations to compare species. However, this result probably is influenced by the demographic inertia created by the long lifespan of many trees. Tree populations without any seedling recruitment may nevertheless persist for many centuries (and so have little extinction risk apparent in our 100-year simulations). Removal of seeds by seed harvests could exacerbate long-term declines caused by other recruitment limitations. However, in many trees, fecundity can be huge, and recruitment limitations ultimately reflect competition or the disturbance regime (Platt et al. 1988; Alvarez-Buylla 1994). Even in species sensitive to some levels of seed harvests, there are generally safe seed harvest levels that emerge from modeling based on empirical data. Harvesting 10 percent of the seeds in 10 percent of the years typically does not increase extinction risks. We call this the 10/10 rule, which is a safe seed harvesting level for all the species we investigated. However, many species are sensitive to higher levels of harvests. For example, harvesting 50 percent of seeds in 50 percent of the years is generally an unsafe harvesting level (the 50/50 rule). Only populations larger than 500 can tolerate this level of harvesting without significant extinction risk over 100 years. Our third seed harvesting rule concerns the frequency and intensity of harvesting, which we call “slow but sure.” In this study, we contrast 9–10 percent overall harvest but with divergent harvesting schedules. Less intense harvests (e.g., 10 percent) that occur more frequently (e.g., 90 percent of years) produce lower extinction risks than heavier harvests (e.g., 100 percent) that occur infrequently (e.g., 10 percent). This occurs partly because some populations may depend on consistent fecundity to maintain population sizes, especially if they are short-lived and lack a seed bank.
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Intensive seed removal in one year may cause a population crash and subsequently produce higher extinction risk. In addition, intensive seed collection in consecutive years, possible in the simulations because of their stochastic nature, increased extinction risk in some species. To the extent that multiple seed collections are not coordinated among workers, this may be a realistic scenario. Smaller populations are most sensitive to concentrated harvests, and they may also experience greater collection pressure because of the lower overall abundance of seeds. Slow but sure collections, spread out over time, not only produce lower extinction risk but also may have other benefits that we have not modeled. If seedling recruitment conditions vary widely, having propagules available in favorable years lowers extinction risk. Also, spread-out collections can be integrated with monitoring of wild populations to help assess their viability and the potential effects of seed collection. Finally, seed collections made over a period of years may have advantages in sampling genetic variation that is sporadically available because of the emergence of new genotypes from the seed bank. Our simulations of collection patterns include random variation, but actual collections may be more systematic or concentrated in time. For example, seed collections of palms for forest garden plantings in Veracruz, Mexico, are heavy but concentrated for a few years and then subsequently lighter and more sporadic (T. Ticktin, pers. comm., 2000). Although our modeling suggests that concentrated harvests can threaten populations, such harvests for only a few years would be unlikely to have long-term effects on long-lived species. Unfortunately, data sets detailing actual harvesting patterns are uncommon (but see Nantel et al. 1996). Many other questions about harvesting seeds were not addressed in our study. Genetic considerations suggest sampling seeds equitably from many individuals and from a diversity of microhabitats, avoiding oversampling from the largest and most fecund plants (Falk and Holsinger 1991; Holsinger and Gottleib 1991; Frankel et al. 1995). Multiple collections through a fruiting season will sample phenological variation among and within plants, and perhaps different mixtures of selfed and outcrossed progeny. Spreading collections spatially and temporally is consistent with collecting smaller numbers of seeds more frequently to minimize demographic risks (Falk and Holsinger 1991). Isolated plants should not be the focus of collection efforts if these plants produce genetically inferior progeny because of inbreeding depression (Burrows 2000).
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This study did not consider the demographic effects of harvesting plants or plant parts (other than seeds) from wild populations. However, demographic considerations give us the basis to make some general comments. First, most plant species’ lambdas are more sensitive to variation in mortality and growth than to fecundity (Menges 1998; Silvertown et al. 1996). This sensitivity varies with the life history of the species and the context of the individual population, being highest in woody plants and herbs of shaded and late-successional habitats (Silvertown et al. 1996; Oostermeijer 2000). Any harvests of plant parts or whole plants ideally should be monitored to detect any increases in mortality. Increased mortality should be avoided in sensitive species or populations. In such situations, it may not be possible to harvest plants and still avoid elevated extinction risks (Nantel et al. 1996).
Acknowledgments We thank Paulette Bierzychudek, Tom Kaye, Pedro Quintana-Ascencio, and Tamara Ticktin for helpful reviews of earlier drafts of this manuscript. References Alvarez-Buylla, E. R. 1994. Density dependence and patch dynamics in tropical rain forests: matrix models and applications to a tree species. American Naturalist 143:155–191. Bengtsson, K. 1993. Fumana procumbens on Öland: population dynamics of a disjunct species at the northern limit of its range. Journal of Ecology 81:745–758. Bierzychudek, P. 1982. The demography of jack-in-the-pulpit, a forest perennial that changes sex. Ecological Monographs 52:335–351. Bierzychudek, P. 1999. Looking backwards: assessing the projections of a transition matrix model. Ecological Applications 9:1278–1287. Brown, A. H. D., and J. D. Briggs. 1991. Sampling strategies for genetic variation in ex situ collections of endangered plant species. Pages 97–119 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Burrows, G. E. 2000. Seed production in woodland and isolated trees of Eucalyptus melliodora (yellow box, Myrtaceae) in the south western slope of New South Wales. Australian Journal of Botany 48:681–685. Byers, D. L., and T. R. Meagher. 1997. A comparison of demographic characteristics in a rare and a common species of Eupatorium. Ecological Applications 7:519–530. Caswell, H. 1989. Matrix Population Models. Sunderland, MA: Sinauer. Eberhart, S. A., E. E. Roos, and L. E. Towill. 1991. Strategies for long-term management of germplasm collections. Pages 133–145 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press.
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Falk, D. A., and K. E. Holsinger (eds.). 1991. Genetics and Conservation of Rare Plants. New York: Oxford University Press. Fieberg, J., and S. P. Ellner. 2001. Stochastic matrix models for conservation and management: a comparative review of methods. Ecology Letters 4:244–266. Fiedler, P. L. 1987. Life history and population dynamics of rare and common mariposa lilies (Calochortus Pursh: Liliaceae). Journal of Ecology 75:977–995. Fiedler, P. L., B. E. Knapp, and N. Fredricks. 1998. Rare plant demography: lessons from the Mariposa lilies (Calochortus: Liliaceae). Pages 28–48 in P. L. Fiedler and P. M. Kareiva (eds.), Conservation Biology for the Coming Decade. Boston: Chapman & Hall. Frankel, O. H., A. H. D. Brown, and J. J. Burdon. 1995. The Conservation of Plant Biodiversity. New York: Cambridge University Press. Gordon, D. R. 1994. Translocation of species into conservation areas: a key for natural resource managers. Natural Areas Journal 14:31–37. Greenlee, J., and T. N. Kaye. 1997. Stochastic matrix projection: a comparison of the effect of element and matrix selection methods on quasi-extinction risk for Haplopappus radiatus (Asteraceae). Pages 66–71 in T. N. Kaye et al. (eds.), Conservation and Management of Native Plants and Fungi. Corvallis: Native Plant Society of Oregon. Guerrant, E. O. Jr. 1999. Comparative demography of Erythronium elegans in Two Populations: One Thought to Be in Decline (Lost Prairie), and One Presumably Healthy (Mt. Hebo). Final Report on Five Transitions, or Six Years of Data. Unpublished report prepared for the U.S. Department of the Interior Bureau of Land Management. Holsinger, K. E., and L. D. Gottleib. 1991. Conservation of rare and endangered plants: principles and prospects. Pages 195–208 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Horvitz, C. C., and D. W. Schemske. 1995. Spatiotemporal variation in demographic transitions of a tropical understory herb: projection matrix analysis. Ecological Monographs 65:155–192. Lesica, P. 1995. Demography of Astragalus scaphoides and effects of herbivory on population growth. Great Basin Naturalist 55:142–150. Lesica, P., and J. S. Shelly. 1995. Effects of reproductive mode on demography and life history of Arabis fecunda (Brassicaceae). American Journal of Botany 82:752–762. Menges, E. S. 1990. Population viability analysis for an endangered plant. Conservation Biology 4:52–60. Menges, E. S. 1998. Evaluating extinction risks in plants. Pages 49–65 in P. L. Fiedler and P. M. Kareiva (eds.), Conservation Biology for the Coming Decade. Boston: Chapman & Hall. Menges, E. S. 2000. Population viability analyses in plants: challenges and opportunities. Trends in Ecology and Evolution 15:51–56. Menges, E. S., and R. W. Dolan. 1998. Demographic viability of populations of Silene regian in midwestern prairies: relationships with fire management, genetic variation, geographic location, population size and isolation. Journal of Ecology 86:63–78. Moloney, K. A. 1988. Fine-scale spatial and temporal variation in the demography of a perennial bunchgrass. Ecology 69:1588–1598.
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Nantel, P., D. Gagnon, and A. Nault. 1996. Population viability analysis of American ginseng and wild leek harvested in stochastic environments. Conservation Biology 10:608–621. O’Connor, T. G. 1993. The influence of rainfall and grazing on the demography of some African savanna grasses: a matrix modelling approach. Journal of Applied Ecology 30:119–132. Olmsted, I., and E. R. Alvarez-Buyulla. 1995. Sustainable harvesting of tropical trees: demography and matrix models of two palm species in Mexico. Ecological Applications 5:484–500. Oostermeijer, J. G. B. 2000. Population viability analysis of the rare Gentiana pneumonanthe: importance of demography, genetics, and reproductive biology. Pages 313–334 in A. Young and G. Clarke (eds.), Genetics, Demography, and Viability of Fragmented Populations. London: Cambridge University Press. Pascarella, J. B., and C. C. Horvitz. 1998. Hurricane disturbance and the population dynamics of a tropical understory shrub: megamatrix elasticity analysis. Ecology 79:547–563. Piñero, D., M. Martinez-Ramos, and J. Sarukhán. 1984. A population model of Astrocaryum mexicanum and a sensitivity analysis of its finite rate of increase. Journal of Ecology 72:977–991. Platt, W. J., G. W. Evans, and S. L. Rathbun. 1988. The population dynamics of a long-lived conifer (Pinus palustris). American Naturalist 131:491–525. Ratsirarson, J. A., J. A. Silander Jr., and A. F. Richard. 1996. Conservation and management of a threatened Madagascar palm species, Neodypsis decaryi, Jumelle. Conservation Biology 10:40–52. Reinartz, J. A. 1995. Planting state-listed endangered and threatened plants. Conservation Biology 9:771–781. Silvertown, J., M. Franco, and E. Menges. 1996. Interpretation of elasticity matrices as an aid to the management of plant populations for conservation. Conservation Biology 10:591–597. Valverde, T., and J. Silvertown. 1998. Variation in the demography of a woodland understorey herb (Primula vulgaris) along the forest regeneration cycle: projection matrix analysis. Journal of Ecology 86:545–562. Werner, P. A., and H. Caswell. 1977. Population growth rates and age versus stagedistribution models for teasel (Dipsacus sylvestris Huds.). Ecology 58:1103–1111.
Chapter 16
Hybridization in Ex Situ Plant Collections: Conservation Concerns, Liabilities, and Opportunities Mike Maunder, Colin Hughes, Julie A. Hawkins, and Alastair Culham
Many plant evolutionary biologists, plant breeders, gardeners, and horticulturalists view hybridization as a constructive process resulting in interesting and potentially useful new diversity in the form of both artificial and spontaneous natural hybrids. However, the plant conservationist seeking to retain the “natural” genetic architecture of wild plant species and populations may view hybrids in a quite different light: as potential contaminants. There is a growing awareness of the impact of hybridization on the conservation of in situ populations and its role in plant extinction (Rieseberg 1991; Levin et al. 1996; Rhymer and Simberloff 1996; Carney et al. 2000; Wolf et al. 2001), but there has been little discussion of the implications of hybridization for ex situ conservation collections. Ex situ facilities serving plant conservation operate under the premise that plant material, cultivated or stored in that facility, can be used for species recovery, habitat restoration, or crop genetic development activities. However, poor management can mean that the genetic makeup, integrity, and value of ex situ material can be compromised by hybridization, thereby undermining these services. Managers of ex situ facilities and their clients, like their in situ conservation colleagues, need to be alert to the risks and opportunities posed by hybridization both within the bounds of their facilities and in surrounding areas. We consider ex situ collections in the broad sense to include not only those specifically managed for conservation purposes but also facilities where assemblages of congeners are cultivated in close proximity over significant periods of time. Furthermore, ex situ conservation collections sensu stricto are a recent invention. Accordingly, a broader perspective is needed to 325
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benefit from important historical insights into the likely outcomes of spontaneous hybridization, which often become apparent only after several decades or even centuries. However, having argued for a broad definition of what we mean by ex situ collections, it does appear that gardens have been, and continue to be, probably the most important source of spontaneous hybrids and invasive plants. Horticultural introductions have been the source of one-third to one-half of the world’s invasive plants (Bight 1998; Cronk and Fuller 1995; Low 1999), dominating the weed floras in the United Kingdom (Rob 1973; Nelson 1994), South Africa (Stirton 1978), Australia (Low 1999), and elsewhere. Much of our discussion therefore focuses on hybrids arising in or derived from horticultural or garden collections. Attention has been given to the influence of breeding systems, population structure, and levels of genetic diversity on threatened species management (Weller 1994). However, hybridization within or resulting indirectly from ex situ collections has received little attention. Ex situ collections, such as botanic gardens, house some of the most concentrated samples of plant diversity, species and varietal, in the world. Hybridization problems can arise through the use of collections managed for display or education as sources of material for conservation or breeding, creating conflicts between routine botanic garden management and ex situ conservation programs. For example, in reviewing the ex situ needs for threatened plants in botanic gardens, Poppendieck (1976) and Snogerup (1979) identified hybridization as a major problem. The traditional botanic garden collection or arboretum, where samples from taxa and populations that are geographically isolated in nature are brought together, can be viewed as an extreme form of artificial sympatry. We explore ex situ conservation in terms of managing collections to facilitate the safe or “pure” propagation of threatened taxa as part of threatened species recovery and preventing the generation and escape of novel and potentially invasive hybrids. Hybrids must be considered in relation to ex situ plant conservation for five distinct reasons. • Natural hybrids, an important component of natural diversity, may
themselves be the focus of conservation efforts. It has been suggested that plant hybrid zones support diverse and unique assemblages of biodiversity and act as evolutionary foci (Whitham et al. 1994; Martinsen and Whitham 1994) and therefore represent legitimate conservation targets. A clear understanding of the nature and
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table 16.1 Hybridization pathways involving ex situ collections. Issue
Examples
Hybridization in ex situ collections (both parents in cultivation)
Echium, Argyranthemum, Saintpaulia, x Rhododendron
Gene flow between collections and wild populations
Lantana, Echinacea, Schlumbergera
Novel hybrid taxa occurring outside ex situ collections after escape
Senecio cambrensis, Fallopia, Heracleum, Quercus
Possible Conservation Implications
Devaluation of cultivated stocks. Inappropriate material for repatriation and reintroduction. Introgression with wild threatened populations. Generation of novel invasive genotypes. Potential generation of novel hybrid invasives. New taxa of “conservation” interest.
dynamics of natural hybrid populations is needed for successful ex situ conservation of such material. • Hybridization has been shown to be an important stimulus to the evolution of invasiveness (Abbott 1992; Ellstrand and Schierenbeck 2000; Vilà et al. 2000), and hybrids arising within or from ex situ collections could contribute to it. Ex situ collections of various sorts have often been implicated as sources of invasive plants, but routes from such collections toward invasion that involve hybridization are rarely taken into account in ex situ conservation programs. Ex situ managers need to look beyond the bounds of their collections to consider possible invasive hybrids arising from all combinations of cultivated ex situ stock, naturalized escapes, and wild populations (Box 16.1; Table 16.1). • Artificial hybridization can also be used in ex situ conservation programs in some, albeit extreme, circumstances as a last resort to salvage some of the genetic information present in a species on the verge of extinction. However, hybridization can also influence the success of ex situ conservation in a number of detrimental ways. • Spontaneous hybridization in ex situ facilities can undermine the genetic integrity of ex situ collections and potentially contaminate open-pollinated seed or seedlings destined for reintroduction.
328
ecological/evolutionary context of ex situ conservation • Spontaneous hybridization can also occur between plants in ex situ
collections and adjacent wild populations, leading to their contamination and potentially contributing to genetic assimilation and extinction of natural populations of rare and threatened species. This type of spontaneous hybridization may occur directly between wild and cultivated material or indirectly via naturalized escapes, which can provide important stepping stones between ex situ collections and surrounding natural and seminatural habitats (Table 16.1).
box 16.1 Management Guidelines We propose a number of collection management guidelines designed to reduce spontaneous hybridization within and the release of potential invasive hybrid taxa from ex situ plant collections and suggestions to reduce hybridization in management of threatened taxa. Collection Management • Enact policies to minimize undesirable spontaneous hybridization in
•
• •
• •
terms of what species are grown and how collections are designed, laid out, and horticulturally managed. Plan collections to minimize hybridization by planting congenerics as far apart as possible (e.g., the Townsville Palmetum was specifically designed with this in mind, with congeneric palms separated within the collections). Manage collections to reduce the production or persistence of hybrids (e.g., by effective weed control and dead-heading of seeding plants). Where hybridization is anticipated or expected (see Table 16.3), bag inflorescences or remove them from congenerics around plants used for seed production, and produce seed using controlled pollination. Physically separate conservation and horticultural display facilities. Avoid the use or promotion of geographically or taxonomically themed collections as conservation resources, use them as educational displays only, and practice strict prevention of seed production and distribution.
• Avoid planting single-genus collections (e.g., endemic Hibiscus
species of the Indian Ocean Islands or Echinacea species of the American prairies) near important wild populations of congenerics. Particular concern should be given to the planting of hybrids and cultivars derived from indigenous and local wild species. • Identify high-risk plant groups in terms of hybridization and invasive tendencies and avoid cultivating species from this group. Material Exchange • Tailor acquisition, exchange, distribution, and release policies to min-
imize unnecessary escape or release of potentially invasive species, and warn recipients about risks associated with species introductions, including hybridization. Impact Assessment and Monitoring • Assess and monitor the likely and actual impacts of ex situ collections
on adjacent natural or seminatural habitats in collaboration with management authorities for nearby habitats and threatened species. Training • Make all horticultural and scientific staff associated with ex situ col-
lections aware of the risks and possible consequences of hybridization within or derived from ex situ collections. Education • Use living collections to demonstrate and interpret what hybrids are,
how pollen flows, and the ecological and economic dangers of invasive species. Species Management Propagation of threatened species must aim to minimize the negative impacts of spontaneous hybridization by • Establishing the taxonomic status and hybridity of the threatened
taxon (i.e., is the target taxon a natural hybrid or part of a hybrid swarm or complex of interfertile taxa?). • Establishing the susceptibility of the taxon to hybridization. Does the taxonomic, biosystematic, or horticultural literature indicate that
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•
• •
•
•
this taxon (or genus) is susceptible to hybridization in the wild and in cultivation? Clarifying the conservation objectives. Does the recovery plan specify the role of the ex situ facility in terms of necessary propagule types (seed or vegetative) and genetic status? Where possible, use vegetative propagation for high-risk taxa. Dedicating isolated propagation facilities for the plant. Produce seed under controlled and insect- and pollen-screened conditions. Removing or reducing any congenerics or sister genera in the immediate environs (wild and cultivated) that could hybridize with the target taxon. Assessing origins of all founder stocks, with particular attention to seed-derived founders from unmanaged ex situ collections where hybridization may have occurred. Genetic screening is recommended in this situation. Minimizing the number of reproductive generations and using seed storage to store propagules. Where appropriate, use clonal or vegetative duplication for propagation.
Research • Liaise with researchers to promote and stimulate research studies of
hybrid problems, invasive origins, and hazards and basic reproductive and population biology of threatened taxa.
This chapter discusses these different facets of hybridization and their implications for ex situ plant conservation practices. We attempt to look at the aspects of hybridization that are relevant to ex situ conservation, particularly to the management of ex situ collections. We briefly review hybridization and the importance and frequency of natural hybridization and discuss the ex situ conservation of natural hybrids. We explore crossability, crossing barriers, and sympatry in more detail as a basis for evaluating the importance of hybridization in ex situ collections and the possible conservation impacts of spontaneous hybridization involving ex situ collections (Table 16.1, 16.2). Finally, we attempt to draw some conclusions about hybrids in ex situ collections as the basis for new management guidelines (Box 16.1). Although artificial hybrids have been and continue to be important in horticulture, forestry, and agriculture and can have important conservation impacts (e.g., as invasives, or allowing the survival of some highly threatened lineages), their impacts on ex situ conservation
table 16.2 Estimated levels of natural and artificial sympatry for some genera with recorded hybrids. Genus
Number of Globally Threatened Taxa a
In Situ Sympatry in Natural Populations (species)
Artificial Sympatry in Ex Situ Collections
27
2–5
6 species and 75 cultivars/hybrids, NCCPG National Collection.
Cotoneaster
9
2–10
Echium section Simplicia
3
None; essentially allopatric, separated by geography and habitat 2–5
300 species and 25 cultivars, NCCPG National Collection. All three species (E. simplex, E. wildpretii, and E. pininana) are widely grown in botanic gardens where they hybridize. 113 species and 32 cultivars, NCCPG National Collection. 80 species and 170 cultivars/ hybrids, NCCPG National Collection.
Argyranthemum
Euphorbia Iris
302 53
2–5
Hybrid Issues
Hybrids recorded from disturbed habitats in the wild; extensive artificial hybridization for horticulture. Hybrids recorded in wild and in garden stocks. No wild hybridization recorded; extensive hybridization found in botanic garden populations. Spontaneous hybrids recorded from botanic gardens. Spontaneous hybrids recorded from botanic gardens.
table 16.2 (continued ) Estimated levels of natural and artificial sympatry for some genera with recorded hybrids. Genus
Number of Globally Threatened Taxa a
In Situ Sympatry in Natural Populations (species)
Lithops
27
2–6
Opuntia
71
2–8
Rhododendron
63
2–10
34 species and 255 “forms,” NCCPG National Collection. 150 species at Palermo Botanical Garden, Italy. 632 species at NCCPG National Collection.
Salix
36
Often 4–6 species
260 species and 75 cultivars, NCCPG National Collection.
Sarracenia
10
2–5 species
10 species and 25 hybrids, NCCPG National Collection.
Artificial Sympatry in Ex Situ Collections
Source: NCCPG, National Council for the Conservation of Plants and Gardens, UK. a Sensu International Union for the Conservation of Nature (Walters and Gillett 1997).
Hybrid Issues
Spontaneous hybrids recorded from botanic gardens. Spontaneous hybrids recorded from botanic gardens. Hybrids recorded in wild and in garden stocks. Hybrids common in both wild and gardens, sometimes involving many species. Hybrid swarms common in some species pairs. Deliberate hybridization resulting in the registration of many hybrids under cultivar names. Hybrids are fully fertile.
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appear to be much less important and more straightforward than those of spontaneous hybrids, and we do not discuss artificial hybrids in detail.
What Is a Hybrid? Geneticists and taxonomists use the term hybrid in different ways. Geneticists refer to the offspring of any genetically distinct line as a hybrid, whereas taxonomists define hybrids as the offspring of taxonomically distinct parents (Stace 1975, 1989; Harrison 1993). The taxonomist’s definition of hybridity depends on the definition of species, a term that is itself controversial (Harrison 1993). McDade (1990) overcame this limitation by defining hybrids as organisms or lineages of reticulate history at a level at which divergence is expected. Both the geneticists’ and taxonomists’ definitions of hybridization are important to conservationists. It is worth noting that it has been suggested that the term hybrid not be used in U.S. conservation legislation and the term inter-cross be adopted (Federal Register 1996). We are concerned here with hybridization between lineages that have diverged more widely, particularly with interspecific hybridization in seminatural and managed environments associated with or influenced by cultivated populations. However, conservationists also need to be aware of the implications of hybridization between genetically distinct populations. This is important for conserving infraspecific diversity through avoiding potential outbreeding depression, which may affect species viability (Keller et al. 2000), and, conversely, reinstating beneficial gene flow between isolated populations (Newman and Tallmon 2001). Furthermore, Ellstrand and Schierenbeck (2000) implicate hybridization between disparate populations, as well as between species, as a potential stimulus to the evolution of invasiveness.
Frequency and Importance of Hybridization Natural hybridization is a common event and has played an important role in plant evolution (Stebbins 1959; Arnold 1992, 1997; Arnold et al. 1999; Rieseberg 1995, 1997; Rieseberg and Noyes 1998; Jackson et al. 1999; Hardig et al. 2000). Hybridization, both spontaneous and artificial, has also played a fundamental role in crop evolution and domestication (Roberts
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1965; Small 1984; Ellstrand et al. 1999), with plant breeders often using wide interspecific crossing for crop, horticultural, and tree genetic improvement (Kalloo and Chowdhury 1992; Pickersgill 1993). A number of studies have assessed the frequency of interspecific hybridization in wild plant populations. For example, Knobloch (1971) identified 23,675 putative interspecific and intergeneric hybrids based on an extensive literature survey. In a study of five biosystematic floras, Ellstrand et al. (1996) showed the average frequency of hybridization among vascular plants to be about 11 percent. Extrapolating from the occurrence of interspecific hybrids in the well-studied British flora, Stace (1989) estimated approximately 78,000 naturally occurring interspecific hybrids across all angiosperms. Stace (1975) noted an overrepresentation of hybrids in Onagraceae, Orchidaceae, Pinaceae, Rosaceae, and Salicaceae, and Ellstrand et al. (1996) identified the Scrophulariaceae, Salicaceae, Rosaceae, Poaceae, Asteraceae, Cyperaceae, and herbaceous outcrossing perennials as groups with a preponderance of hybrids. However, these surveys are heavily biased toward well-studied temperate floras. The recent proliferation of molecular studies not only supports these broad estimates of the high frequency of natural hybridization but also suggests that many cases of hybridization may have previously gone undetected (Rieseberg and Brunsfeld 1991; Rieseberg and Ellstrand 1993; Wendel and Doyle 1998). Introgression or divergence after hybridization may obscure hybrid origins (McDade 1990; Wolfe and Elisens 1994). Cryptic or ancient hybrids may be identified only through incongruence between morphological, chloroplast, and nuclear gene phylogenies. For example, introgression, the transfer of genetic material between hybridizing taxa through repeated backcrossing of hybrids, is revealed in molecular studies by finding the chloroplast DNA haplotype of one species in individuals that appear to belong to another species in both morphology and nuclear markers (Rieseberg and Brunsfeld 1991; Wendel and Doyle 1998).
Management of Natural and Artificial Hybrids for Conservation The high frequency of natural hybrids indicates the importance of hybrids as a component of overall biodiversity worthy of conservation effort alongside nonhybrid taxa. However, the conservation of hybrids can be more difficult and complex than for nonhybrid taxa and can raise a number of
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specific dilemmas in terms of appropriate conservation action and introduce somewhat spurious distractions from more important conservation priorities. At least four different scenarios involve the conservation of hybrids and the management of hybrid material in ex situ collections and are discussed in more detail in this chapter: the conservation of naturally occurring hybrid populations or swarms as part of a habitat conservation or restoration program; the management of naturally occurring hybrids mixed within, and possibly contributing to the demise of, populations of a threatened species; the management of highly threatened taxa of hybrid origin; and the use of artificial hybridization to salvage genetic variation of threatened species on the verge of extinction.
Threatened Taxa of Hybrid Origin There are uncommon hybrids that merit conservation attention. For instance, Agave arizonica, endemic to Arizona, appears to be a natural hybrid and is threatened by overgrazing from cattle (De Lamater and Hodgson 1987). Molecular and morphological evidence suggests that the threatened tree, Eucalyptus graniticola, known from a single individual on a granite outcrop in Western Australia, is a hybrid. Accordingly, it is likely that the conservation of this rare hybrid will depend on ex situ propagation and storage of the single genotype rather than on in situ recovery of a population (Rossetto et al. 1997). In other cases hybrids seem to present more of a distraction from more serious conservation priorities than a legitimate conservation objective in their own right. For example, Senecio cambrensis is listed in the British Red Data Book for vascular plants (Wigginton 1998). S. cambrensis is the allopolyploid product of S. x baxteri, a hybrid that resulted from introduction of S. squalidus from Mount Etna (Italy) to the Oxford Botanic Garden and its subsequent escape along newly built railways to meet up with the native British ragwort S. vulgaris (Lowe and Abbott 1996). Isozyme and cpDNA data indicate independent origins of S. cambrensis in Scotland and Wales (Harris and Ingram 1991; Ashton and Abbott 1996). Although S. cambrensis is now one of the few vascular plant species endemic to the United Kingdom, its conservation seems to be spurious, driven by an interest to preserve the outcomes of an interesting example of human-made evolution in real time, especially given that these outcomes could be viewed in quite a different light, as contamination and disruption of a native species by an introduced invasive.
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Hybrid Swarms One obvious problem with conservation of hybrids is that they may be poorly circumscribed taxonomically and morphologically compared with nonhybrid taxa. Many hybrids are morphologically cryptic and revealed only by detailed investigation using molecular markers (e.g., Milne and Abbott 2000). Poor definition can be particularly problematic in the case of hybrid swarms in which the morphological distinctions that define the two parent species are blurred by a spectrum of morphological intermediacy. Hybrid swarms are common (Leebens-Mack and Milligan 1998; Wang and Cruzan 1998; Milne et al. 1999), particularly in disturbed habitats, and are likely to occur in disturbed habitats that become the focus of habitat restoration efforts (Lesica and Allendorf 1999; Neuffer et al. 1999). In hybrid situations, traditional morphologically defined conservation units do not suffice, and conservation should ideally be guided by a clear understanding of the evolutionary and ecological dynamics of the hybrid zone (Whitham et al. 1999; Carney et al. 2000), but this often necessitates detailed data that are not always available. The conservation of economically important domesticated and wild complexes of crop relatives (Villani et al. 1999), which often involve hybridization, presents similar challenges (Small 1984; Ellstrand et al. 1999).
The Management of Hybrids within Populations of a Threatened Species The management of hybrids within populations of threatened species can pose other dilemmas for conservationists, not least because the hybrids may make up a significant fraction of the extant genetic diversity and may be part of the threat themselves. Wild populations of the rare Cercocarpus traskiae, endemic to Santa Catalina Island, California, which contain hybrids between C. traskiae and the nonthreatened C. betuloides var. blanchae, provide a good example of these sorts of dilemmas. Detailed study of the 11 known adult C. traskiae trees using random amplified polymorphic DNA indicated that six appear to be pure C. traskiae, the other five are hybrids, and half of the genetic diversity of the species would be lost if the sympatric hybrids were removed from the wild populations (Rieseberg and Gerber 1995). Two options were discussed: removal of all hybrid plants and the associated loss of genetic diversity or propagation of nonhybrid individuals for translocation, with the hybrid population maintained.
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Hybridization as a Conservation Tool In extreme circumstances hybridization can be used as a conservation tool to restore fertility to a highly threatened species that is no longer producing seed, to confer resistance to a species threatened by an introduced pest or disease, or simply to salvage some of the genetic variation in species on the verge of extinction that are no longer amenable to other conservation measures (Carney et al. 2000). The responsible use of hybridization, or the propagation of a natural hybrid, may allow the survival of genes from highly threatened lineages. However, such experimental hybridizations could be undermined by outbreeding depression (Waser and Price 1994). Perhaps the best-known example in which hybridization has been used in this way is the St. Helena redwood, Trochetiopsis erythroxylon, which became extinct in the wild in the 1950s. Although seed was collected before extinction, cultivated material reliant on self-pollination showed signs of catastrophic inbreeding depression, prompting artificial hybridization with the closely related St. Helena ebony, T. ebenus (S. Goodenough, pers. comm., 1998). The hybrid (T. x benjamini) shows extreme heterosis, being more vigorous than either parent, and has been successfully planted on the island, thereby preserving genes of both species (Cronk 1995); however, this hybrid is reported to be backcrossing with the ebony (Maunder 1995). Spontaneous and fertile hybrids within the highly threatened Hawaiian endemic genus Hibiscadelphus (Malvaceae) have promoted discussion on the use of deliberate hybrids as a conservation tool (Baker and Allen 1977). A natural hybrid of Argyroxiphium (Asteraceae) found in Hawaii appears to represent a cross between A. virescens (extinct) and A. sandwicense and as such is an opportunity to maintain genetic representation of the extinct A. virescens (Carr and Medeiros 1998). Examples of artificial hybrids raised from threatened species include hybrids of Pennantia baylissiana (Icacinaceae, New Zealand) with one surviving wild tree and P. corymbosa (Anonymous 1997); the Franklin tree, Franklinia alatamaha (Theaceae), extinct in the wild in the United States and crossed with Gordonia lasianthus (Theaceae; Orton 1977); and a number of threatened Macaronesian taxa (e.g., Lotus and Argyranthemum) hybridized to create new garden cultivars (Cunneen 1995). These examples suggest that this approach could be used more widely as a last resort when all other options have failed. The cycad Encephalartos woodii (Zamiaceae) survives only as a male plant; crossing with female congenerics has been recommended as a management option (C. Dalzell, pers. comm., 2002).
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Crossability and Sympatry The extent and outcomes of hybridization are influenced by a wide range of internal and external factors. These include the occurrence of species in close proximity with either overlapping (sympatric) or adjoining (parapatric) distributions; at least partially overlapping flowering times and common pollination vectors; successful transfer and germination of pollen on the foreign stigma, pollen tube growth, fertilization, and development of the hybrid embryos into mature seeds; successful germination of hybrid seeds and establishment of hybrid seedlings that are able to compete with seedlings of one or both parent species; and formation of adult hybrid plants that are in turn able to reproduce. Barriers to hybridization can occur at any stage, and often more than one barrier prevents hybridization between particular species pairs. Potentially interfertile species often are isolated by distance, phenology, environment, or ecological niche (Grant 1949). The extent and importance of spatial isolation as a barrier to crossing are indicated by high artificial crossability between many groups of allopatric species. For example, detailed artificial hybridization experiments for several woody genera including Erythrina (Neill 1988), Salix (Mosseler 1990), Eucalyptus (Griffin et al. 1988), Labordia (Motley and Carr 1998) and Leucaena (Sorensson and Brewbaker 1994) show high crossability between species with essentially allopatric distributions. Studies of reproductive barriers on oceanic islands have reached similar conclusions about the importance and frequency of spatial (between islands) and environmental (habitat) isolating mechanisms (Carr and Kyhos 1986; Marrero-Rodríguez 1992; Carr 1995; Smith et al. 1996; Stuessy et al. 1998). Indeed, within the Hawaiian flora all large genera contain species that readily hybridize when sympatric (C. Morden, pers. comm., 2002). In a detailed study of hybridization in the Canary Islands, Marrero-Rodríguez (1992) found that although many spatially isolated lineages are interfertile, three main patterns of hybridization occur: species that are rarely sympatric but form hybrids when brought together in cultivation, such as Limonium (Plumbaginaceae) and Bencomia (Rosaceae); sympatric taxa that form natural hybrids in the wild, such as Argyranthemum (Asteraceae), Echium (Boraginaceae), and Micromeria (Lamiaceae); and sympatric species that rarely form hybrids, such as Aeonium (Crassulaceae), Sonchus (Asteraceae), and Euphorbia (Euphorbiaceae). Given the importance of spatial isolation for many cross-compatible species, it is not surprising that juxtaposition of pre-
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viously isolated species in cultivation—that is, in artificial sympatry, the hallmark of ex situ collections—opens up many opportunities for hybridization to occur. Other barriers to interspecific crossing include the pollen-pistil interaction, which often prevents pollen germination (Rieseberg et al. 1995; Klips 1999), and genetic barriers. Finally, there can also be significant intraspecific variation in interspecific crossability. For example, wild accessions of Phaseolus vulgaris and P. coccineus (Fabaceae) cross spontaneously, whereas embryo rescue is needed to obtain hybrids from crosses between domesticated accessions (Pickersgill 1993). Even if the resulting hybrid seed is viable, the F1 progeny may be wholly or partially sterile or subject to hybrid breakdown (Stace 1975), and even within genera with high interspecific crossability, there is often a wide spectrum of hybrid fertility and fitness. For example, in the genus Leucaena, with 70 percent crossability between species, hybrids can be highly sterile, partially or fully fertile, or either self-incompatible or self-fertile (Sorensson and Brewbaker 1994). Similar results were found in crossing experiments on Erythrina (Neill 1988). This variation means that many F1 hybrids do not reproduce and cannot persist, and numerous such ephemeral hybrids have been described especially among annuals (Stace 1975). However, even in the face of low fertility or near sterility, because they persist for many years, F1 hybrids of perennial species may in themselves make significant ecological impacts, especially if able to propagate vegetatively, and may produce occasional fertile pollen, leading to backcrossing. Four evolutionary outcomes of hybridization are possible (Abbott and Milne 1995; Rieseberg 1997): introgressive origin of new intraspecific taxa, origin of new fertile homoploid hybrid derivative species, the extinction of one or both taxa through assimilation of one taxa into the other, and origin of new allopolyploid species. A growing number of studies have documented these different outcomes in plants across a range of genera (e.g., Helianthus [Rieseberg 1997], Senecio [Abbott et al. 1995], and Cardamine [Urbanska and Landolt 1998]).
Artificial Sympatry When plants are brought together into artificial sympatry, as in ex situ collections, or when one species increases its range after introduction, one of the most important and common barriers to crossing, that of geographic
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isolation, is removed, thus establishing new opportunities for hybridization between the previously isolated taxa (Table 16.2). Ecological or temporal (phenological) barriers to hybridization can also be altered and eroded in cultivation. For example, flowering and fruiting behavior of cultivated stocks may be distorted by altered growing conditions, weather, and day length. In addition, exotic species in cultivation may be exposed to novel pollinators, again opening up new opportunities for hybridization. For example, flowering specimens of Sophora toromiro (Fabaceae), a tree extinct in the wild from Rapa Nui (Maunder et al. 2000), in cultivation in the Melbourne Botanic Gardens, Australia, are regularly visited by honeyeaters, a group of birds not known from Rapa Nui (S. Glissman-Gough, pers. comm., 2001). Conversely, species with highly specific pollination syndromes may be deprived of their pollinators in ex situ environments. Specialized pollination syndromes are rarer in temperate floras, which depend largely on opportunistic social bees (Waser et al. 1996) or wind, than in tropical and subtropical regions (Johnson 1996), and there is great variation between and within plant families in their degree of pollinator specialism. For example, Asteraceae and Ranunculaceae are generalists, whereas the Asclepiadaceae and Orchidaceae are largely specialists (Johnson and Steiner 2000). In Table 16.2 we attempt to quantify, albeit in a preliminary way by means of selected examples, levels of artificial sympatry in ex situ collections and compare them with levels of natural in situ sympatry as well as other forms of circa situm artificial sympatry, such as may be encountered as a result of indigenous domestication. In Table 16.3 we indicate high-risk genera, containing threatened species and vulnerable to hybridization. Artificial sympatry is at its extreme in collections of horticultural taxa and arboreta. For example, the National Plant Collections Directory of the National Council for the Conservation of Plants and Gardens (NCCPG 1999) in the United Kingdom lists more than 600 collections growing over 50,000 plant species or cultivars. We suggest that artificial sympatry in ex situ collections is typically at least 10–30 times greater than levels of natural in situ sympatry, and for certain genera—where collections are very large indeed (e.g., Rhododendron) or where natural sympatry is very low or absent (e.g., Leucaena)—may be much higher than that. Although artificial sympatry within ex situ collections is inevitably high, sympatry between such collections and congeners in surrounding natural and seminatural habitats can be extremely limited in the immediate vicinity of the collection.
table 16.3 Examples of high-risk genera containing threatened taxa with recorded introgression (derived from Rieseberg and Wendel 1993), indicating genera with a high risk of hybridization in extensive horticultural collections (National Council for the Conservation of Plants and Gardens [NCCPG], United Kingdom) and conservation collections (Center for Plant Conservation [CPC], United States). Threatened taxa sensu IUCN (Walters and Gillett 1998). Genus
Family
Abies Acer Achillea Aquilegia Arctostaphylos Argyranthemum Asclepias Aster Ceanothus Cercocarpus Cistus Clarkia Cucurbita Delphinium Eucalyptus Fuchsia Geum Helianthus Heuchera Ipomopsis Iris Juniperus Penstemon Phlox Pinus Populus Potentilla Primula Quercus Ranunculus Salix Salvia Solanum Tradescantia Vaccinium Viola
Pinaceae Aceraceae Asteraceae Ranunculaceae Ericaceae Asteraceae Asclepiadaceae Asteraceae Rhamnaceae Rosaceae Cistaceae Onagraceae Cucurbitaceae Ranunculaceae Myrtaceae Onagraceae Rosaceae Asteraceae Saxifragaceae Polemoniaceae Iridaceae Cupressaceae Scrophulariaceae Polemoniaceae Pinaceae Salicaceae Rosaceae Primulaceae Fagaceae Ranunculaceae Salicaceae Lamiaceae Solanaceae Commelinaceae Ericaceae Violaceae
CPC Collections
NCCPG Collections
Globally Threatened Taxa
31 13 26 30 58 32 35 46 38 4 4 39 2 65 174 6 6 18 19 11 53 27 124 21 81 5 58 48 70 74 37 76 128 2 21 71
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However, this can increase when cultivated stocks escape and naturalize outside collection facilities. The assessment of sympatry is not always straightforward because it varies for different plant groups, being dependent on pollination mechanisms and hence distances over which effective pollen flow is likely. Two species pollinated by large long-distance Xylocopa bees may be sympatric in terms of potential for pollen flow even when they are separated by several kilometers, whereas two species pollinated by small Centris bees would be judged to be sympatric only if they occur within a few 100 meters of each other. Detailed assessments of sympatry of this type can be important for planning the layout and management of ex situ collections to minimize hybridization and for assessing risks of hybridization between plants in ex situ collections and nearby populations of rare and threatened congeners in natural and seminatural habitats surrounding collections.
Hybridization Pathways and Conservation The negative impacts of hybridization for conservation are documented for both animals (Green and Rothstein 1998; Allendorf et al. 2001) and plants (Rieseberg 1991; Levin et al. 1996; Rhymer and Simberloff 1996; Carney et al. 2000). The extent and significance of hybridization involving rare and threatened species suggest that this is not an uncommon or isolated threat (Stace 1975). Carney et al. (2000) suggest that a total of 130 rare plant taxa in the United Kingdom, 39 in California, and 38 in Hawaii are involved in or result from hybridization. In California, 90 percent of all threatened and endangered plants occur sympatrically or parapatrically with at least one congener (Anonymous 1989; Ellstrand 1992). Similarly, in the United Kingdom, about 10 percent of the listed protected plant species naturally hybridize with related nonthreatened species (Stace 1975). An assessment of botanic garden collections of threatened European species suggests that ex situ collections may be equally or even more vulnerable to hybridization than in situ populations of such taxa. In a review of threatened plants (as defined by the Berne Convention) cultivated in European botanic garden collections, there was a strong correlation between representation in botanic gardens and availability from commercial horticultural sources (Maunder et al. 2000), suggesting that horticulturally robust taxa are favored in botanic gardens. The most abundant 74
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species (in 25 genera) were checked for the occurrence of horticultural hybrids (Lord 1999), revealing 12 of the 25 genera recorded as having at least one hybrid available from commercial horticulture. Although this does not distinguish between artificial and spontaneous hybrids, it does provide a rough indication of potential and actual crossability within genera of horticultural value. Natural or artificial hybrids have been recorded from 12 of the 25 genera: Aster (Allen and Eccleston 1998), Cypripedium (Klier et al. 1991), Echium (Bramwell 1972), Euphorbia (Turner 1998), Fritillaria (Jefferson-Brown and Pratt 1997), Geranium (Yeo 1985), Limonium (Morgan et al. 1999), Primula (Richards 1993), Pulsatilla (Lindell 1998), Saxifraga (Webb and Gornall 1989; McGregor 1995; Holderegger 1998; Steen et al. 2000), Tulipa (Van Raamsdonk et al. 1995), and Typha (Kuehn et al. 1999). The outcomes of hybridization can have important implications for conservation that may be divided into three broad categories: loss of genetic integrity or genetic contamination of ex situ collections; gene flow from ex situ collections into wild populations, leading to contamination and potentially contributing to genetic assimilation and extinction of rare and threatened species; and the generation of novel invasive weeds. Spontaneous hybridization may involve cultivated stocks only, interaction between wild and cultivated stocks, or permutations of interactions between escaped naturalized, cultivated stocks and wild populations, involving flow of pollen both from and into ex situ collections (Table 16.1). These complex pathways are analogous to Small’s (1984) matrix of domesticated and wild plant hybridization events: wild x wild, domesticate x domesticate, wild x domesticate, and domesticate x weed. We provide examples of most of these pathways, and all must be considered as possibilities. A clear understanding of the importance of different pathways is clearly needed to underpin guidelines (Box 16.1) for management of ex situ collections that aim to minimize risks posed by hybridization. However, there are few data with which to assess the relative importance of different pathways in terms of their respective conservation impacts. Indeed, unraveling the exact pathways leading to hybridization can be extremely difficult (see Hollingsworth et al. 1998; Milne and Abbott 2000). However, given the limited direct contact or sympatry between ex situ collections and wild populations and very small size of populations in ex situ collections, it is clear that indirect contact involving escapes can act as stepping stones for gene flow and for the generation of novel invasives (Table 16.1). Escapes of all sorts must be considered,
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including incipient local and more widely naturalized taxa as well as taxa that escape through intentional or unintentional release into wider horticultural, agricultural, or forestry use and cultivation. Studies attempting to predict the impact of hybridization on the likelihood and rate of extinction (Huxel 1999; Wolf et al. 2001) indicate that extinction of a population or species is more likely if the native taxon lacks competitive advantage, is uncommon, and has weak reproductive barriers. The extinction process can be very rapid; in some cases extinction may occur in as few as five generations (Wolf et al. 2001).
Loss of Genetic Integrity of Ex Situ Collections Since the eighteenth century, botanic gardens and commercial nurseries have regularly produced new hybrids for both ornamental and commercial purposes (Wilks 1900, 1907; Zirckle 1968; Leapman 2000). This is demonstrated to a remarkable extent in “hobby” groups of plants such as cacti and succulents (Rowley 1982). It is estimated that about 100,000 artificial orchid hybrids have been produced, some involving crosses between three or more genera (e.g., X Brassolaeliocattleya [Brassavola x Cattleya x Laelia]). Artificial hybridization between geographically distant congenerics has led to the creation of cultivars in important horticultural genera such as Rhododendron, Syringa, Magnolia, Hamamelis, and Rosa (Pringle 1981; Williams et al. 1990; Callaway 1994; Strand 1998). New horticultural hybrids of indigenous Proteaceae taxa are being developed in Western Australia and South Africa by both botanic gardens and horticultural research facilities (Considine 1993; Jansen van Vuuren et al. 1991; Jansen van Vuuren 1995; Sedgley 1995). Although botanic gardens are still involved in the breeding and release of new cultivars for horticulture (Winter and Botha 1994), hybridization work is largely now the concern of commercial horticultural breeders (Tobutt 1992; Uosukainen 1992). The most obvious impact of hybridization on ex situ conservation activities is the loss of genetic integrity of material in the collections, of openpollinated seed material, and seedlings derived from such collections that are intended for use in reintroduction or habitat restoration projects. The Convention on Biodiversity regulates and controls the collection of material for ex situ conservation “so as not to threaten ecosystems and in situ populations of species” (Glowka et al. 1994: 55). The release of hybrid plant material that could undermine a species recovery project or act as a
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damaging invasive would be contrary to Convention on Biodiversity biosafety concerns (Article 8h, 14; Glowka et al. 1994). High levels of artificial sympatry between species of any group with moderate or high crossability inevitably entail a high risk of spontaneous hybridization, and such hybrids have long been known to occur frequently in many garden and other collections (see Table 16.2). One of the first recorded spontaneous garden hybrids was Primula x kewensis, a fertile polyploid derivative from P. floribunda and P. verticillata (Newton and Pellew 1929), and there are many other examples of spontaneous hybrid origins of useful plants in cultivated garden and agricultural collections. Indeed, during the early development of garden plants before artificial hybridization techniques were understood and applied, when a keen eye for promising variants along with selection and skillful cultivation were the only tools available to gardeners, spontaneous hybridization is thought to have played a prominent role in the development of many garden plants (Meikle 1973). Among trees in arboreta, spontaneous hybridization has spawned new hybrids that have become widely used in horticulture and forestry. For example, Leyland cypress, X Cupressocyparis leylandii, a widely cultivated fastgrowing hedge tree, originated as a spontaneous hybrid between two introduced North American species in an arboretum at Welshpool in 1888. Similarly, ‘hybrid larch’, Larix x marschlinsii is a spontaneous hybrid between two introduced species which was first noted at Dunkeld, Scotland in 1904. This long history of spontaneous garden and arboretum hybrids highlights the obvious dangers of contamination of material of rare and threatened species, or other valuable genetic material, derived from ex situ collections. Spontaneous hybrids derived from garden stocks have already been noted for a number of threatened plant groups, including palms (Maunder et al. 2001b), Saintpaulia (Eastwood et al. 1998), Euphorbia (Turner 1998), Primula (Richards 1993), Lilium (Synge 1980), and Meconopsis (McAllister 1999). Cultivated stock of the Juan Fernandez endemic Wahlenbergia lorrainii (Campanulaceae, later treated as a synonym of W. fernandeziana), which is thought to be extinct in the wild (Ricci and Eaton 1994), is probably of hybrid origin, but the origin of the hybrid is unknown (W. fernandeziana x W. grahamiae; Lammers 1996). The majority of ex situ botanic garden collections and arboreta were neither planned nor designed with conservation in mind but rather as facilities for public display, education, and recreation. These have traditionally featured geographic displays (some geographically specific [e.g., Canary
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Island] or broad [e.g., Mediterranean] displays) or taxonomically focused displays (e.g., orchid, cacti, fern house, or oak collection) incorporating sets of congeners in close proximity. For example, normally allopatric species of Echium, Argyranthemum, or Limonium might be found in a single Canary Island display. Large specialist collections such as the United Kingdom National Plant Collections, with very large numbers of congeners assembled on one site (examples listed in Table 16.2), provide even more extreme examples of artificial sympatry. Many of these garden collections hold stocks of threatened taxa alongside numerous closely related nonthreatened taxa, but they are rarely managed specifically as part of conservation or recovery projects. Use of open-pollinated seed or seedlings derived from this sort of material in species recovery programs carries the clear risk that hybrid material from mixed collections could be unwittingly reintroduced (Maunder 1992; Maunder et al. 2000). Hybridization can severely compromise the utility of botanic garden collections as ex situ conservation resources. A good example is provided by the genus Echium. The propensity of the Macaronesian Echium species to hybridize in cultivation is well known. Indeed, a number of species described in the nineteenth century have subsequently been recognized to be hybrids of garden origin (Bramwell 1972). Echium pininana, endemic to the Canary Islands, Spain, is threatened in the wild but is a common component of the European botanic garden flora. It survives in cultivation as apparently secure botanic garden stocks and as extensive feral populations that can contain thousands of plants. However, molecular data indicate that these cultivated stocks of E. pininana have hybridized with other cultivated members of the section Simplicia (Maunder et al., unpublished data, 1996), making them superfluous for in situ conservation of the species and wasting limited ex situ resources. The ex situ conservation of other plant groups of conservation concern will face similar problems. For example, in the Aizoaceae, the dominant family in the Succulent Karoo biodiversity hotspot (sensu Myers et al. 2000), natural hybridization is limited by geographic isolation, flowering time, flower structure, and sometimes different levels of ploidy (Ihlenfeldt 1994), but distantly related species can be artificially hybridized (Hammer and Liede 1990), and the family is prone to hybridization in cultivation or after species introductions (e.g., in Carpobrotus [Gallagher et al. 1997; Vilá and D’Antonio 1998]). In contrast, the extensive ex situ glasshouse collections of highly crossable orchids (Peakall et al. 1997; Nielsen 2000), generally are less prone to spontaneous
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hybridization because of the lack of pollinators and the limited opportunities for seedling establishment. Other types of ex situ plant collections are similarly prone to problems posed by spontaneous hybridization. For example, research arboreta are maintained for a number of tree genera (e.g., Salix, Erythrina, and Leucaena; see Table 16.2). The primary objective of these collections is as a source of material for breeding and artificial hybridization programs aiming to improve these economically important tree crops. However, these collections also usually involve at least an implicit genetic conservation role as well. Ex situ collections of the tropical forage tree genus Leucaena illustrate the problems posed by spontaneous hybridization in such collections. When Leucaena species are brought together in cultivation, there is a 1 in 3 chance of any two species being cross-compatible (Sorensson and Brewbaker 1994). Two documented hybrids, L. x mixtec and L. x spontanea, and several other putative hybrids are thought to have resulted from indigenous circa situm cultivation in Mexico (Hughes and Harris 1994, 1998; Hughes 1998), and there are numerous reports of hybridization in ex situ field trials and germplasm collections (e.g., Bray 1986). The outcomes of spontaneous hybridization of Leucaena in ex situ collections include contamination of open-pollinated seed released from collections (Bray 1986); generation of new taxa, such as L. x spontanea, which may pose weediness hazards as great as those resulting from introduction of L. leucocephala (Hughes and Jones 1999); and taxonomic problems for plant identification (Hughes 1998). This set of problems is very similar to those documented for the genus Salix in Australia (Cremer at al. 1995). In contrast to strict ex situ conservationists, Leucaena and Salix collection managers seem to accept that some contamination of open-pollinated seed is inevitable. Indeed, novel Leucaena hybrids usually are seen as positive and interesting additions to collection and plantation diversity rather than as threats. In such examples, if the distribution of undesirable hybrids is to be avoided, growers should examine and rogue out any atypical hybrid seedlings.
Gene Flow from Ex Situ Collections and Genetic Contamination of Wild Populations The mixing of gene pools of formerly distinct taxa by introgression has been variously called contamination, genetic assimilation, infection, genetic deterioration, genetic pollution, genetic swamping, genetic takeover, and genetic
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aggression (Rhymer and Simberloff 1996). Contamination is problematic because in general we seek to conserve morphologically defined and taxonomically labeled conservation units, but contamination would remain equally problematic using other criteria such as management of evolutionary lineages, sensu Newton et al. (1999). The pejorative terms mentioned earlier reflect the view of hybridization as a destructive rather than constructive process and emphasize the role that hybridization can play in species extinction. The idea that hybrids have reduced conservation value is also enshrined in legislation (e.g., the hybrid policy of the U.S. Endangered Species Act; O’Brien and Mayr 1991; Soltis and Gitzendanner 1999). Hybridization has been implicated as an important factor contributing to plant extinction by causing reductions in population viability through production of hybrid seed at the expense of conspecific seed; production of fertile hybrids, the decrease in proportional representation of the rare species, and the associated decline in proportion of “pure” progeny; competition with hybrids for establishment of microsites and resources; and increase in both herbivore and pathogen pressure, decreasing population growth, given that hybrids may support a greater diversity of pests and may act as staging posts for pest colonization of parental species (Brochman 1984; Levin et al. 1996; Rhymer and Simberloff 1996; Huxel 1999; Carney et al. 2000). Ex situ collections can contribute to contamination of populations of rare and threatened species either directly through hybridization between cultivated stocks, between cultivated stocks and adjacent wild populations, or indirectly through hybridization between naturalized escapes and wild populations (Table 16.1). We have found very few documented cases of hybridization or gene flow between ex situ collections and wild populations of threatened species. This is not surprising given the small areas occupied by ex situ facilities, the very limited direct contact they have with surrounding wild populations, and the small size of ex situ populations. However, there are a number of cases of hybridization involving garden escapes and wild populations of rare and threatened species. Again, this is not surprising given the much wider spread of naturalized or cultivated escapes that enhances their chances of coming into contact with wild populations of congeners. A number of examples illustrate this. The restricted California endemic Oenothera wolfii (Onagraceae) is threatened by hybridization with the garden escape O. glazioviana (Imper 1997). In turn, it is thought that O. glazioviana is a widely naturalized European horticultural hybrid between two North American taxa, O. grandiflora and O.
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elata (Anonymous 1992). Planting of European horticultural hybrids of the epiphytic cactus genus Schlumbergera in Brazilian gardens adjacent to threatened populations of congenerics has spawned a number of putative hybrids (N. Taylor, pers. comm., 1999). Studies by Shapcott (1998) on the threatened Australian palm Ptychosperma sp. have revealed hybrids between wild populations and exotic congenerics in nearby urban gardens. Similarly, in the Mascarene Islands there are fears that introduced ornamental plantings of Hyophorbe palms will hybridize with the endemic Hyophorbe populations (Strahm 1989; Maunder et al. 2002). A particular risk occurs where threatened species are being cultivated in a landscape with high concentrations of congenerics, both wild and cultivated. Experience at the Chicago Botanic Garden has shown that two threatened Echinacea species (E. tennesseensis and E. laevigata) can cross with any other Echinacea species (J. Ault, pers. comm., 2002) and that horticultural selections of Echinacea can cross with adjacent wild populations (van Gaal et al. 1998), raising the possibility of introgression and the transfer of insects between host plants via “hybrid bridges” (Floate and Whitham 1993). Ironically, the growing popularity of naturalistic landscaping using indigenous species may increase the opportunities for hybridization. As urban development expands and establishes new garden populations adjacent to wild habitats, the opportunity for gene flow between wild populations and horticultural congenerics increases. This phenomenon is of particular concern around expanding urban developments in areas of diverse mediterranean scrublands in California, the Cape Province of South Africa, and Western Australia (Low 1999). On oceanic islands the introduction or escape of interfertile congenerics can have dramatic consequences. Examples from the Juan Fernandez Islands include the establishment of robust and fertile Gunnera hybrids (Pacheco et al. 1991) and hybridization between the introduced Acaena argentea and the endemic Margyricarpus digynus (Rosaceae) (Crawford et al. 1993). Francisco-Ortega et al. (2000) review hybridization between endemic Macaronesian taxa and introduced congenerics, listing examples from Arbutus (Ericaceae) and Phoenix (Arecaceae).
Ex Situ Facilities as a Source of Novel Hybrid Invasives Exchange and movement of plants between collections have not only released invasive species into new areas but also multiplied opportunities
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for the generation of novel hybrids, which can also spread and in turn pose new threats of invasion (Abbott 1992; Abbott and Milne 1995; Hollingsworth et al. 1998; Ellstrand and Schierenbeck 2000; Milne and Abbott 2000). The importance of biological invasions as a conservation threat of global concern is now widely appreciated (Williamson 1996; Vitousek et al. 1996; Bight 1998; Ewel et al. 1999; Low 1999). Furthermore, it is likely that invasives will become increasingly prominent and important as changing land use patterns and climate change affect biodiversity (Davis et al. 2000; Sala et al. 2000). Ex situ collections of one type or another have played a part in the distribution of invasive plants around the world (Cronk and Fuller 1995). It is well known that many ex situ collections, of all types, are inherently leaky. Deliberate exchange of material between collections promoted by seedlists (Indices Seminum) has been and remains a major activity of many gardens and other collections. Indeed, species introductions via botanic gardens acting as staging posts for worldwide exchange have been and continue to be one of the most important sources of invasive plants (Cronk and Fuller 1995; Bight 1998; Low 1999). Formal exchange is augmented by informal and sometimes illegal collection of material by gardeners, horticultural enthusiasts, local farmers, research station workers, and other visitors to grow in their farms or gardens. Of equal or greater concern and impact are plants that jump the garden fence to escape into the surrounding habitats (Nelson 1994). Spartina anglica (Poaceae), the autopolyploid derivative of S. x townsendii that arose as a spontaneous hybrid between the introduced S. alterniflora and the indigenous S. maritima, has invaded large areas of salt marsh in the United Kingdom (Thompson 1991). Spartina hybrids between introduced S. alterniflora and native S. foliosa are invading and replacing the native parent species on the West Coast of North America (Daehler and Strong 1997). The spectacular and invasive Heracleum mantegazzianum (Apiaceae), introduced to the United Kingdom as a garden ornamental via botanic gardens, has hybridized with the native H. sphondylium to produce localized hybrid populations (Stace 1975). The history of the Lantana camara (Verbenaceae), one of the world’s worst invasive weeds (Cronk and Fuller 1995), has been complicated by multiple introductions into cultivation, hybridization among cultivated stocks, escapes from cultivation, backcrossing in new territories (e.g., in South Africa), and hybridization between introduced and wild indigenous
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taxa (e.g., in South and Central America; Stirton 1999). Populations of the Mediterranean thistle Onopordum (Asteraceae) that have become important weeds in Australia are products of extensive hybridization and introgression that have occurred after introduction (O’Hanlon et al. 1999). Ellstrand and Schierenbeck (2000) postulate that hybridization may result in critical evolutionary changes that create the opportunity for increased invasiveness, citing 28 well-documented examples of invasive taxa that have come into being after hybridization. They cite four reasons why hybrids may stimulate invasiveness: evolutionary novelty arising through recombination producing variants that are better adapted to certain environments; increase in genetic variation, which can be especially significant for introduced plants that often contain low levels of genetic diversity; fixed heterosis (as in Spartina anglica); and increased fitness resulting from the dumping of mutational load. Rhododendron ponticum (Ericaceae) is an example of an invasive garden escape that has gained additional genetic variation through hybridization, thereby extending the range of environments invaded. Spontaneous hybrids between R. ponticum from the Mediterranean and the more coldtolerant North American R. catawbiense have occurred in or around horticultural collections in the United Kingdom and seem to have greater cold tolerance than R. ponticum, apparently assisting spread into colder environments in eastern Scotland (Milne and Abbott 2000). Two other spontaneous Rhododendron hybrids were discovered among naturalized R. ponticum populations in the United Kingdom during the same study by Milne and Abbott (2000). Given the very high levels of artificial sympatry within and around the various large ex situ collections of Rhododendron in the United Kingdom and the known high crossability within the genus, there is clear potential for other spontaneous hybrids to arise and contribute further to the evolution of invasiveness in R. ponticum. Hybridization can also enhance invasiveness by restoring fertility in hybrid derivatives of naturalized or invasive species that are sterile and currently limited to vegetative spread. The genus Fallopia (Polygonaceae) in the United Kingdom provides a cogent example of this type of spontaneous hybridization that could have particularly profound effects on invasiveness. Fallopia japonica subsp. japonica, the Japanese knotweed, was introduced to the United Kingdom from Japan between 1825 and 1850 and widely cultivated as a garden ornamental in Victorian times and has subsequently escaped and spread to become one of the most serious invasive plants in the
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United Kingdom. It forms dense thickets in many areas, posing a serious threat to native riparian habitats. All species of Fallopia are gynodioecious, that is, they occur as separate hermaphrodite and male sterile (functionally female) plants. In Britain, only male sterile plants of F. japonica subsp. japonica are known, and a high proportion of the other widely cultivated and weakly naturalized species, F. sachalinensis, are also male sterile. Molecular evidence suggests that the plants across the United Kingdom and Europe are ramets of a single clone. This supports the view that F. japonica subsp. japonica does not currently reproduce by seed in the United Kingdom. Hybrids (F. x bohemica) between F. japonica subsp. japonica, F. sachalinensis, and the infamous “mile-a-minute plant,” F. baldschuanica, occur in areas where the species are now sympatric, apparently arising through the sporadic occurrence of functional gametes. These hybrids have partially or fully restored fertility, raising the possibility of seed production and dispersal in F. japonica after introgression, further enhancing its invasive tendencies, with potentially far-reaching consequences for eradication or control (Hollingsworth et al. 1998). Long time lags, often of several decades or even more than a century, between introduction and invasive spread are a common feature of many invasion trajectories (Lodge 1993; Scott and Panetta 1993; Hobbs and Humphries 1994; Williamson 1996; Reichard and Hamilton 1997; Ewel et al. 1999). In some cases the stimulus of hybridization needed for the evolution of invasiveness provides a possible explanation for such time lags (Ellstrand and Schierenbeck 2000). Simply put, few introduced species (or spontaneous hybrids) naturalize, and few of those become serious invaders, but those that do spread and invade can take decades or longer to manifest themselves or to evolve their invasive tendencies. This suggests that given more time, more invasive hybrids are likely to evolve. Many species introductions, garden escapes, and releases are still in the early stages of establishment, spread, and possible invasion. In addition, it is clear that many spontaneous hybrids go unnoticed or are misidentified pending the sort of very detailed studies (e.g., Hollingsworth et al. 1998; Milne and Abbott 2000) that are needed to reveal some more cryptic hybrids or basic taxonomic work on less well-known groups (see Hughes and Harris 1994, 1998; Hawkins et al. 1999). The 28 invasive hybrids listed by Ellstrand and Schierenbeck (2000) may be no more than the first indicators of what is to come.
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Once again, ex situ conservation collections can play their part in precipitating new spontaneous invasive hybrids either directly or indirectly (Table 16.1), and it is the often complex, indirect pathways via escapes, intertwined by multiple introductions, that appear to be the most important source of new invasive hybrids, rather than hybrids arising directly in or around collections themselves. Multiple introductions, both repeated introductions from the same source at different times and separate introductions of independent origin, which are common for many species introduced to or released from ex situ collections (see Urbanska and Landolt 1998; Milne and Abbott 2000 for discussion of multiple introductions of Rhododendron ponticum), can also enhance opportunities for invasion and hybridization (Ellstrand and Schierenbeck 2000).
Conclusions Hybrids are an important but underappreciated and somewhat poorly understood component complicating many facets of ex situ conservation programs. As might be expected under the extreme conditions of artificial sympatry encountered in ex situ collections, there is abundant evidence that spontaneous hybridization is common in many collections and that this can have important consequences for ex situ conservation programs because of the potential loss of genetic integrity of open-pollinated seed material or seedlings derived from collections for species recovery work or other purposes. One might expect that some of these hybrids could pose risks to wild populations outside collections, but in practice it appears that most hybrids arising within collections remain confined to ex situ facilities. This is probably because most do not survive to reproduce simply because they are weeded out as part of routine management. Furthermore, low levels of propagule pressure may limit invasions (Williamson 1996) so that spontaneous hybrid individuals, unless highly self-fertile or readily spread vegetatively, are unlikely to establish outside and spread from gardens. One might also expect a proliferation of hybrids arising directly between cultivated stocks and adjacent natural and seminatural populations. However, there seems to be a surprising dearth of documented examples of this type of direct interaction between ex situ collections and surrounding habitats. This is likely to result from the small areas occupied by ex situ facilities, which limit direct contact or sympatry with surrounding wild popula-
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tions, and from the small size of populations held in collections, which often comprise only a handful of individuals. Small source populations of this kind tend to result in very low levels of gene flow into larger sink populations nearby; of far greater importance are indirect routes (Table 16.1) between ex situ collections and wild populations that involve the much more widespread and abundant populations of naturalized escapes of various kinds. These much larger naturalized or released populations, often numbering many millions of plants, can result in artificial sympatry over much larger areas and in large pollen sources, greatly increasing the chances of hybridization compared with the extremely localized small populations represented by the ex situ collections themselves. These conclusions provide a clear basis for the practical guidelines we propose here. First, when the genetic integrity of seed derived from ex situ collections is important for conservation or other purposes, hazards of hybridization within collections must be borne in mind and measures taken to reduce those risks. Second, the risk of hybrid interactions between ex situ collections (including gardens) and wild populations are greatest after escape, so that general measures to reduce leakiness from ex situ collections will be the most important and effective measures for reducing conservation risks posed by hybridization. The need for ex situ management of threatened species and other particularly rare or valuable genetic resources will increase as wild populations continue to decline. Ex situ collections of threatened plants have a vital role to play in supporting species conservation. It would be a terrible irony if the long-term legacy of ex situ conservation included new hybrids and invasives rather than recovered species.
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Chapter 17
Accounting for Sample Decline during Ex Situ Storage and Reintroduction Edward O. Guerrant Jr. and Peggy L. Fiedler
Many events between propagule collection and successful reintroduction can influence the quantity and quality of an ex situ sample. To the degree that ex situ collections are to be successful in enhancing the long-term survival of sampled populations and species, the expected losses during various parts of the process must be anticipated and accounted for in both the original collection and subsequent management efforts. As part of an effort to improve its own expanding practice of ex situ plant conservation, the Center for Plant Conservation (CPC) developed a set of genetically based guidelines for collecting seed and other propagules for conservation purposes (CPC 1991). These guidelines have found wide application in ex situ plant conservation programs. These genetic sampling guidelines continue to evolve as conservationists worldwide contribute insights gained by practical experience working with threatened plant populations (Touchell et al. 1997). Despite these and other advances (Brown and Marshall 1995; Schoen and Brown 2001), some important issues remain to be resolved. In this chapter, we focus on two particularly troublesome issues that can dramatically affect estimates of appropriate sample sizes for conservation collections. The first is sample decline during ex situ storage, with the emphasis on detecting mortality in seed collections. The other concerns anticipating and minimizing losses that are likely to occur during the establishment phase of a reintroduction, when a sample is returned to the wild. In this study we assume that propagules are handled in the best possible way from the time of collection to their processing at an ex situ facility (Smith 1995) and that they are processed promptly and maintained under optimal conditions (Chapters 6 and 7 and Appendixes 2 and 3, this 365
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volume). Furthermore, we assume that germination (Chapter 8, this volume) and propagation procedures for the species are well established. Plant material is needed to learn how to maintain, germinate, and propagate a species before successful storage, retrieval, and reintroduction. Developing such knowledge often entails the use of large amounts of seed or plant material, and these needs must be factored into a collection plan. Challenges during storage are fundamentally different for plant taxa that must be maintained as arrays of growing plants and for those that can be maintained as dormant seed in a seed bank. Storing seed has many advantages over growing plants as a means to maintaining germplasm off site, but ease of monitoring survivorship is not among them. In contrast to growing plants, for which a simple visual observation usually is adequate to detect mortality or even reduced vigor, seeds often do not change appreciably in appearance when they die. Monitoring survivorship of dormant seeds in seed banks therefore is much more difficult than it is for growing plants. Detecting mortality rates or reduced vigor of stored seed entails repeated testing for viability or ability to germinate. Assessing mortality during storage is perhaps best viewed as a statistical sampling problem. What sample sizes are necessary to detect a given decline in survivorship with desired levels of statistical significance and power? The other major phase in which attrition can occur is during reintroduction attempts between initial planting and the times at which survivors reach reproductive maturity and begin to reproduce on their own. Such losses, which vary greatly in magnitude, are here called the demographic cost of reintroduction.
Ex Situ Storage as Seed Unlike growing plants after they die, dead seeds often are indistinguishable from living ones in their outward appearance. Even though survivorship and germinability are different, as a practical matter the most direct and effective way to monitor the potential usefulness of a seed collection is to germinate samples when they enter the seed bank and periodically thereafter. This is not as simple a task as it might seem. First, it is necessary to know how best to germinate the sampled population (Chapter 8, this volume). Germination requirements often are assumed to be species specific, yet they may differ significantly between populations, at least in widespread species. For example, Meyer (1992) demonstrated that the cold stratification requirement varied significantly among populations of Penstemon
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eatonii (Scrophulariaceae) collected from different parts of its range. To assess changes in germination behavior over time it is necessary to subject all samples of the selected seedlot to comparable conditions in each trial. Otherwise, germination rate differences could be attributable to environmental causes. Given that ambient outdoor conditions are not sufficiently similar between years to eliminate environmental differences as a potential cause, accurate and precise information about the survivorship of seed bank collections is best obtained by using controlled environment chambers to germinate seeds. The next hurdle is to interpret accurately the results of comparisons between different trials. Although the magnitude of what constitutes a significant decline is a subjective decision, sample sizes necessary to detect a given decline are amenable to analysis. In their Guidelines for the Maintenance of Orthodox Seeds, the CPC (Weiland 1995) suggest a 15 percent decline as a reasonable threshold to trigger a contingency action, either recollection or regeneration grow-outs. Weiland (1995) attributes this figure to crop scientists, noting that even with the large sample sizes with which they have to work, the viability often declines more severely before action can be taken. Ideally, statistical test results to determine declines in germinability of a seed sample accurately reflect the true condition of a seedlot (Figure 17.1). However, it is possible, through chance alone, that germination tests will indicate a decline when none has occurred. This is a type I, or false change error (Elzinga et al. 1998). The probability of making it can be considered the significance of the test. This is the p value commonly cited when a statistically significant difference is found. Alternatively, and again through chance alone, a test may fail to indicate a decline when one has occurred. This is known as a type II, or missed change error. Our ability to avoid it is known as the power of a test. In other words, the power of a test is a measure of how likely our test is to detect a given decline if there really is one. Of course, it is easier to detect a large decline than a small one, so it is necessary to designate the minimum detectable change when specifying the power of a test. There is no single sample size necessary to detect a given decline. Sample size can vary according to how tolerant you are of making the two kinds of errors, among other things. These are subjective decisions that involve tradeoffs. As the desired significance of a test increases, with all other factors held equal, power declines. The conservation implications of the two error types differ as well. Whereas a false
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Figure 17.1 Possible combinations of actual viability changes and test results. A sample’s viability will have either declined or not, and germination tests will either detect a decline or not. In two of the four possible combinations of the two variables, the two views concur, and there is no error. Two types of errors are also possible: in a type I error, a change is identified where one did not occur, and in a type II error, a real change is missed.
decline error can lead to unnecessary collection or grow-out, a misseddecline error could result in the extinction of the species by failing to recollect or grow out a sample when those actions would have been possible. Sample sizes necessary to detect a given decline also vary with the initial germinability of a seedlot. As an example, Figure 17.2 illustrates the relationships of statistical power as a function of sample size differences when initial germinability is either 90 percent (0.9) or 50 percent (0.5) and the desired significance of the tests is either p = 0.1 or p = 0.01. The five lines on each of the four graphs represent different germination rates in a subsequent test. Germination rates are indicated by the numerical value on the graph, with the lower, less steep lines indicating greater survivorship. In all cases statistical power to detect a given decline increases with increasing sample size, that is, all lines trend from the lower left to the upper right. The graphs can be used in several ways. For example, if the initial germination rate of a sample is 90 percent (Figure 17.2B, D) and the desired minimum detectable difference is an absolute decline of 20 percent to a subsequent germination of 70 percent, the line indicated by 0.7 is the one to consider. For example, if a statistical power of 0.9 (indicated on the y-
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Figure 17.2 Relationships of statistical power as a function of sample size differences when initial germinability is either 90 percent (0.9; B, C) or 50 percent (0.5; A, C) and the desired significance of the tests is either p = 0.1 (A, B) or p = 0.01 (C, D). Sample sizes refer to the number of seeds used in each test, not the sum of two or more tests.
axis) is desired, a sample size of about 100 seeds for each test would be necessary if the p value chosen is 0.01 (Figure 17.2D). In contrast, if the significance of the test is relaxed from p = 0.01 to p = 0.1, a sample size of only about 50 seeds would be necessary to achieve the same degree of statistical power (Figure 17.2B). Alternatively, the graphs can be used to estimate statistical power given a particular sample size a practitioner is willing to use. If the choice is made to use a sample size of 100 seeds per test, the initial germination is 0.5, the significance of the test is set at p = 0.1 (Figure 17.2A), and minimum power of 0.9 is chosen, the minimum detectable difference is a drop from 0.5 to 0.3. If a p = 0.01 level of significance is desired (Figure 17.2C), the power to detect a drop from 0.5 to 0.3 germination would be about 0.7 rather than 0.9. Beyond the specific uses to which the graphs can be put, there are three general patterns to note in Figure 17.2. First, the power of a test increases
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dramatically as minimum detectable difference increases. Second, to detect a given decline for a given sample size, statistical power is greater if initial germination is 90 percent rather than 50 percent. Tests are least sensitive when initial germinability is 50 percent and are more sensitive toward either extreme. Third, statistical power increases with increased tolerance for making a false change error (e.g., accepting a p = 0.1 rather than p = 0.01 level of significance). Our analysis illustrates several dilemmas confronting seed bank operators. One is that sample size must be chosen before the initial germination rate is known. Pilot studies are helpful, but they use additional seed. Given the large sample sizes often needed to detect changes of a magnitude we might want to detect, sufficient seed may not be available to monitor a collection as closely as would be desirable. The challenge is particularly acute when seeds from each maternal plant are maintained separately, as opposed to bulk collections, resulting in a larger number of smaller accessions. This raises strategic questions about how precisely it is necessary to know the status of a collection and about how best to deal with uncertainty surrounding the status of a collection. Resolution of these issues awaits further discussion in the conservation community. If a sample drops below a critical threshold, it will be necessary to regenerate the sample with surviving seeds by growing them out to increase their numbers or to make additional collections from the wild populations. Retaining the genetic diversity of a sample is of critical importance. Schoen and Brown (2001) have shown that in order to have a 95 percent probability of retaining the alleles captured in the original sample, the size of the grow-out population must triple in each regeneration cycle. This just avoids random changes in allele frequencies and does not take artificial selection, which can be strong, into account.
Ex Situ Storage as Growing Plants Detecting mortality and even reduced vigor is much simpler in growing plants than in stored seed and can often be accomplished by visual inspection. Nevertheless, the overall challenge of maintaining adequate numbers of healthy and genetically appropriate individuals in growing collections is in many ways much more difficult than storing seed in a bank.
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To begin, what mortality rates might be expected in growing collections of ex situ material? The first factors to consider are life history and natural lifespans in the wild. Trees are generally long-lived, but the natural lifespans of woody and herbaceous perennials vary widely. Some might be able to live for decades if not centuries, whereas others live only one to a few years. Attempts to maintain ex situ populations of annual plants by growing them out every year are particularly difficult because each year the genetic constitution of the sample will progressively diverge from that of the source population. Figure 17.3 illustrates the documented survivorship of various accessions of three herbaceous perennials that have been maintained at the Royal Botanic Gardens, Kew, some for almost a half century (Maunder 1997). Superimposed on this array of empirical data are a series of lines, each representing a different constant annual mortality rate, ranging from 0.005 to 0.32, or 0.5 percent to 32 percent. The steeper the line, the greater is the mortality rate. No Caralluma (Asclepiadaceae) accessions survived for as long as 25 years, and all had annual mortality rates greater than 16 percent (i.e., all symbols are below the 0.16 line). In contrast, some accessions of both Nepenthes (Nepenthaceae) and Primula (Primulaceae) were maintained in an ex situ context for up to 35 years, with annual mortality rates of <1 percent. Nevertheless, other accessions of both Nepenthes and Primula had annual mortality rates in excess of 16 percent. Even taxa that can survive well in cultivation may not necessarily do so, despite having the best of horticultural care provided by knowledgeable people in committed institutional settings (see also Elias 1987). Even under the best circumstances, these figures may be overly optimistic because the survival of an accession is not necessarily the same as survival of the wild collected individuals themselves. An accession may represent a group of wild collected genotypes, and artificial selection may weed out particular genotypes, so the accession survives with a much lower level of genetic variation. If the species is undergoing seed regeneration, artificial selection and hybridization may create a semidomesticated population that is more tolerant of artificial conditions; accordingly, the accession survives but the genetic makeup is very different from that of the wild source population. That said, some individual specimens can live for centuries in botanic gardens. For example, cycads planted in the eighteenth century in botanic garden glasshouses in Holland and the United Kingdom, as well as a palm (Chamaerops humilis, Arecaceae) planted in 1585 in Italy’s Orto
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Figure 17.3 Survivorship of various growing collections of three perennial
taxa at the Royal Botanic Gardens, Kew, against a backdrop representing survivorship over time of various constant mortality rates between 0.005 and 0.32 annually. All Caralluma collections suffered average annual mortality rates between 0.16 and 0.32. Mortality rates in different Nepenthes collections vary from zero mortality in a 35-year-old accession to an annual mortality rate greater than 0.16 per year in an accession maintained for only 5 years.
Botanico di Padua, have survived into the twenty-first century (Maunder et al. 2001). Outright mortality is not the only danger to off-site growing collections. Diseases and pests can be acquired, and, perhaps even more insidious, the genetic integrity of the samples can be eroded over time through founder effects, genetic drift, genetic bottlenecks, artificial selection, and hybridization (Chapters 12–16 and Appendix 3, this volume). All of these can render garden-grown plants unsuitable for reintroduction into the wild. For example, the Malayan rubber industry was formed on a foundation of only nine plants, representing a mere 0.01 percent of an original sample of 70,000 Hevea brasiliensis (Euphorbiaceae) seed collected by Wickham and sent to the Royal Botanic Gardens, Kew, and then to the Singapore Botanic Garden (Purseglove 1979). Bennett and Bennett (1992) found significant variation in the amount of nuclear DNA between cultivated and
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wild Milium (Poaceae) samples, which they attributed to selection during cultivation in high-nutrient environments. Briggs et al. (1992) showed that rapid selection for herbicide tolerance developed in different genetic strains of Senecio vulgaris (Asteraceae) in Cambridge, England. When ex situ samples must be cultivated to be maintained, the opportunities for genetic change are manifold.
Minimizing Losses in Growing Collections The obvious foundation for maintaining conservation plants in an ex situ setting is to provide them with consistent, appropriate, high-quality care. This consists of providing not only a suitable environment in which to grow but also consistent, ongoing care. Proper ex situ cultivation and care are beyond the scope of this chapter (see Appendix 3, this volume).
Estimating the Demographic Cost of Reintroduction It is not reasonable to expect that all seeds collected for a conservation project will be used in reintroduction attempts and result in established seedlings, nor will all reintroduced seedlings survive to reproduce successfully. To estimate the range of population size decline that might be expected to occur between planting and successful reproduction of founding individuals, we used empirically derived stage-based transition matrices from the literature as a basis for stochastic modeling. We used data from six field-based studies of taxa representing a range of life histories: iteroparous perennials with tree, shrub, and herbaceous habits (Table 17.1). All populations modeled have a lambda () greater than 1, indicating that they described growing populations projected to increase in size over time. These analyses thus are inherently optimistic and assume an underlying dynamic of a growing population. We used RAMAS/Stage (Ferson 1990) to model the expected attrition in numbers after reintroduction of 1,000 individuals of the smallest size class described in each study. A more detailed description of transition matrix models and their use in conservation is provided in Chapter 15 (see also Caswell 2001; Menges 1986; Guerrant 1996). Stochasticity was introduced into the models at each time step by randomly generating values for each element in the matrix independently, from empirically based distributions for each of the matrix elements. For each matrix, the values were
table 17.1 List of taxa modeled, with life history, transitions, lambda, and source. Taxon
Life History
Astrocaryum mexicanum Calathea ovandensis Calochortus howellii Erythronium elegans Fumana procumbens Panax quinquenifolium
I, P, I, P, I, P, I, P, I, P, I, P,
W, T H H H W, Sh H
Transitions
6 4 6 5 6 4
a
1.012 1.036 1.000 1.042 1.017 1.045
Minimum Mean
Time to Reach Minimum Mean (years)
98 190 590 400 520 15
31 4 16 7 4 3
Source
Pin˜ ero et al. 1984 Horvitz and Schemske 1995 Fiedler et al. 1998 Guerrant 1999 Bengsston 1993 Nantel et al. 1996
Life history: I, iteroparous; P, perennial; W, woody; T, tree habit; H, herbaceous; Sh, shrub habit. All simulations began with 1,000 founders. a Different sites used.
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constrained such that no element or column added to more than unity (except considering the reproductive contributions). This contrasts with the method used by Menges, Guerrant, and Hamzé in Chapter 15, in which they introduced stochasticity by randomly selecting whole matrices from among the available empirically derived alternatives. Greenlee and Kaye (1997) and Kaye and Pyke (2003) discuss some of the implications of these two methods of modeling stochasticity. Ours was simply a pragmatic decision based on the software used and does not imply that element selection method is superior to whole matrix selection. Our results are intended to indicate relative attrition rates and to provide order-of-magnitude estimates of possible attrition between planting and successful establishment that can reasonably be expected to occur. Simulations for each species, or scenario within species, were run over a period of 100 years and replicated 1,000 times. Figure 17.4 illustrates the course of mean sizes of surviving populations over time, for a period of 35 years. Of particular interest here are the minimum mean sizes to which the surviving populations fell and how long it took for each to reach the minimum size. As indicated in Table 17.1 and Figures 17.4 and 17.5, the demographic cost of reintroduction can be substantial. In the most extreme case, an outplanting of 1,000 Panax seedlings would on average be expected to drop by more than 98 percent to just 15 individuals within 3 years before the populations began to rise. But many simulated runs of this and other taxa ended with extirpation before any increase could begin. Panax was also projected to be the most rapidly growing of the modeled taxa, at an annual rate of 4.5 percent ( = 1.045). Note that equilibrium population growth rates and expected pre-reproductive survivorship are not positively correlated. If anything, these limited data suggest a negative correlation, a possibility that should be pursued further. If newly established populations are to have anything like the genetic diversity of the ones from which the founders were collected, expected losses during reintroduction must be accounted and compensated for in the original collection. One way to express the magnitude of losses is by dividing the initial size (1,000) by the minimum population size. This generates a multiplication factor, or coefficient of decline (Figure 17.5), by which the initial estimate of collection size must be multiplied to compensate for expected losses. In the case of Panax, 1,000 divided by 15 means that 67 times as many plants
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Figure 17.4 Simulation results for six perennial taxa, showing mean population size trajectories for modeled populations all beginning with 1,000 founding individuals of the smallest size category of propagule studied for each taxon. Symbols on each line represent the minimum mean size to which the modeled population declined before beginning to rise. Numbers in parentheses after name refer to the value of lambda, or population growth rate. Note that the minimum size to which a population declined before beginning to rise is not necessarily correlated with population growth rate.
as are needed to reach sexual maturity will need to be planted (Figure 17.5A). At this level of attrition, if the reintroduction goal is to have 1,000 founder Panax plants reach sexual maturity, it would be necessary to plant almost 67,000 seedlings. Of course, these data are simulated results based on wild populations with positive growth rates. One assumption of this modeling is that outplanted individuals will behave demographically identically to naturally occurring plants. This may be unduly optimistic, at least without extensive horticultural care provided until the plants successfully reproduce. Figure 17.5B illustrates empirical results of actual reintroduction attempts, some of which fared worse than any of the simulations. Another assumption is that the series of years for which data were gathered in the field accurately reflect what will happen during a reintroduction. Presumably
Figure 17.5 Graphic representation of the coefficient of decline, or the degree to which founding population size collections would
need to be multiplied to compensate for pre-reproductive mortality. Compare (A) model data with (B) selected empirical data on various reintroduction attempts and with (C) the establishment-year results of 27 sets of field germination tests conducted over a 5-year period with the rare herbaceous perennial Erythronium elegans. (Data for Astragalus osterhoutii, C. Dawson, pers. comm., 1999; A. cremnophylax var. cremnophylax, Maschinski 1993; Abronia umbellata ssp. breviflora, Kaye 1998; Warea, Wallace 1990; Lilium occidentale, Guerrant 2002 and unpublished data, 2002; Arabis koehleri var. koehleri, Guerrant, unpublished data, 2002; Erythronium elegans, Guerrant 1999.)
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there will be many stochastic environmental effects that cannot be anticipated but will affect establishment. For example, in a series of 27 separate field germination and seedling establishment trials with Erythronium elegans (Liliaceae) that were set out annually over a 5-year period, seedling establishment varied from 0 to 94 percent (Figure 17.5C; Guerrant 1999). All seeds used were planted the year they were produced, in microsites similar to and in very close physical proximity to where they were collected. Therefore, differences cannot be attributed to genetic causes, but to variation in environmental conditions. Some years were simply better for germination and seedling establishment than others. Attrition can be high and vary greatly between years and also between sites within a year. The implications for collection guidelines intended to support even one reintroduction attempt are daunting. To compensate for losses of these magnitudes, our results suggest that initial sample sizes may need to be orders of magnitude greater than current recommendations. Unfortunately, such collections may be too great for sampled populations to bear or prohibitively expensive in time and other resources needed to collect, store, and monitor. In addition to increased sample sizes, other ways to compensate for potential losses associated with reintroduction must be explored.
Reducing the Demographic Cost of Reintroduction To reduce the demographic cost of reintroduction, it is necessary to increase the proportion of founding individuals that survive to reproduce successfully. There are a variety of ways to pursue this goal (Guerrant 1996). All else being equal and given the resources to produce, outplant, and possibly even care for them, using larger, more mature individuals of some species to found populations, should, theoretically, provide a higher probability of success than using seeds or smaller plants as founders (Guerrant 1996). Unfortunately, not all species lend themselves to this approach because larger plants of some species do not transplant as successfully as seedlings. And even for those that do transplant well, other pragmatic considerations may limit the usefulness of this approach in some situations. Nevertheless, the results of demographic modeling here showed that relative to populations in which smaller founder individuals were used, those founded with larger individuals did not suffer nearly as much mortality before the plants reached reproductive stature and the simulated popula-
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Figure 17.6 All other factors held equal, the use of older or larger size classes
of founders is expected to reduce the demographic cost of reintroduction. Simulation results show that both the demographic cost and the time it took simulated populations to reach their minimum size progressively decreased as the sizes of founding individuals increased. For horticultural reasons not all taxa lend themselves to this approach equally well.
tions began to increase in size (Figure 17.6; see also Guerrant 1996). Of course, there are practical limits to such tactics, but even a marginal difference could well make the difference between success and failure. Any reasonable means by which a greater proportion of an off-site sample can be translated into mature, reproductive plants should be expected to have many benefits, not the least of which is to minimize collection pressure on the original population. Although the theoretical benefits are great, we are aware of few empirical data that bear on the question. In an ongoing experimental reintroduction of the endangered Lilium occidentale, Guerrant (2001, 2003) has been monitoring emergence and growth rates in the wild of plants derived from either new or stored seed or year-old bulbs (Figures 17.5 and 17.7). For each of the 6 years after planting, plants derived from bulbs have consistently emerged in significantly greater proportion (56 percent after 6 years) than plants derived both from new seed (40 percent) and from seed that had been stored for 1 or 2 years (32 percent). In another reintroduction experiment involving Arabis koehleri var. koehleri (Brassicaceae), the establishment rate is lower, and the disparity between the success of propag-
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Figure 17.7 Results of the first 6 years of an experimental reintroduction of the endangered western lily (Lilium occidentale) plotted against a backdrop of modeling data for six perennial taxa (Figure 17.4). Empirical results in which the fates of half-sibling families of plants that were placed out as yearling bulbs or seed that had been stored in a seed bank show that plants put out as bulbs have consistently emerged at higher rates than those set out as seed (Guerrant 1999 and unpublished data, 2002).
ules planted as seedlings or seeds appears initially to be much greater. At the end of the first spring after a fall planting, less than 10 percent of seeds produced seedlings, whereas just over 50 percent of seedlings survived (Figure 17.5, Guerrant, unpublished data, 2002). These preliminary results are consistent with the hypothesis advanced by Guerrant (1996) that using larger, more advanced individual plants as founders should result in higher survivorship and thus lower extinction risk. In their investigation into the effects of planting different life stages, Bell et al. (2003) have taken an additional step by comparing demographic characteristics of restored populations of Cirsium pitcheri (Asteraceae) and Asclepias meadii (Asclepiadaceae) with viable natural populations. For C. pitcheri, a short-lived monocarpic perennial, they estimate that establishing a viable population would entail planting about 1,600 seedlings or about 250,000 seeds, which demonstrates the potential magnitude of the demographic cost of reintroduction. Based on their comparative demographic
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analyses of reintroduced and naturally occurring individuals, the first such data of which we are aware, they urge caution in generalizing the results. Not surprisingly, the demographic characteristics of transplanted and naturally occurring C. pitcheri individuals differed somewhat, with transplants flowering earlier and being less fecund. Likewise, the results for A. meadii, a long-lived herbaceous perennial, seem to be consistent with the hypothesis that it is more efficient to use greenhouse-generated seedlings as founders than it is to use direct seeding. Unfortunately, because it is so longlived and successful reproduction in the wild so infrequent, they had not yet gathered sufficient data on the wild population to make firm comparisons between reintroduced and naturally occurring populations. In addition to the survivorship benefits of outplanting larger individuals, plants may reach much larger sizes and reach reproductive maturity at a younger age when produced in a garden or other off-site setting. The demographic benefits of reducing the age at first reproduction are great and well known. Conservation horticulturists with the New England Wildflower Society, at the Garden in the Woods, were able to grow Potentilla robbinsiana (Rosaceae) plants to reproductive maturity in just a couple of years, which is roughly one-fifth the time it takes in the wild (B. Brumback, pers. comm., 2001). In part because of the successful reintroduction efforts of the New England Wildflower Society, a member of the CPC, P. robbinsiana was removed from the federal list of endangered and threatened species in 2002. Another, complementary approach is to protect plants, especially seedlings, from herbivory. Using the threatened thistle Cirsium pitcheri as a model, Bevill et al. (1999) experimentally excluded insect herbivores from juvenile rosettes in two natural populations. In the population where insect herbivory was high, they were able to increase juvenile plant survivorship from 40 to 93 percent. They also documented a tenfold increase in seed production in juveniles that matured and flowered. They argue that even small-scale manipulation of some interactions between rare plants and their natural enemies can be an effective and low-cost management tool. In a similar vein, Raven (1995) was able to almost double the proportion of naturally established Erythronium elegans flowering plants that produced seed in 1993 from 18 to 34 percent by protecting the plants from large vertebrate herbivores with portable wire cages. The pattern was repeated in 1994 (0 seeds produced in plants subject to herbivory and 162 in caged plants). Thus, it seems that small-scale management can increase not only the survivorship of founding individuals but also the numbers of seeds available for collection.
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Another way in which the demographic costs of reintroduction can be ameliorated is with the use of clonal material. There are a variety of ways in which clonal material can be produced and used for conservation purposes. Clonal material of some taxa can be generated using standard horticultural techniques such as cuttings, air or ground layering, and dividing. More recently, the use of in vitro, or tissue culture, techniques has been applied to rare plant conservation. Sugii and Lamoureux (Chapter 10) and Pence (Chapter 11) describe tissue culture techniques that are used for conservation purposes. These authors typically use in vitro methods not to produce large numbers of genetically identical individuals but rather to grow genetically unique plants from immature seeds, or spores. Though useful or even essential in some cases, in vitro methods, especially those used to produce multiple copies of particular genotypes, run the risk of inducing genetic damage if not done carefully. Clonal multiplication is not a panacea for plants with low population sizes because all individuals produced are genetically identical. Nevertheless, there are several ways in which clonal material can be used to conservation advantage. The basic advantage provided by cloning is to spread the risk of mortality over more individuals (ramets). Clonal material can be used simultaneously in one or more reintroduction attempts, while the genetic material continues to be maintained at the ex situ facility, and be distributed to other facilities. Although the genetic implications must be evaluated on a case-by-base basis, multiple copies of any or all genotypes can be outplanted, thus increasing the chances that any particular genotype will survive to reproduction. Finally, cloning allows the plants themselves to be used as “phytometers” to identify the best available habitat for reintroduction.
Conclusions Ex situ collections can be extremely valuable as source material for use in restoring diversity, literally making the difference between extinction and survival. But for reintroduction to be successful, a sufficiently large quantity of genetically appropriate material must be collected to ensure survival during storage and the rigors of establishment before they can reproduce themselves in their new habitat. Thus, to realize the potential conservation value of an ex situ collection, it is not enough just to collect a genetically representative sample for storage. The probable kinds and magnitudes of some losses can be estimated in advance and thus be compensated
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for by increasing the sample sizes of the original collections. In this chapter we have attempted to identify and consider quantitatively two likely sources of attrition that can occur between collection and successful use. The main purpose of this exercise is to better inform collection strategies and guidelines (Appendix 1, this volume). We cannot reasonably expect all seeds to survive ex situ storage long enough to have conservation value, no matter how good the conditions are. Given that dead seeds generally look like living ones, it is necessary to conduct germination trials to monitor the utility of collections. Sample sizes necessary to detect declines in germinability are a combined function of the minimum detectable change desired and seed bank operators’ tolerance for making both false change and missed-change errors. Overall, the sample sizes required to detect even seemingly large viability declines, on the order of 15 to 25 percent, are remarkably high, even if the significance and power levels of the tests used are moderate (e.g., 0.1 and 0.9, respectively). The implications of and proper response to the realization that we will probably not have precise and accurate knowledge of the status of particular ex situ collections remains to be addressed. Likewise, we cannot reasonably expect that all propagules introduced into the wild will survive and reproduce. Based on modeling results of empirically derived data for plants with diverse life histories, the relative magnitude of expected pre-reproductive mortality—the demographic cost of reintroduction—can be expected to vary widely and at times be very high. Anything that can be done to increase the proportion of founders that survives to reproduce, such as planting larger founders or actively caring for them horticulturally, will decrease the demographic cost of reintroduction and increase the chances of successful reintroduction, the ultimate measure of effective ex situ management. In turn, anything that can be done to improve storage procedures and increase the efficiency with which stored material can be translated into reproductive plants in native habitats will reduce the collection pressure on wild populations.
Acknowledgments Many people contributed to the development of this chapter, and we would especially like to thank Tim Bell, Marlin Bowles, Christopher Dunn, Kayri Havens, and Mike Maunder for their time and insights that have greatly improved the final product.
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Guerrant, E. O. Jr. 1999. Comparative demography of Erythronium elegans in two populations: one thought to be in decline (Lost Prairie), and one presumably healthy (Mt. Hebo). Final report on five transitions, or six years of data. Unpublished report prepared for the Salem District, USDI Bureau of Land Management, and the Siuslaw National Forest, USDA Forest Service. Guerrant, E. O. Jr. 2001. Experimental reintroduction of the endangered western lily (Lilium occidentale). Pages 201–212 in S. H. Reichard, P. W. Dunwiddie, J. G. Gamon, A. R. Kruckeberg, and D. L. Salstrom (eds.), Conservation of Washington’s Native Plants and Ecosystems. Seattle: Washington Native Plant Society. Guerrant, E. O. Jr. 2003. Experimental reintroduction of western lily (Lilium occidentale) at the New River ACEC: Results of the first six years of growth. Unpublished technical report submitted to the Coos Bay District, USDI Bureau of Land Management. Horvitz, C. C., and D. W. Schemske. 1995. Spatiotemporal variation in demographic transitions of a tropical understory herb: projection matrix analysis. Ecological Monographs 65(2):155–192. Kaye, T. N. 1998. Experimental reintroduction of pink sandverbena at four sites on the Oregon Coast: Coos Bay, New River, Tahkenitch, and Siltcoos. Unpublished report submitted by the Oregon Department of Agriculture Plant Conservation Biology Program to the Siuslaw National Forest, the Coos Bay District of the Bureau of Land Management, and the U.S. Army Corps of Engineers. Kaye, T. N., and D. A. Pyke. 2003. The effect of stochastic technique on estimates of population viability from transition matrix models. Ecology 84(6):1464–1476. Maschinski, J. 1993. Integrated conservation strategies for recovery of sentry milkvetch at the South Rim of Grand Canyon National Park. Pages 101–107 in P. G. Rowlands, C. van Ripper III, and M. K. Sogge (eds.), Proceedings of the First Biennial Conference on Research in Colorado Plateau National Parks. Transactions and Proceedings Series NPS/NRNAU/NRTP-93/10, USDI National Park Service. Maunder, M. 1997. Botanic Garden Response to the Biodiversity Crisis: Implications for Threatened Species Management. Unpublished doctoral thesis, Botany Department, University of Reading, UK. Maunder, M., B. Lyte, J. Dransfield, and W. Baker. 2001. The conservation value of botanic garden palm collections. Biological Conservation 98:259–271. Menges, E. S. 1986. Predicting the future of rare plant populations: demographic monitoring and modeling. Natural Areas Journal 6(3):13–25. Meyer, S. E. 1992. Habitat correlated variation in firecracker penstemon (Penstemon eatonii Gray: Scrophulariaceae) seed germination response. Bulletin of the Torrey Botanical Club 119(3):268–279. Nantel, P., D. Gagnon, and A. Nault. 1996. Population viability analysis of American ginseng and wild leek harvested in stochastic environments. Conservation Biology 10:608–621. Piñero, D., M. Martinez-Ramos, and J. Sarukhán. 1984. A population model of Astrocaryum mexicanum and a sensitivity analysis of its finite rate of increase. Journal of Ecology 72:977–991. Purseglove, J. W. 1979. Tropical Crops: Monocotyledons. London: Longman. Raven, A. 1995. A two-year study on the impact of herbivory, salal and pollination success on the Lost Prairie Erythronium elegans population. Unpublished report submitted to the Salem District of the USDI Bureau of Land Management.
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part four
Using Ex Situ Methods Most Effectively Part IV is a synthesis of current thinking and practice in ex situ plant conservation. It begins with a chapter placing the practical application of ex situ conservation in the context of global declines in plant diversity and the constraints of an often inadequate conservation infrastructure for most areas of the world. Following in the tradition of the previous two volumes that have come from the Center for Plant Conservation (CPC) experience (Falk and Holsinger 1991; Falk et al. 1996), we offer a series of practical guidelines intended to provide guidance for the collection, ex situ storage, and cultivation of wild plant samples. The volume finishes with an appendix that reviews agencies and organizations currently undertaking or promoting ex situ plant conservation. Chapter 18 assesses the role of ex situ plant conservation in dealing with the ongoing biodiversity crisis, particularly examining the relevance of applying a technique largely developed in the temperate regions to high-diversity hotspots. Ex situ plant conservation in the tropics and other high-diversity regions is vital as part of an integrated approach. It must not only recognize the technical and horticultural limitations of the techniques in dealing with an expanding roster of threatened plants but also firmly link plant conservation to human needs, most notably poverty alleviation and welfare. The authors firmly believe that although ex situ conservation is, paradoxically, oversold by some agencies and special interest groups, it is also profoundly undervalued by the majority of conservation agencies and funding groups. The chapter describes the urgent need for investments in both infrastructure and horticultural skills to enable ex situ conservation to tackle accelerating species loss, and provides an overview of where the needs are most acute. The appendixes crystallize the best scientific information into practical guidelines that can be used to optimize the collection, storage, and cultivation of plant samples. Appendix 1 builds on the existing CPC
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guidelines for sampling from wild populations of rare and threatened plant species. The CPC guidelines have been updated in response to the latest research and the practical experience of the CPC network. They are designed for adoption as a practical tool for managers and plant conservationists. Appendix 2 provides scientifically tested, practical guidelines for seed storage developed through experience at the U.S. Department of Agriculture National Seed Storage Laboratory, which is part of the National Center for Genetic Resources Preservation in Fort Collins, Colorado. These guidelines will enable managers to assess the storage needs of wild collected seeds and thereby prolong the effective life of often scarce and small seed samples. Appendix 3 provides guidance on managing living plant collections to minimize genetic or demographic problems that can profoundly limit the conservation value of a collection. This is a new set of guidelines intended to enable managers to assess and manage risks associated with living collections. Often unrecognized or underappreciated, these risks can be reduced through careful population and horticultural management. In addition to the “Guidelines for a Rare Plant Reintroduction Plan” (Falk et al. 1996), the CPC now offers practical guidelines for all three major components of an ex situ capability: gathering a genetically representative sample; storing samples off site, as seeds or growing plants, for a long period of time; and restoring samples to their native habitats should that be needed for recovery. The CPC is not the only organization involved in such activities or the only one that seeks to provide guidance to practitioners around the world. Appendix 4 reviews the major national and international agencies and organizations associated with ex situ plant conservation, illustrating the variety of institutions and objectives involved and great strides that have been made. References Falk, D. A., and K. E. Holsinger (eds.). 1991. Genetics and Conservation of Rare Plants. New York: Oxford University Press. Falk, D. A., C. I. Millar, and M. Olwell (eds.). 1996. Restoring Diversity: Strategies for Reintroduction of Endangered Plants. Washington, DC: Island Press.
Chapter 18
Realizing the Full Potential of Ex Situ Contributions to Global Plant Conservation Mike Maunder, Edward O. Guerrant Jr., Kayri Havens, and Kingsley W. Dixon
Toward a More Integrated Approach Ex situ plant conservation is a portfolio of scientifically based techniques available to support the primary objective of retaining plant diversity in the wild. Each has its own strengths, weaknesses, and spheres of applicability (Box 18.1). There is an increasing recognition among both practitioners and authors that to be most effective plant conservation efforts should incorporate a wide variety of complementary techniques. This integrated approach (sensu Falk 1990) is overcoming the traditional schisms between the institutions and processes of in situ and ex situ plant conservation. Most global investment for ex situ conservation has been for genetic resource conservation, that is, the retention of plant material of value for crop breeding (Cohen et al. 1991). These activities are recognized as legitimate investments by both governments and development agencies. We maintain that comparable or greater levels of investment are essential to successfully retain wild plant diversity, on which sustainable development and, ultimately, life depend. The majority of the world’s plant diversity can be retained only through habitat and ecosystem conservation, a response that will require immense social, economic, political, and scientific investment (Cowling and Pressey 2001). Opportunities for extensive ecosystem reserves, though extremely valuable, are decreasing as modern agriculture, logging, plantation, and mining concessions damage high-diversity ecosystems, including ancient anthropogenic landscapes and their agricultural heritages. Tropical areas 389
box 18.1 Techniques for Ex Situ Management of Wild Plant Diversity: Opportunities and Limitations Cryopreservation: Seeds, pollen, or tissue is stored frozen in liquid nitrogen. This technique is used for the long-term storage of agricultural and horticultural taxa and increasingly for wild species. It is effective for the conservation of material over long time periods, but it entails a significant initial capital investment and access to trained technicians and supplies of liquid nitrogen. Cryopreservation is expected to become more useful as techniques for tropical species are developed; over time it should reduce the need for field gene banks. Seed banking: Seeds are stored in conditions of low moisture and temperature. This technique is routinely used for orthodox seeds of crops and wild species. It is an effective storage technique for orthodox seeds, and it allows storage of many genotypes over extended periods of time. Facilities can vary from desktop (seed stored in airtight conditions over silica gel) to freezers and large-scale gene banks with walk-in seed vaults. Tissue culture storage: Somatic tissue is stored in vitro under temperature- and light-controlled conditions of slow growth. This technique is effective for the conservation of material over short time periods, but it entails a significant initial capital investment and access to trained technicians and laboratory supplies. It should be used as a propagation rather than storage technique, with cryopreservation replacing tissue culture storage for long-term storage. Tissue culture propagation: Somatic tissue and seed are propagated in vitro; this technique is used for the proliferation of clonal plants and controlled seedling production. It is effective for the propagation of difficult material (e.g., small amounts of vegetative tissue, immature seeds), but it entails a significant initial capital investment and access to trained technicians and laboratory supplies. Cultivation in a dedicated conservation facility: Plants are cultivated under a specific horticultural regime with the aim of cultivating and propagating threatened species. Conditions are managed to minimize artificial selection, hybridization, and disease transmission. This is usually a shortterm activity to produce material for recovery activities or to bulk up material for long-term seed storage. Specialist cultivation in a controlled environment: Plants are cultivated in an artificial environment (e.g., tropical species in heated glasshouses in temperate regions). A very high horticultural investment is needed, and space limitations often preclude adequate genetic representation.
Cultivation in mixed display or reference collections: Plants are cultivated as part of reference collection under ambient environmental conditions. The majority of holdings in botanic gardens and arboreta are held in such collections, where the focus is on taxonomic representation or horticultural display. This method carries a high risk of artificial selection, hybridization, genetic drift, and disease transmission. Field gene bank: An open-air extensive planting used to maintain genetic diversity, most often for woody species. Field space allocation allows extensive genetic representation but represents a very long-term horticultural investment with high recurrent maintenance costs. There is a risk of hybridization between accessions and taxa. Commercial cultivation: Plants are cultivated as a profit-generating horticultural or agricultural crop, where management is dictated by commercial pressures. There are few opportunities for genetic management or control of provenances. However, this method can generate large numbers of individuals of threatened taxa that can be used for the landscape and horticulture trade, thereby encouraging the planting of native taxa and possibly reducing collecting pressure on wild populations. Community garden: Plants are cultivated in village- or communitymaintained plots. Management is dictated by local community needs and available resources, with few opportunities for genetic management and control of provenances. Community gardens are effective in maintaining valued plant resources, but carry a high risk of artificial selection. Inter situ: Plants are cultivated in horticulturally maintained nearnatural conditions, such as a managed population within restored seminatural vegetation. This technique is effective in maintaining populations of threatened plants when natural habitats are extensively degraded. This technique allows horticultural management (e.g., weeding) in seminatural conditions, such as weeded and fenced exclosures. In situ horticulturally managed populations: Wild plants undergo some degree of species-specific horticultural management, such as controlled hand pollination. This method allows the horticultural, genetic, and demographic management of threatened taxa without the inherent risks of moving plants or propagules to ex situ facilities. In situ managed wild populations: Wild plants grow in managed habitats that are subject to community-level management such as burning of grasslands. In situ: Wild plants that are subject to prevailing or natural ecological processes and influences.
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especially are experiencing increasing human population growth accompanied by dramatic social and economic changes. The surviving major wilderness areas (sensu Mittermeier et al. 1998), as large undisturbed natural areas, offer the best opportunities for retaining ecosystem and evolutionary processes. Necessary as such areas are, these are not adequate in terms of scale and distribution. Furthermore, even relatively pristine habitats are becoming increasingly permeable to invasive species, pests and diseases, overharvesting, civil unrest, climate change, and other problems. Nevertheless, opportunities do exist to secure large areas of high-diversity habitat; a study by Pimm et al. (2001) suggests that a one-time investment of $4 billion could secure 2 million km2 of habitat and manage the 2 million km2 already protected.
Applying Ex Situ Techniques in Biodiversity-Rich Regions Ex situ techniques should play an increasingly important role in supporting habitat and ecosystem conservation initiatives. Ex situ plant conservation as both a scientific discipline and a practical management response has, for the most part, been honed and developed in temperate North America and Europe. These parts of the world are characterized by a fortunate combination of relatively low levels of plant diversity and high levels of economic resources. The traditional botanic garden collection approach, based on maximum alpha diversity, has not proved particularly effective in conserving either high-diversity plant groups (Maunder et al. 2001b) or hotspot floras (Maunder et al. 2001a). The effective and widespread application of ex situ tools, especially in biodiversity hotspots, will entail major changes in both philosophy and technology (Maunder 1994). A large, if not the largest, proportion of the world’s botanical diversity resides in economically poor nations. Many of these nations have suffered a serious decline in resources available to inventory and conserve plant diversity (Barrett et al. 2001; Whitesell et al. 2002). In some cases, however, the infrastructure to support plant conservation has improved dramatically over the last 30 years, such as the growth in botany in Brazil. In many other parts of the world, such as central Africa, the botanical infrastructure has been at best static. This imposes severe limitations on the institutional infrastructures that house the small and underresourced cadres of taxonomists, systematists, ecologists, and other biodiversity specialists. Only 6 percent of the world’s scientists live in the countries that
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house 80 percent of the planet’s biodiversity; of the 7,000 taxonomists globally, the megadiverse countries have only about 500 in residence (Sarukhán and Dirzo 2001). In addition, the application of those limited resources may not be most effectively directed to supporting and influencing conservation management (Sheil 2001). It is hoped that as economies grow and stabilize, national investment in plant and habitat conservation will also grow, but it is also expected that as economies grow, so will the number of threatened habitats and species (Naidoo and Adamowicz 2001). A review of temperate and tropical hotspots illustrates the challenges and opportunities for ex situ conservation in supporting wild plant diversity. The biodiversity hotspots (sensu Myers et al. 2000) are characterized by high numbers of species, high concentrations of endemics, diverse habitats (often including tropical forest ecosystems), and a high degree of cultural diversity. Such tropical and Mediterranean floras are characterized by large numbers of point endemics or microareal species. The high levels of botanical diversity are threatened by increasing levels of habitat destruction and overexploitation. Gentry (1986) documented how about 38 plant species endemic to a single 20-km2 area of Andean foothill forest in Ecuador were apparently lost through habitat conversion. A large proportion of the surviving biodiversity is held under communal ownership; for instance, nearly 75 percent of the areas listed in Mexico as conservation priorities are under communal ownership or management (Sarukhán and Dirzo 2001). Although such areas may have been under traditional, often sustainable land use for generations, with increased population pressures they are now vulnerable to degradation by a dynamic characterized as the tragedy of the commons, whereby cooperative management of the whole is replaced by individual exploitation (Hardin 1968). The hotspots (sensu Myers et al. 2000) cover a diverse array of situations with varying levels of loss and plant conservation infrastructure. For instance, the five mediterranean hotspots (central Chile, the Cape Floristic Region, the Mediterranean Basin, California, and Southwest Australia) are characterized by both high levels of species diversity (e.g., 25,000 species for the Mediterranean Basin) and high levels of endemism (exemplified by the Cape Floristic Region, with 5,600 endemic species). The main threats to these hotspots include habitat loss, fragmentation, invasives, and changes to fire ecology. Most of the indigenous flora for these hotspots exhibits orthodox seed behavior, allowing seed banking to proceed as a long-term conservation backup. Only a few governments in these
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hotspot areas have produced thorough and updated conservation assessments (e.g., South Africa, California, and Southwest Australia). Each hotspot has differing infrastructures for ex situ conservation. For instance, the California Floristic Province is found in an affluent developed nation, whereas the Mediterranean Basin encompasses both developed (e.g., France and Spain) and less developed (e.g., Morocco and Albania) nations. The Californian and Cape hotspots have an established resource of botanic gardens and ex situ facilities, including some focusing on regional issues. In California, for example, the Santa Barbara Botanic Garden, the University of California at Berkeley Botanical Garden, and the Rancho Santa Ana Botanic Garden, all members of the CPC, have significant programs dedicated to threatened native plants, as does the National Botanical Institute in South Africa. In contrast, the Mediterranean Basin has a shortage of ex situ facilities. A number of nations lack an adequate infrastructure for ex situ wild plant conservation (e.g., Egypt, Albania, Algeria, and Tunisia). Others (e.g., Morocco, Turkey, Greece, and Lebanon) are developing an infrastructure. Only three, Spain, France, and Italy, have advanced and effective infrastructures. Regional networking is variable, with effective networking in California, Southwest Australia, and southern Africa and a nascent network developing in the Mediterranean Basin. Southwest Australia exemplifies a coordinated regional approach that includes substantial investments in seed banking (Chapter 3, this volume). Their task is made somewhat easier by being in a single administrative region with a state government committed to integrated plant conservation. This area supports 5,710 plant species with 4,524 endemics (Paczkowska and Chapman 2000). The area is threatened by habitat loss caused by agriculture, the long-term effects of fragmentation and salinization, and the developing impact of introduced pests and disease, most notably the fungal pathogen Phytopthora cinnamomi. The flora of the area is well known, and studies of the flora by regional conservation agencies (e.g., Conservation and Land Management) and conservation facilities (Kings Park and Botanic Garden [KPBG]) are ongoing. However, the pace of additions to the regional list of threatened taxa far exceeds the physical capacity of the two key facilities associated with ex situ management of conservation taxa (KPBG through the Science Unit and Seed Technology Centre; Threatened Flora Seed Centre). Research at KPBG has highlighted the need for a fully integrated and pragmatic program of research and development. Strategic breakthroughs include new molecular tools
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for provenance evaluation (Krauss 2000); protocols for tissue culture (Rossetto and Dixon 1992); cryostorage techniques for plant tissue, symbiotic organisms (e.g., mycorrhiza), and seed (Touchell and Dixon 1994; Batty et al. 2001; Turner et al. 2001); and tools for ecosystem restoration that facilitate the reintroduction of threatened taxa (Batty et al. 2002). In most tropical regions the disparity between need and existing facilities is even more extreme. These areas support huge numbers of endemic plants, often with poorly developed ex situ resources. Even in developed nations there is a disparity between resources and needs. For instance, Hawaii as part of the Polynesia and Micronesia hotspot has by far the largest number of threatened plants found in any state of the United States (Dobson et al. 1997). The collapse of many wild populations and the impact of invasives on natural habitats suggest that a large proportion of Hawaii’s endemic plant diversity is unlikely to survive as wild populations beyond the near future (Maunder et al. 2002a). The Hawaiian botanic gardens are in emergency mode, salvaging the last samples of threatened lineages. Fortunately, most of the endemic flora exhibits orthodox or at least nonrecalcitrant seed behavior (A. Yoshinaga, pers. comm., 2002), allowing the use of seed banking. The islands’ five main ex situ facilities, all members of CPC, are largely on lowland sites, whereas the majority of threatened taxa are upland species. The ex situ facilities have adopted the following immediate responses (see Chapter 10, this volume): salvaging of threatened taxa through seed and tissue banking; restoration of wild populations to areas of surviving habitat, including active pest management and weeding of fenced habitat plots; demonstration activities promoting habitat restoration and species recovery; and active promotion of Hawaiian conservation issues to the visiting public. In Hawaii, ex situ conservation has been adopted as an essential working tool by a variety of national (U.S. Fish and Wildlife Service [USFWS] and U.S. Army), state (Department of Land and Natural Resources), and local bodies (community restoration projects) (Bruegmann et al. 2002). The Eastern Arc hotspot of Kenya and Tanzania faces a biodiversity crisis resulting from commercial logging, expansion of plantation and village agriculture, and largely unregulated coastal tourism development. Although it shares many features with the Hawaiian crisis in terms of the scale of biodiversity loss, this region is poorly resourced for ex situ conservation. The three regional botanic gardens are developing local and national conservation initiatives (Amani, Tanzania; National Museums of
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Kenya; and the Nairobi Arboretum) with smaller ex situ facilities associated with protected forest areas in Kenya (Maunder et al. 2002b). There are excellent seed bank facilities in the region, but they focus largely on agricultural and forestry taxa. Although a proportion of the flora exhibits orthodox seed behavior, the wet forest flora of the region is largely unassessed in terms of the storage behavior of seeds. We recognize that the majority of the area’s botanical diversity would be best preserved in protected areas, if available. New ex situ facilities associated with protected areas, forest reserves, and community projects can potentially support both species recovery and habitat restoration. Conservation and restoration activities should focus on supporting local livelihoods and plant conservation and perpetuating traditional knowledge in plant use and habitat management (Maunder et al. 2002b; Box 18.2). In many biodiversity-rich developing countries, candidates for ex situ conservation can and should be identified in conjunction with the designation of protected areas. Once candidate taxa have been identified for ex situ management, the realities of tropical horticultural practice must be realized. In general, horticulture as a scientifically based profession is poorly developed in the tropics. Improving horticultural capacity, in terms of both functioning facilities and trained professional staff, will take high levels of investment. For a number of plant groups (notably cacti and succulents, bromeliads, orchids, palms, cycads, and geophytes) there already exists a body of horticultural experience often linked to the horticultural appeal of the group. For example, orchid conservation has both suffered at the hands of irresponsible collectors and benefited from the horticultural expertise developed by amateur and professional growers. However, the broader range of candidate taxa—many rarely, if ever, cultivated before— will require a significant investment in horticultural research, professional training, and institutional capacity (Dixon 1987). For species with orthodox seed behavior, the opportunities for effective, genetically representative ex situ conservation can be best realized through seed storage. However, the vegetation in many of the world’s hotspots is characterized by high levels of recalcitrant seed behavior that precludes using conventional seed bank techniques. Conventional wisdom has held that species from tropical wet forest habitats have predominantly recalcitrant seed, whereas those from arid and mediterranean regions for the most part have orthodox seed behavior and can be stored in conventional seed banks. Recent research indicates that the uncritical use of conventional
box 18.2 The Role of the Ex Situ Conservation Facility in Supporting Wild Plant Diversity Goal • The conservation of wild plant diversity in both protected areas and
human-modified landscapes. Objectives • Preserving local and regional plant diversity, primarily in situ. • Promoting the conservation of large-scale habitat and ecosystem
areas. • Providing ex situ resources for local and regional conservation agen-
cies and local communities in support of species recovery and habitat restoration. • Promoting local and national awareness about threatened species and ecology, protected areas, ecological dynamics, and the dependence of human communities on plant resources. • Forging productive relationships between conservation managers and neighboring constituencies. Requirements • Institutional strategy based on a rigorous and viable business plan
derived from participatory processes and full consultation. • Institutional mission and role linked directly to established national
•
• • • •
and local priorities for conservation, capacity building, and sustainable development. Formal agreements established between local stakeholders, such as communities, financiers, managers, local authorities, and technical advisory groups. Institutions managed on sound scientific principles in pursuit of species and habitat conservation objectives. Managerial culture and display systems sympathetic with local culture and traditions. Facility supported by strategic alliances at the local, national, and international levels. Biodiversity restoration projects designed to benefit local livelihoods and plant conservation and ensure the perpetuation of traditional knowledge.
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Benefits Species Conservation Impact • Plant germplasm effectively stored and made available for species
recovery and habitat restoration, meeting international, national, and local legislative obligations. • Plant germplasm stored to support conservation of local agricultural diversity. • Functional links between the display facility, local protected areas, and surrounding lands, promoting landscape-level natural resource planning and conservation. • Cost-effective ex situ and in-country propagation facilities. Conservation Knowledge and Awareness • Key species displayed to many citizens and overseas visitors. • Displayed species used to explain and promote the conservation of
larger natural areas and wild populations. • Displayed species promoting sustainable land use and harvesting
practices. • Displayed species and landscapes promoting the conservation work of
the institution and collaborators. Habitat and Landscape-Level Conservation • Greater awareness of the values of protected areas and their roles in
landscape-scale resource conservation. • Strong functional links with local community and land managers. • An area-based approach to public interpretation and awareness. • Increased awareness in many groups of the benefits of landscape-level
approach. • Scope for testing, demonstrating, and undertaking habitat restoration. Based on Maunder et al. 2002b.
seed storage for highly endemic floras, whether tropical or temperate, may be problematic. This is because they may harbor taxa with unpredictable seed storage behavior unless significant baseline research is undertaken to ascertain storability criteria (Merritt et al. 2000; Tieu et al. 2001; Chapter 7, this volume). National and regional projects must focus on the basic needs for seed storage of threatened taxa as a key priority in developing
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effective ex situ seed storage programs; a failure to realize the need for research into storage and regeneration criteria could result in expensive gene morgues rather than effective gene banks. Two challenges—one sadly proven and the other predicted—are political instability and global climate change. Large areas of the world are subject to the upheavals of war and political and social instability. Political and economic instability can profoundly disturb ex situ programs through the collapse of government and financial infrastructures, damage to collections and facilities, and even violence toward staff. During the Siege of Leningrad in World War II, a number of staff starved to death while attempting to curate the priceless seed bank accessions in what is now known as the Vavilov Institute (Loskutov 1999). More recently, the seed bank in Kabul, Afghanistan, was looted twice in 10 years (Charles 2002). Increasingly, the long-term stewardship of ex situ resources must take into account the damaging impacts of war and political change on both wild plant resources and ex situ institutions (Richards and Ruivenkamp 1995; Sperling 2001). Similarly, the impact of climate change must be built into ex situ strategy, as we plan to store samples of populations threatened by climate change and potentially translocate threatened populations to new sites (Crumpacker et al. 2001). The Northern Hemisphere model for ex situ conservation, with storage of material in remote facilities, can diminish the link between people and their local plant resources. If ex situ conservation is to be effectively applied in the world’s hotspots, facilities need to work with and be directly supported by local communities (Burgess 1994; Muller 1994). This entails a focus on the relationship between local people and plant resources, with an emphasis on managing plant resources of valued utility, such as timber resources, medicinals, and traded ornamentals (Vovides et al. 2002; Jaenicke et al. 2002; Sunderland et al. 2002; O’Neill et al. 2001). Species loss in biodiversity hotspots is not limited to developing countries. The California Floristic Province, Hawaii, and New Caledonia (overseas territory of France) are suffering significant losses. So, too, the biodiversity hotspot of southwest Western Australia (sensu Myers et al. 2000), a region of immense mineral, agricultural, and intellectual wealth and that has an estimated 2,300 species in conservation crisis (Brown et al. 1998). These examples represent a different yet familiar scenario in which species decline is linked to economic drivers overriding ecological and conservation considerations, demonstrating that the extinction of plant species is not exclusive to the developing world.
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The Ex Situ Load: How Many Species Can Ex Situ Facilities Manage? As the ability of habitat areas to support current levels of indigenous biodiversity declines, more species will need some form of ex situ support. It is difficult to gauge how many species in the coming decades will need storage and cultivation in secure and effective facilities. However, we can be reasonably sure that global changes will continue to erode plant diversity at an accelerating rate (Tilman and Lehman 2001) and that present levels of ex situ investment will be grossly insufficient to counteract expected extinction loss. Currently, there are few national lists accurately identifying taxa needing ex situ support, in part because of the current ambiguities surrounding the identification and categorization of threatened species (Gärdenfors 2001) and an insufficient understanding of the available tools for ex situ support. We anticipate that as more countries undertake Red Listing exercises using the World Conservation Union (IUCN) criteria, more accurate lists will be available to guide ex situ practitioners. We base our discussion on the recommendations for ex situ representation outlined in the Global Strategy for Plant Conservation (CBD 2002), namely that at least 60 percent of threatened plant species be maintained in accessible ex situ collections and that 10 percent of threatened plant species be included in recovery and restoration programs. We interpret recovery management as species-specific management to improve or reinstate the viability of wild populations. Oldfield et al. (1998) document more than 7,300 globally threatened tree species, which represents around 10 percent of the world’s trees. If this pattern is representative for the world’s flora (estimated at 300,000 to 420,000 species), then 30,000 to 42,000 plant species may be threatened. Recent studies (e.g., Pitman et al. 2002; Pitman and Jørgensen 2002) suggest that a large proportion of endemic tropical species will be threatened. These studies suggest that almost 30 percent of the world’s flora is threatened with extinction; this equates to about 90,000 to 126,000 taxa. Ensuring a 60 percent representation means that 54,000 to 75,600 species must be maintained in ex situ facilities. Similarly, if 10 percent of threatened plant species are to be included in recovery and restoration programs, then 9,000 to 12,600 species must be targeted. Assuming arbitrarily that one resourced conservationist can manage a maximum of four plant recovery projects, this suggests the global need for about 3,000 plant conservation practitioners.
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We expand this discussion to look at necessary national ex situ infrastructure for megadiverse nations. Because the current assessments of numbers of threatened species are neither sufficiently accurate nor consistent across regions, we use the total number of endemic species as the basis for calculating potential ex situ loads. We also recognize that ex situ storage and management will also serve threatened nonendemics and the nonthreatened taxa needed for habitat restoration activities. If we assume that 60 percent of Brazil’s endemic plant species (16,500–18,500 species, a mean of 17,500) may need to be held ex situ, this could equate to 10,500 species, of which only a small proportion can probably be seed banked; the rest will have to be cultivated as growing plants. If we assume that 10 percent of Brazil’s 17,500 threatened plants are managed for recovery, this equates to 1,750 species. The Center for Plant Conservation (CPC) network (United States) manages an average of 18 threatened plants per facility; we will assume, perhaps optimistically, that 25 taxa per facility can be managed. This provides a figure for Brazil of 70 botanic gardens needed for recovery management, with a dedicated staff of around 435. We have identified five countries (Indonesia, Mexico, Madagascar, Philippines, and Malaysia) that will need a particularly high level of investment for single-species management. These countries have an annual forest loss of more than 4 percent, less than 10 percent of land area allocated to protected areas, and more than 40 percent endemism. Assuming that 10 percent of threatened plants are managed for recovery, it is evident that Indonesia and Madagascar will need new facilities and a huge increase in effort (Table 18.1). However, assuming, perhaps optimistically, that existing facilities can support an increase in activities, China, Mexico, and India have an apparent surplus capacity. However, it is likely that these existing facilities will need more resources and staff capacity. In addition, it is likely that these facilities will not be located in the regions where they are needed. For instance, of the 32 Mexican states, 10 do not have botanic gardens; however, four of these states have facilities planned or in preparation (Rodríguez-Acosta 2000). Focusing attention on particular nations in this way may mean that significant concentrations of biodiversity will not receive the necessary investment in ex situ infrastructure. Figure 18.1 illustrates the relationship between the numbers of botanic gardens as a function of the number of endemic plant taxa for the various biodiversity hotspots identified by Myers et al. (2000). If sufficient investment were made in four of the five coun-
table 18.1 Ex situ conservation load for selected high-endemism nations. Country
Plant Endemism
Brazil Indonesia Colombia Mexico Australia Madagascar China Philippines India Peru Papua New Guinea Ecuador United States Malaysia Republic of South Africa Democratic Republic of the Congo
16,500–18,500 14,800–18,500 15,000–17,000 10,000–15,000 14,500 8,800–9,600 10,000 3,800–6,000 7,000–7,900 5,400 10,500–16,000 4,000–5,000 4,000 6,500–8,000 16,500 3,200
Mean
60%: Ex Situ Storage
10%: Recovery Management
No. of Existing Botanic Gardens
Required Botanic Gardens (25 species per facility)
17,500 16,650 16,000 12,500 14,500 9,200 10,000 4,900 7,450 5,400 13,250 4,500 4,000 7,250 16,500 3,200
10,500 9,990 9,600 7,500 8,700 5,520 6,000 2,940 4,470 3,240 7,950 2,700 2,400 4,350 9,900 1,920
1,750 1,665 1,600 1,250 1,450 920 1,000 490 745 540 1,325 450 400 725 1,650 320
29 5 25 89 128 2 106 10 122 7 4 6 296 10 19 0
70 67 64 50 58 37 40 20 30 22 53 18 16 29 66 13
Source: Plant endemism data from Davis et al. (1994, 1996, and 1997). Botanic garden numbers from Wyse Jackson 2001.
Figure 18.1 Graphic representation of the relationship between ex situ capacity (i.e., botanic gardens) and need (number of endemic plants) in the 25 global biodiversity hotspots identified by Myers et al. (2000). Botanic garden numbers are from Wyse Jackson (2001). Diagonal lines indicate constant values, or isoclines of the relationship between capacity and need. Capacity is greatest in the upper left (fewest endemics per botanic garden) and diminishes toward the lower right. Note that scales are logarithmic. Hotspot symbols, names, and in parentheses, number of endemics per botanic garden (in or within 150 miles of hotspot) are BAF, Brazil’s Atlantic Forest (800); BC, Brazil’s Cerrado (880); Cal, California Floristic Province (49); Cap, Cape Floristic Region (1136); Car, Caribbean (250); Cau, Caucasus (133); CC, Central Chile (321); CDE, Choco, Darien, and western Ecuador (107); EAM, Eastern Arc Mountains (136); IB, Indo-Burma (318); MA, Mesoamerica (161); Mad, Madagascar (4,852); MB, Mediterranean Basin (146); MSWC, Mountains of southwestern China (700); NC, New Caledonia (2,551 endemics, 0 botanic gardens); NZ, New Zealand (93); P, Philippines (583); PM, Polynesia and Micronesia (278); SK, Succulent Karoo (647); Sun, Sundaland (1,000); SWA, Southwest Australia (866); TA, Tropical Andes (444); WAF, Western African Forests (83); Wal, Wallacea (1,500); WGS, Western Ghats and Sri Lanka (136).
tries listed—Indonesia, Madagascar, Philippines, and Malaysia—they would presumably be able to take care of the ex situ needs in three of the four most underresourced biodiversity hotspots (Philippines, Sundaland, Wallacea, and Madagascar). Unfortunately, a nation-based system of evaluating need for investment would miss the most underresourced hotspot, New Caledonia, which has more than 2,500 endemic taxa and no botanic gardens. Other hotspots in critical need of significant international investment in ex situ infrastructure—those with more than 400 endemic taxa per botanic garden—include the Cape Floristic Region (1,136), Brazil’s Cerrado (880), Brazil’s Atlantic Forest (800), the mountains of southwestern China (700), Southwest Australia (866; but see Chapter 3, this volume), Succulent Karoo (647), and Tropical Andes (444). Modest but carefully targeted investment by the international community in ex situ facilities in hotspot areas could go a long way toward slowing the rate of species loss and result in substantial benefits to humanity in the decades and centuries to come.
Practical Recommendations If ex situ conservation is to play an effective role in conserving wild plant diversity as a support to habitat conservation (Box 18.2), a number of actions are needed to ensure that appropriate levels of infrastructure and capacity are established. We foresee the development of ex situ facilities and services as part of community-based conservation initiatives (sensu Getz et al. 1999), and as playing a fundamental role in supporting habitat and landscape-scale conservation (Box 18.2; introduction and Chapter 5, this volume). Furthermore, the development of national guidelines for seed storage and ex situ mechanisms for conservation can provide a basis for standardization of approaches that will ensure adequate genetic representation and maximum longevity of the ex situ collections. For example, the Australian Network for Plant Conservation has released national guidelines for germplasm storage and translocation activities (ANPC 1997; Touchell et al. 1997; Brown et al. 2003). These guidelines have had wide audience appeal across a broad section of stakeholders including farmers, nongovernment organizations, and conservation agencies. The development of similar guidelines for key biodiverse countries and regions would provide broad benefits for stimulating best practice in conservation activities.
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Ex Situ Policy The Convention on Biological Diversity (CBD) and national laws on access to and ownership of plant genetic resources have introduced legal and ethical obligations for ex situ conservation facilities (Glowka et al. 1994). The CBD recognizes the value of ex situ conservation (Article 9) with an emphasis on undertaking these activities “preferably in the country of origin” and as a support to the “recovery and rehabilitation of threatened species and for their reintroduction into their natural habitats” (Glowka et al. 1994: 52–53). The IUCN Species Survival Commission policy on ex situ conservation (IUCN 2002: 1) recognizes that “the primary objective of maintaining ex situ populations is to help support the conservation of a threatened taxon, its genetic diversity, and its habitat. Ex situ programmes should give added value to other complementary programmes for conservation.” The International Agenda for Botanic Gardens in Conservation (BGCI 2001) defines the global mission of botanic gardens worldwide in conservation as follows: • Stem the loss of plant species and their genetic diversity worldwide. • Focus on preventing further degradation of the world’s natural
environment. • Raise public understanding of the value of plant diversity and the
threats it faces. • Implement practical action for the benefit and improvement of the
world’s natural environment. • Promote and ensure the sustainable use of the world’s natural
resources for present and future generations.
The integral role of ex situ conservation has been emphasized in the Global Strategy for Plant Conservation (CBD 2002). However, the next stage, the implementation of truly integrated plant conservation strategies, is hampered by a shortage of funds and an insufficient understanding of the potential contribution of ex situ conservation.
Improving Professional Capacity Integrated plant conservation, sensu Falk (1990), seeks to combine the protection of plants in their native habitat with an ex situ conservation program
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that will both provide an insurance policy against extinction in the wild and provide tools and resources to better manage the wild populations. We suggest that the definition be expanded to include the integration of theoretical academic approaches with applied practical approaches, the integration of coarse-filter habitat-level conservation with fine-filter singlespecies work, the integration of professional and volunteer efforts, and the integration of a public education component into all conservation programs. Many botanic gardens are embracing this approach with a wider perspective on conservation and outreach. For instance, the Chicago Botanic Garden’s Institute for Plant Conservation in the United States and KPBG in Australia both have comprehensive applied plant conservation research and training programs focused on delivering high-quality science in the context of applied conservation management of plant diversity. This shift is being reflected in institutional management processes, such as the identification of appropriate key performance indicators for ex situ facilities (Bartos and Kelly 1998), and in interdepartmental planning processes that integrate the conservation perspective into institutional and garden activities. To cope with the increased range of threatened taxa needing ex situ management, increased emphasis on applied research to develop horticultural protocols for wild taxa will be needed, particularly in the tropics. Ongoing applied research is moving toward allowing cost-effective storage facilities for seed, tissue, and pollen and, in some facilities, the storage of symbiotic organisms such as mycorrhizal fungi and beneficial bacteria. These are more secure ways to store important accessions compared with using conventional active growth horticultural collections for long-term cultivation. Partnering and twinning activities between economically and technologically more developed regions and developing countries provide a mechanism that is already delivering a wide variety of benefits across a broad range of habitats, species, and regions. For example, the activities of Botanic Gardens Conservation International (BGCI) through development of the Global Strategy for Plant Conservation endorses opportunities for shared activities between botanic gardens. Excellent examples of the benefits of local empowerment include the development of the Limbe Botanic Garden, Cameroon (Sunderland et al. 2002), and the broad international activities of the Millennium Seed Bank project of the Royal Botanic Gardens, Kew (Williams 2001; Smith et al. 2003).
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Species Selection The selection of threatened species for ex situ management will be guided by a variety of criteria including threat level, legislative and institutional responsibilities, likelihood of successful reintroduction, cost-effectiveness, relevant social and economic issues, and curatorial preferences for “pet” taxa. At a crude level, facilities can use the five “E” criteria to help select species priorities: endangerment (selecting the most threatened taxa needing ex situ support), endemism (selecting threatened taxa that represent a unique local or regional responsibility), economic (selecting taxa that provide local or regional economic or social resources, such as medicinal plants), ecological (selecting taxa that have a role in maintaining ecological processes or supporting habitat restoration), and emblematic (selecting threatened taxa that can be used as flagships for promoting landscape- and habitat-level conservation). Another important criterion for selecting taxa for storage is phylogenetic position. If species are chosen for collection on the basis of their phylogenetic position, then the relative threat to future diversity of losing 10 species from a group of closely related species, such as the loss of 10 of the 2,000 or so Astragalus (Fabaceae) species or the 55 Hawaiian Cyanea (Lobeliaceae) species, is very different from losing 10 monotypic genera or families, that is, species with deep phylogenetic roots. Sampling for storage of crop plants is currently better covered than sampling for phylogenetic representation, and both will be important in the distant future. Target species for ex situ management should be identified in advance of catastrophic decline to avoid attempts to salvage from genetically depauperate remnant populations, the so-called living dead, when the chances of successful recovery and reintroduction can be severely reduced (Maunder et al. 1999).
Support to In Situ Management Although in situ efforts are clearly primary, they face some practical challenges that can be tackled using ex situ techniques. For instance, when habitats are affected by events that eradicate local populations, ex situ facilities can provide both landscape species for habitat restoration and threatened taxa for reintroductions and, often, the skills and resources to support effective reestablishment of species. Habitat management may be
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influenced by political and economic priorities or the needs of high-profile taxa. This may damage other components of the habitats, such as the impact of increasing herbivore populations in small forest reserves. Accordingly, there are increasing opportunities to establish ex situ techniques as part of habitat and ecosystem management systems. This will mean developing effective collaborations and adding ex situ techniques to the portfolio of techniques used by the habitat manager, such as seed banking and reintroduction. Wherever possible, conservation efforts for individual species should be integrated with regional conservation packages for both ecosystems and suites of species. This approach provides a cohesive and identifiable product that is readily marketed and establishes conservation tools and approaches across a wide range of stakeholders. As habitat areas decline in both quality and area and the species load for ex situ management increases, ex situ facilities will increasingly need to coordinate ex situ custody with habitat restoration. This will require tools for the cost-effective restoration of extensive landscapes, tools that effectively promote natural regeneration and serve the goals of biodiversity restoration while ideally providing some economic benefit to resident communities. There is an urgent need for habitat-specific restoration techniques in the tropical world. Where such manuals or protocols exist, the impact has been significant; for instance, a manual on prairie restoration for the Chicago region (Packard and Mutel 1996) has resulted in many successful restoration initiatives. The Australian Network for Plant Conservation has produced benchmark guidelines to assist in a more cohesive approach to the key areas of translocation and germplasm storage of threatened species. These guidelines have assisted in the development of a national approach to the recovery of threatened species and ecosystems (ANPC 1997; Touchell et al. 1997; Brown et al. 2003). An example of a successful regional initiative is the Chicago Wilderness Project, encompassing botanic gardens, zoos, museums, protected areas, and community groups in an effort to secure, manage, and conserve native habitats in the greater Chicago region (Chicago Region Biodiversity Council 1999).
Fundraising and Outreach We propose that ex situ plant conservation facilities, such as botanic gardens, can and must play an increasingly important role in generating funds
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and support for in situ conservation. Out-of-country ex situ facilities, serving an affluent and interested visiting public, can increasingly provide scientific, financial, and managerial support to overseas facilities where indigenous species can be maintained in a more cost-effective manner (Chapter 5, this volume). In addition to acting as centers for biodiversity conservation, there is a need for botanic gardens to engage their stakeholders and to both intellectually and emotionally challenge them to modify behavior and values. The opportunity exists to use investment in ex situ display facilities to raise public concern and finances for supporting in situ biodiversity projects. For instance, the Utrecht Botanical Gardens, Netherlands, uses its rainforest exhibit as the base for fundraising activities to purchase rainforest reserves in French Guyana. Opportunities exist for crossborder and international joint projects, such as the link between British-based institutes (Durrell Wildlife Conservation Trust, Royal Botanic Gardens, Kew, and Fauna and Flora International) and Mascarene conservation initiatives and the role of the Sonora Desert Museum and Native Seeds/SEARCH in promoting cooperative projects with neighbors in Mexico. In a world of increasing political and social complexity, botanic gardens are increasingly recognized as politically neutral and noncontroversial agents for conservation. This special position provides botanic gardens with a unique edge in garnering financial and political support for conservation activities.
Partnerships Given their inherent supporting role, ex situ facilities will be most effective in conserving plant diversity when working in active partnership with both land managers and regulatory agencies. Indeed, ex situ facilities need to actively integrate themselves into local, national, and international conservation infrastructures and, where appropriate, foster links with in situ owners and managers. In practice, the actual arrangements will have to be tailored to the context in which an ex situ provider operates (Guerrant and Raven 2003). For example, the CPC was built on an original vision of a strategic alliance among independent organizations, each of which would maintain ex situ samples of their region’s rarest plants. The national scope and credibility of the CPC have enabled it to enter into a series of important memoranda of understanding (MOUs) with several agencies of the U.S.
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federal government, such as the U.S. Department of the Interior (Bureau of Land Management [BLM] and USFWS) and the U.S. Department of Agriculture (U.S. Forest Service [USFS]). These greatly facilitate the work of individual CPC participating institutions (PIs) in their work with the regional and local offices of those same agencies. In turn, the work of the PIs at the local and regional levels, using a common set of standards and guidelines, greatly facilitated the ability of the CPC to enter into the national agreements. At the core of the CPC’s tangible ex situ conservation resources is the National Collection of Endangered Plants, which currently numbers almost 600 of the most threatened U.S. plant taxa. Among the greatest benefits of any partnership are those that emerge from the interaction between otherwise independent people and organizations, each with their own experiences, strengths, limitations, and points of view. One particularly important benefit of a partnership organization such as the CPC is that the original institutions invited to participate could be strategically chosen, partly on the basis of their geographic location. It is no accident that California, Florida, Texas, and Hawaii each have multiple PIs. Over time the diverse experiences and expertise of the PIs have become reflected in an evolving set of standards and guidelines. The national CPC office provided the original model and directions for each of the PIs to apply in their region, which ensured a level of consistency and quality control across the nation. Over time the growing bodies of experience among the PIs have been shared within the CPC, and the recommended standards and guidelines have been modified to reflect their experience. The series of books and their recommendations (Falk and Holsinger 1991; Falk et al. 1996; this volume) that have emerged from the CPC are tangible conservation resources. The Berry Botanic Garden (BBG), a charter member of the CPC, exemplifies the benefits derived from being part of the CPC partnership. Located in the U.S. Pacific Northwest, the BBG is responsible for more than 60 taxa in the CPC National Collection (Guerrant 1997). Almost all have orthodox seed, so they can be stored frozen in the Seed Bank for Rare and Endangered Plants of the Pacific Northwest. The seed bank staff have established a close cooperative relationship with the major land managers in the region: the BLM and the USFS, which jointly control more than half the land in the state of Oregon, and the USFWS, which oversees the federal Endangered Species Act. The majority of the more than 10,000
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accessions of more than 300 taxa were collected by BLM or USFS personnel. As in all good relationships, both partners benefit: the Seed Bank becomes larger and more representative, often with associated funding, and the agencies are able to reduce the risk of losing sensitive taxa in their care and have the biological resources for reintroduction if necessary. The state office obtains annual reports for taxa on their sensitive species lists outlining the status of ex situ collections. The partnerships between the BBG and relevant land management and regulatory agencies have produced a regional system whereby ex situ collections can be accumulated in a systematic and prioritized manner that can strategically target the region’s most threatened taxa and populations.
Managing Ex Situ Liabilities Ex situ facilities have a legal and moral responsibility to manage the inherent risks associated with living plant collections. Collection managers should increasingly assess their collections to identify high-risk taxa and take basic measures to reduce the risk of escape (Wittenberg and Cock 2001; Baskin 2002). Botanic gardens, and ex situ collections in general, have given insufficient attention to pest management (Clement 1999). Ex situ collections have unwittingly acted as vectors for introductions of a number of pathogens and pests, such as the distribution of fungus Armillaria to South Africa via a botanic garden (Coetzee et al. 2001) and the distribution of New Zealand flatworm via botanic gardens and nurseries (Boag et al. 1994).
Institutional Continuity The commitment to ex situ conservation entails long-term investment in both facilities and staff. This could be measured in decades and possibly centuries for some taxa. For example, Franklinia has been in cultivation since extinction in the wild around 1790; similarly, plants of Erica verticillata, extirpated in South Africa, have been repatriated from Austria, derived from eighteenth-century collections (A. Hitchcock, pers. comm., 2002). As the importance of ex situ collections increases relative to surviving wild populations, there is a concomitant need to improve the continuity of management and ensure the long-term commitment of institutions and individuals to species or habitat management projects. This
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includes a number of biological, financial, and managerial challenges. By adopting an ex situ responsibility, institutions are committing themselves to a service that lasts beyond traditional contract periods and in some cases represents an intergenerational responsibility that must persist through institutional changes in directorship and strategy. The use of long-term storage techniques, such as seed banks, will greatly facilitate collection longevity. However, for growing collections there is an urgent need to develop protocols to ensure long-term genetic, horticultural, phytosanitary, and data security.
Contribution to Sustainable Use Ex situ facilities have an opportunity, perhaps even an obligation, to align themselves with the global imperative of supporting sustainable livelihoods and alleviating poverty (Box 18.2). Wild plants are essential to many rural communities in developing countries, supplying building, furniture, and craft materials, firewood, livestock forage, aromatics, and medicinals (Campbell and Luckert 2002). It is estimated that 80 percent of people in developing countries rely chiefly on traditional, plant-based medicine for their primary healthcare (Farnsworth and Soejarto 1991), and the use of wild-collected herbal medicine in developed countries is growing (Lange 1997). Indeed, the trade and associated commerce in wild-collected plant products can form a significant part of the income of rural people (Olsen 1998). An increasing number of ex situ facilities, including botanic gardens, are developing applied conservation programs for medicinal and agricultural plants (Introduction, this volume). The Native Seeds/SEARCH, based in Arizona, and the Foundation for the Revitalization of Local Health Traditions (FRLHT), based in Bangalore, India, demonstrate the potential for a multidisciplinary approach. Native Seeds/SEARCH has a seed bank for traditional cultivars, projects promoting traditional desert foods to combat diabetes, a culture memory bank, a conservation farm, and a wild Capsicum (Solanaceae) conservation area jointly managed with the Coronado National Forest. Similarly, FRLHT supports a network of 17 ex situ facilities and 50 in situ sites as part of the Medicinal Plant Conservation Network. Ex situ institutions can align their strategic objectives for plant conservation with recognized international measures for development, thereby opening the door to increased funding opportunities and improved politi-
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cal status. Established measures for development, well-being, and environmental sustainability include Agenda 21, the Human Development Index of the United Nations Development Programme (UNDP), the World Summit on Sustainable Development Plan of Implementation, the United Nations Millenium Development Goals, the Human Wellbeing Index proposed by a consortium including the Food and Agricultural Organisation of the United Nations, and the International Development Targets (IDT), set by the Organisation for Economic Co-operation and Development (DFID 2001; Prescott-Allen 2001).
Conclusions It is likely that existing protected area networks in regions of high plant diversity will be insufficient to ensure the long-term survival of all plant diversity; accordingly, there will be a continued need for ex situ activities to directly support in situ conservation. We argue that conservation facilities and institutions can provide quantifiable evidence of the value of their work and resources in terms of species and populations recovered, areas and resources restored, education (outreach to public and stakeholders and beneficial change implemented), and services to the local community (local employment and revenues used for local resources; Box 18.2). These activities depend on long-term funding, appropriate economic analysis, and long-term planning to ensure institutional and project viability. A relatively small financial investment, in global development terms, could build a series of ex situ facilities such as seed banks and train locals in their operation and in horticultural skills. The places in most need on a global scale, and the units in which success can be tracked, have been described. The potential returns on investment are enormous. What we do or do not do now will reverberate for centuries to come. Ex situ facilities, such as botanic gardens, have already shown an ability to manage plant resources, and these facilities are undergoing active evolution as the historical collection priorities of alpha taxonomy are being supplemented and modified by the imperatives of biodiversity conservation and public education. Although the traditional gardens have retained Franklinia and dozens of other plant species now extinct in the wild, the current generation of gardens should not be measured in terms of the living dead but rather in terms of how well they support living populations and landscapes.
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Sperling, L. 2001. The effect of the civil war on Rwanda’s bean seed systems and the unusual bean diversity. Biodiversity and Conservation 10:989–1009. Sunderland, T. C. H., P. C. Blackmore, N. Ndam, and J. Nkefor. 2002. Conservation through cultivation: the work of the Limbe Botanic Garden, Cameroon. Pages 395–419 in M. Maunder, C. Clubbe, C. Hankamer, and M. Groves (eds.), Plant Conservation in the Tropics: Perspectives and Practice. London: Royal Botanic Gardens, Kew. Tieu, A., K. W. Dixon, K. A. Meney, K. Sivasithamparam, and R. L. Barrett. 2001. Spatial and developmental variation in seed dormancy characteristics in the fireresponsive species Anigozanthos manglesii (Haemodoraceae) from Western Australia. Annals of Botany 88:19–26. Tilman, D., and C. L. Lehman. 2001. Human caused environmental change: impacts on plant diversity and evolution. Proceedings of the National Academy of Sciences of the United States of America 98(10):5433–5440. Touchell, D. H., and K. W. Dixon. 1994. Cryopreservation for seedbanking of Australian species. Annals of Botany 74:541–546. Touchell, D. H., M. Richardson, and K. W. Dixon (eds.). 1997. Germplasm Conservation Guidelines for Australia. Canberra: Australian Network for Plant Conservation. Turner, S. R., T. Senaratna, E. Bunn, B. Tan, K. W. Dixon, and D. H. Touchell. 2001. Cryopreservation of shoot tips from six endangered Australian species using a modified vitrification protocol. Annals of Botany 87:371–378. Vovides, A. P., C. G. Iglesias, M. A. Pérez-Farrera, M. Vázquez Torres, and U. Schippmann. 2002. Peasant nurseries: a concept for an integrated conservation strategy for cycads in Mexico. Pages 421–444 in M. Maunder, C. Clubbe, C. Hankamer, and M. Groves (eds.), Plant Conservation in the Tropics: Perspectives and Practice. London: Royal Botanic Gardens, Kew. Whitesell, S., R. J. Lilieholm, and T. L. Sharik. 2002. A global survey of tropical biological field stations. BioScience 52(6):55–64. Williams, N. 2001. Getting precious seed on hold. Current Biology 10(1):42. Wittenberg, R., and M. J. W. Cock (eds.). 2001. Invasive Alien Species: A Toolkit of Best Prevention and Management Practices. Oxford, UK: CAB International. Wyse Jackson, P. 2001. An international review of the ex situ plant collections in botanic gardens of the world. Botanic Gardens Conservation News 3(6):22–33.
appendix 1
Revised Genetic Sampling Guidelines for Conservation Collections of Rare and Endangered Plants Edward O. Guerrant Jr., Peggy L. Fiedler, Kayri Havens, and Mike Maunder
The ultimate goal of collecting and storing seed or other plant material off site (ex situ) is to enhance the long-term survival prospects of sampled populations and species in their native habitats. The solution to this seeming paradox is that off-site samples are a means to an end—namely the continued survival of these threatened species in the wild—and not an end in themselves. Ex situ collections are most valuable when they function as a part of a more comprehensive, integrated conservation strategy to reduce the rate at which plant diversity is being lost (Falk 1987, 1990). Insofar as we are able to maintain off-site samples in a healthy condition, two primary potential benefits emerge. First, ex situ collections reduce the probability that the sampled individuals, populations, and species will become irrecoverable. Second, off-site samples provide material for use in reintroduction, research, or other collection-related activities. Despite the potentially critical conservation value of ex situ samples, such collections have inherent risks and costs. Removing plant material from wild populations potentially reduces, however minimally, the short-term survival prospects of sampled populations. Both the potential benefits and the biological and other resource costs of gathering and maintaining ex situ collections must be included in the calculus of sampling decisions.
History and Direction of These Guidelines The revised guidelines offered here are built on the foundation of the Center for Plant Conservation (CPC) Guidelines for Conservation Collections 419
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of Rare and Endangered Plants (CPC 1991). The CPC guidelines, which have found wide application globally, are organized around a hierarchical series of four practical questions to be answered and finish with a concern for the potential impact collection might have on the sampled population. The questions are: 1. Which species should be collected? 2. How many populations should be sampled per species? 3. How many individuals should be sampled per population? 4. How many propagules should be collected from each individual? 5. Under what circumstances is a multiyear collection plan indicated? The guidelines offered here represent evolutionary, not revolutionary change and attempt to incorporate new knowledge and lessons learned in the last decade. The five questions originally posed remain central to any sampling strategy for conservation collections of threatened plants. Perhaps the most obvious change is a shift in emphasis toward a more explicit evaluation of the multifaceted context in which the basic sampling questions must be addressed. The revised guidelines are organized around the following list of contextual questions to consider. They are intended to assist practitioners in the process of balancing the many factors that must be taken into account in collecting material of threatened plant taxa. what purpose is the material intended to serve? • What purpose(s) is a conservation collection intended to serve? For
example, many fewer propagules are needed to develop germination and cultivation protocols, and the genetic considerations are very different from those for acquiring a genetically representative sample for long-term storage or reintroduction. what material is available? • What is the nature of the sampling universe? Sample sizes appropriate
for a species limited to one or a few small populations are very different from those for a species known from 50 locations, each with a large population. • Is seed storage an option, or must samples be maintained as growing plants? It is generally much easier and more economical to store large numbers of seeds in a seed bank than it is to maintain fewer actively growing plants in a botanic garden or other nonnative setting.
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However, depending on available resources, seed storage may not necessarily be a realistic option, even for taxa with orthodox seeds. what will it take to have enough material for use when needed, and is the benefit worth the cost? • What sources and magnitude of attrition in a collection might be
expected during storage and later use to restore diversity to the wild? Not all propagules collected will survive ex situ storage; indeed, some may be used to monitor their condition during storage, and of those that survive, not all propagules planted out will successfully reproduce. • When is the short-term danger posed by collection high enough to indicate that collection should be spread over 2 or more years? In order for collection to be justified, the expected potential value of the sample must outweigh the short-term impact of collection. Under what circumstances might such restrictions not apply and emergency salvage collection be justified? • If material might be used for reintroduction, maintain seed from each maternal parent plant separate from seeds from other plants. Unless seed from each maternal parent is kept separate during collection, it will be impossible to equalize the contribution of different parent plants in a reintroduction attempt. We begin by reviewing the CPC guidelines (Table A1.1) and then consider in more detail the additional questions just posed. Unlike that of the original guidelines, the order of these additional questions is somewhat arbitrary. The task of determining sample size is made more difficult because these questions are interrelated, and the answer to one may affect others. The final sample size decisions are then the result of an iterative process that ends only when we are satisfied that the probability of increasing the longterm survival prospects sufficiently outweighs the short-term impact to the sampled populations. Our goal is to give conservationists a logical framework with which to decide for themselves what sample sizes are appropriate in ex situ collections by providing a series of factors for them to consider in arriving at their decisions (Table A1.1). Ultimately, an appropriate sample size is highly context dependent and must reflect simultaneously the intended purposes for which a collection is being made, the number and sizes of extant populations, and the likelihood of being able to store them off site in good condition for as long as they are needed.
table a1.1 Summary of Center for Plant Conservation (CPC 1991) genetic sampling guidelines, to which one additional question has been added (“which community or habitat?”), along with alternative benchmark values recommended by Brown and Marshall (1995). Questions/ Decisions
Which community or habitat?
Which species?
How many and which populations per species?
How many and which individuals per population?
How many propagules per individual? 1–20 Survivability of propagules
CPC recommended ranges and factors to consider
n/a
Degree of endangerment
1–5 Degree of gene flow among populations
Target level of biological organization Key considerations
Community
Species
Ecotype, population
1–50 Diversity among individuals within each population Individual
Selection of pioneer and colonizing species to reinstate original taxonomic composition and ecological processes n/a
Potential loss of unique genetic lineage, or gene pool
Degree of genetic difference among populations, population history
Genetic communication within population, law of diminishing returns on additional samples
Survivability of propagules, long-term use of collection
—
50
50
50
Brown and Marshall (1995) benchmark guidelines
Allele
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The CPC Guidelines: An Overview The Center for Plant Conservation’s Genetic Sampling Guidelines for Conservation Collections of Endangered Plants (CPC 1991) represent the first comprehensive attempt to create general guidelines for conservation collections of threatened plants (Table A1.1). The Australian Network for Plant Conservation adapted the CPC guidelines for its own needs (Touchell et al. 1997) as part of a larger, more comprehensive statement about what is involved in germplasm conservation. The original recommended ranges for the size of a conservation collection were designed to aid practitioners in determining how many propagules would be needed to capture a genetically representative sample of the population in question. Reflecting the state of the art at the time, the primary purpose of most collections of rare and endangered plant germplasm was amorphous and open-ended. For example, conservation collections would be “established for the purpose of contributing to the survival and recovery of a species” (CPC 1991: 225). The experience of the last decade has shown us that the task of using ex situ samples to contribute to the survival and recovery of populations and species is both more complicated and more difficult than originally conceived (see Falk et al. 1996). In addition to external factors, such as the marked increases in what constitutes a minimum viable population, the original guidelines did not sufficiently reflect various propagule costs, which must be accounted for in advance of use. Therefore, the figures offered in the original CPC guidelines might better be viewed as minimum estimates of what should survive in a reintroduced population after some of these additional factors are taken into consideration. Though groundbreaking, the CPC guidelines did not emerge from nothing. Rather, they drew on a variety of sources, most notably a foundational series of discussions on germplasm sampling strategies offered by A. H. D. Brown and colleagues (Marshall and Brown 1975, 1983; Brown and Briggs 1991). After the publication of the CPC guidelines, Brown and Marshall (1995) offered targets for germplasm collections in a larger and more comprehensive volume devoted to collecting plant genetic diversity (Guarino et al. 1995). Marshall and Brown (1975) suggested that the objective of a conservation collection for a genetically representative sample should be to include at least one copy of 95 percent of all alleles that occur in a (large) population at frequencies greater than 0.05 (5 percent). Brown and Marshall (1995) note that either increasing the certainty level above 95 percent or dropping the critical allele frequency below 0.05 drastically increases sam-
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ple size with only marginal gains in total genetic representation. They note that a random sample of 59 unrelated gametes from a population is sufficient to achieve this objective. Their benchmark criterion of 50 individuals per population reflects this analysis. They discuss a series of factors that might increase or decrease the sample size, depending on circumstances. In summary, although the general framework of the original CPC guidelines proved very useful in the 1990s, we perceive there to have been a growing feeling in the ex situ conservation community that the recommended ranges for collection may seriously underestimate what is needed to ensure a representative, long-term sample (Guerrant and Pavlik 1998), at least for some purposes. This concern is fueled in part because estimates of minimum viable population size have increased dramatically over the last decade (Lande 1995; Lynch et al. 1995). Thus, in addition to an increased appreciation of the challenges associated with preserving representative germplasm for use in reintroduction, we are also gaining a better understanding of the numbers needed to make a population viable and how to undertake a successful reintroduction effort.
Arriving at an Appropriate Sample Size A complex network of interconnected factors must be considered in the process of arriving at an appropriate sample size for a conservation collection of an endangered plant species. One way to organize the network is to view it as a two-step process driven by two independent classes of factors, both of which feed into an evaluation cycle (Figure A1.1). The major classes of input factors are the taxon being considered and the purposes for which samples are to be used. The choice of a taxon determines the sampling universe (i.e., how many populations are known and how large are they?) and also strongly influences the type of propagules that can be used. The other major driver concerns the various purposes that an ex situ collection is intended to serve. Once the taxon and purposes have been chosen, initial sample size estimates can be made. However, not all propagules collected can reasonably be expected to survive in good condition during the period of time between collection and use. Therefore, sufficient additional propagules are needed to mitigate expected attrition, and revised estimates must be made. The revised sample size estimates, taking attrition into consideration, are then evaluated for their potential impact on the sampled population. If the estimated impact is judged too great, then this additional
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Figure A1.1 Conceptual flow chart illustrating how collection size decisions might be made. Illustrated are two major input factors: the choice of taxa with which to work and the purposes that collections are intended to serve. The information about taxa and purposes together are fed into an evaluation cycle that considers attrition to collections and the potential impact on sampled populations. If the impact is judged to be too great, then the evaluation cycle is repeated until the impact is judged acceptable. Refer to accompanying worksheets (Figures A1.2–A1.4).
factor is added to the sum of inputs, opportunities, and constraints, and the process of evaluating needs and impact is repeated. Only when the perceived benefit of collection is judged to be sufficiently high and the impact on the sampled population sufficiently low is a final sample size determined.
Inputs Two main categories of factors drive the process of determining a proper sampling design: the choice of taxon with which to work and the purposes
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collections are intended to serve (Figure A1.1). Each of these two primary drivers has associated opportunities and constraints.
Choice of Taxon In addition to the intended purpose for ex situ collections, the other major categorical input to sample size decisions is the choice of taxon. The CPC guidelines focus attention on degree of threat and the potential for loss of unique gene pools as primary determinants of which taxa are chosen for ex situ treatment, and those recommendations are still valid. In practical terms, national or more narrowly focused lists of threatened plants can be used to help prioritize taxa for collection. The choice of a taxon establishes two sets of opportunities and constraints. One is the sampling universe: how many populations of that taxon are known, and how large are they? The other concerns our ability to store and cultivate the taxon: is seed storage an option, and, if so, how well and economically can it be stored in good condition for long periods of time, or must vegetative material be used? If vegetative material must be used, are adequate resources and personnel available to care for the material?
Sampling Universe: Number and Size of Populations It is one thing to have an ideal target range for the number of propagules we would like to have available for use, but the actual optimal number to be collected is contingent on many factors, the foremost of which is the nature of the sampling universe. The number, size, dynamics, and threat status of known populations circumscribe the range of what can be collected. Much of the sampling theory on which these guidelines are built was developed to inform collections of sample sizes necessary to obtain some specified likelihood of capturing alleles occurring in some specified minimum frequency, from an unspecified but large population. As noted in the original guidelines, many threatened species are found in few, often very small populations. As grim as many gross population size figures are, often they may seriously overstate the number of individuals from which seeds can be gathered. Seeds can be gathered only from successfully reproducing plants, and not all plants in a population are reproductive. Recommendation: For species with 50 or fewer populations, collect from as many populations as resources allow, up to all 50. For species
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with more than 50 populations, collect from as many populations as is practical, up to 50. For populations with 50 or fewer individuals, collect from all known individuals; for populations with more than 50 individuals, collect from 50. These recommendations may appear to be absurdly high, and in many ways they are. We intend for them to represent the ideal sample meant to serve the broad range of expected purposes. These figures also assume a very large sampling universe, which is often not the case with threatened species. In practice, therefore, sample sizes will almost always be much smaller than these benchmark guidelines, reflecting the context in which particular taxa are found, our ability to work with them, and our purposes for collecting samples. Before extensive collections of any one taxon are made, however, many factors must be considered, including the needs of other species. It is not necessarily desirable to finish collecting from one taxon before beginning on another. Rather, it may be better to spread collection resources over many populations of many taxa over an extended period of time. In so doing, limited resources can be allocated in a prioritized manner, focusing on the most threatened populations and species first. After initial collections of as many of the most threatened populations and taxa have been made, subsequent efforts can be made to make the collections larger and more representative. This has the added benefit of minimizing collection pressure on individual populations.
Storage Options: Seeds or Growing Plants? Perhaps the most basic consideration in deciding how many propagules to collect concerns how effectively and economically a particular species can be maintained over time. Our ability to maintain plant material off site in a healthy condition therefore dramatically influences the collection targets. If we cannot reliably maintain material off site in good condition for long periods of time, the biological cost to sampled populations may be very high relative to the conservation value of any collection made. Therefore, a key factor in determining how much to collect concerns our ability to maintain the plant material off site. Collection of seed for off-site storage is generally preferable to gathering plant parts or whole plants for off-site storage. There are two main arguments supporting this contention. First, seed removal is considered by many, but
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not all (K. W. Dixon, pers. comm., 2001), to be less damaging demographically than is removing vegetative plant parts. This conclusion is based in part on demographic modeling by Menges (1992: 226), in which he wrote, “The threat posed to population survival by environmental variation appeared almost entirely due to variation in mortality, growth, and reproduction status and not to variation in reproductive output.” Just as seed collection increases environmental variation in reproductive output, taking cuttings increases the variation in growth rate and possibly mortality. The second reason is more pragmatic. At this time, it is generally much easier and more economical and effective to store large numbers of seeds in a seed bank (assuming, of course, that such facilities are available) than it is to maintain even much smaller numbers of actively growing plants in a botanic garden or other off-site setting. Emerging technologies, such as cryopreservation of vegetative tissues (e.g., stem segments with lateral buds), when more widely evaluated and available, may provide additional storage options. However, seed storage is not necessarily a realistic option, even for taxa with orthodox seeds. If there is a choice of propagule type (seeds versus cuttings), and seeds can be stored alive for long periods of time, we believe that sampled plants and populations are much better off if seeds, rather than plant parts or, worst of all, whole plants are taken. We also acknowledge that this is not always possible or desirable. But the choice itself may require samples for viability testing, which brings us to another basic consideration, the explicit consideration of which marks perhaps the most significant development since the original guidelines. For what purpose is a sample being taken?
Purposes of a Conservation Collection The purpose for which a collection is made strongly influences both sample size and sampling design. At one extreme, for some collection purposes, such as obtaining material to learn details of a species’ germination and propagation needs or the storage behavior of its seed, little material may need to be gathered, without much regard to the genetic makeup of the collection. At another extreme, some collection purposes, such as salvaging genetic material from a doomed population for use in storage and reintroduction, may necessitate large samples taken from all individuals. Note also that it may not always be necessary to collect additional material from the wild. For some purposes, suitable material (seeds, growing plants, or plant parts) may be available from other sources, such as existing samples in seed banks, in vitro cultures, field gene banks, or other cultivated
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sources. If genetically and otherwise biologically appropriate material is available in suitable off-site storage, it may mean that new collections need not be made from wild populations.
Purpose: To Develop Germination and Propagation Protocols or to Determine Seed Storage Behavior As a general rule, it is prudent not to collect material in volume before reliable methods are available to maintain it, given the potential negative impact of collection on sampled populations. Collection pressure can also be reduced by conducting pilot studies on closely related but more common congeners. In practice, however, there are taxa and situations for which the threat of extirpation in the wild is so great that more extreme measures might be justified. In most cases involving narrowly distributed endemic plants, it is reasonable to assume that different populations of a taxon would have similar, if not identical germination and propagation requirements and seed storage behavior. Therefore, and in the absence of obvious ecological differences between populations, there is no a priori need for a statistically representative sample, as there is, for example, in collecting for long-term storage or rare plant reintroduction. Samples collected for the latter two purposes should be taken from sources that are least likely to harm species or population survival prospects in the wild. In other words, obtain seeds (or cuttings) from the largest, most secure (or at least most dispensable) sources possible. Seeds from properly identified and documented cultivated specimens generally are acceptable for developing propagation protocols. Recommendation: For developing germination and propagation protocols, or to determine seed storage behavior, use existing ex situ material if available. For extremely rare taxa, it may be advisable to begin with pilot studies using closely related but more common congeners. If wild populations must be sampled, begin with small collections from the largest and most secure. Actual sample sizes should be determined in consultation with those who will be working with or who are familiar with the material in question.
Purpose: Ex Situ Storage The collection and maintenance of large, genetically representative samples can serve as a hedge against catastrophic loss in wild populations and can provide material for reintroduction and other conservation efforts. Such col-
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lections should be made without unduly compromising sampled populations, and this is clearly easiest, most economical, and most effective to accomplish for taxa with long-lived, orthodox seeds. The numbers and genetic diversity of such collections are strongly influenced by the number and size of extant populations from which to collect. And again, the numbers needed for storage depend greatly on what purposes the stored seeds are intended to serve. Should an off-site collection be expected to support a single reintroduction attempt? Two? Ten? Are there other purposes, such as unanticipated scientific research efforts, that an off-site collection might be expected to support? Relative to stored seed, the cost to maintain off-site samples as actively growing plants is much greater, and the probability of successfully perpetuating the genetic integrity of stored material is much less (Appendix 3, this volume). Presumably somewhere in between are ex situ samples maintained as slow-growing tissue culture or somatic tissue stored cryogenically. The genetic integrity of stored seed or other frozen samples also is much more easily and effectively maintained than it is for population samples maintained as actively growing plants. This is thought to be true for several reasons. First, the expected longevity of stored seed generally is much greater than for growing plants. Second, assuming that proper seed storage facilities and techniques are available, both the absolute and relative costs of maintaining the original genetic array of a collection are much less for seeds than for growing plants. Third, it is extremely difficult, if not impossible, to provide habitats off site that are sufficiently similar to those experienced in the wild to avoid artificial selection for unknown properties. Therefore, in addition to the deleterious genetic effects resulting from random genetic drift caused by the small population sizes of a living collection, the genetic adaptedness of growing samples is expected to deteriorate much more quickly in growing collections than in dormant seed collections. Finally, there are phytosanitary and related considerations that apply to growing plants but not stored seed. Recommendation: Maintain ex situ collections as dormant seed, if possible. For seed storage, sample size is limited mostly by available resources, such as the size of the sampling universe, the impact of collection on wild populations, and the technical capability to store seeds for a long period of time. The limit for the number of growing plants is set more by the practical constraints in handling a species than by other factors, so the total number generally is lower for growing plants than for seed storage.
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Purpose: Reintroduction, Including Conservation Introduction, or Augmentation Reintroduction is not a simple one-size-fits-all procedure. For example, experience suggests that the reintroduction of a rare population into a habitat from which it has been extirpated and introduction to a new, apparently suitable but as yet unoccupied habitat hold different ecological considerations, including reasons for extirpation in the former case and reasons for the absence of colonization in the latter case. If biologically and genetically suitable and appropriate propagation material is already stored off site, it should be considered for use before new collections from wild populations are made. Actual sample sizes depend heavily on the management questions being asked of the experiments, and other aspects of the reintroduction plan being considered, such as monitoring and contingency plans. Recommendation: To develop reintroduction protocols, begin with the smallest collections necessary to address the management questions being posed in the experimental reintroductions. Sample sizes necessary to support actual reintroductions, enhancements, or augmentations can vary widely. In general, the larger the founding population, the greater will be the chance of it surviving to become an established, self-sustaining population (Guerrant 1996). But it is important to bear in mind that not all reintroduction attempts will succeed, even of species for which propagation protocols have been established empirically. The number of reintroduction attempts that a collection is intended to support and their geographic limitations also greatly affect the sample size needed. Recommendation: To increase the probability of successfully reintroducing self-sustaining populations of threatened plant species, collect from as large and diverse an array of suitable founders as is prudent, given the sampling universe. Collect and maintain separately seeds from each maternal line. Only in this way can representation of the different founders be known and controlled intentionally.
Purpose: Other Potential Uses Sample sizes necessary to satisfy other uses, such as scientific research, education, and interpretation, vary so widely that no general recommendations are possible.
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Recommendation: Collection for other purposes should be evaluated in light of their intrinsic estimated conservation value and in light of the cumulative impact of all collection activities anticipated for those species and populations.
Evaluation Cycle With the choice of taxa made and purposes of a collection articulated, the information can be collated and integrated, and preliminary estimates of sample sizes necessary to serve them can be made (Figure A1.1). Before a final collection strategy can be adopted, initial sample size estimates must be put through two additional filters. The first is the various sources and magnitude of attrition that can reasonably be expected before use. To account for expected losses, sample size estimates must be increased. If appropriate material already exists in ex situ collections, however, the estimated sample sizes may be reduced accordingly. The preliminary revised sample size estimates reflect expected attrition rates and extant ex situ samples. The revised estimate is then put through the last filter, which is an assessment of the probable impact such a collection has on the sampled population. If the expected impact is judged to be too great, then this conclusion is added to the original mix of inputs, some accommodation made, and the cycle repeated. Only when the impact of further collection is judged acceptable should collection proceed.
Sources of Attrition from Collection through Successful Use It is one thing to collect a genetically representative population sample and quite another to maintain such a collection until it is needed to serve the intended purpose, such as establishing a new, genetically comparable population. There are many steps along the way in which mortality and other losses can occur, in terms of both sheer numbers and genetic diversity. In this section, we consider various sources of attrition, what it takes to monitor them, and how losses can be mitigated. Two major sources of attrition are considered. The first is mortality while in ex situ storage itself, and the other might be considered the demographic cost of reintroduction: the residual fraction of propagules planted that survive and reproduce successfully (Chapter 17, this volume).
Monitoring Survival Rates of Stored Seed Although estimates of longevity for dormant seed maintained properly are very encouraging—perhaps on the order of decades to centuries for many
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orthodox seeded species—some mortality is to be expected. Unlike growing plants, a dead seed generally appears indistinguishable from a live one. In addition to destructive tests such as cut tests or tetrazolium staining tests, it is generally necessary to attempt to germinate subsamples of a collection to ascertain its viability or at least germinability (Chapter 17, this volume). Unless large sample sizes are available (which is typically not the case, especially if seeds from each maternal parent are maintained separately), monitoring viability with a high degree of precision is not likely to be possible. As far as we know, neither this dismal statistical fact nor its implications have been fully assimilated by the ex situ community. Recommendation: If a species of conservation concern exists in ex situ collections, the survivorship, health, and genetic status of the off-site collections should be monitored. To minimize genetic changes in ex situ conditions, emphasis should be placed on improving storage or cultural conditions rather than or at least before additional wild collection.
Demographic Cost of Reintroduction Population size targets, often specified as numbers of reproductively mature plants, are commonly indicated in reintroduction plans for specific projects. It is not reasonable to expect that all propagules planted will survive to reproduce, but what is a reasonable expectation? To estimate the range of postplanting decline in population size that might be expected during reintroduction, Guerrant and Fiedler (Chapter 17, this volume) used empirically derived stage-based transition matrices for a variety of plants with life histories as a basis for stochastic modeling. Not surprisingly, they found that the demographic cost during reintroduction can be substantial. In the most extreme case, an outplanting of 1,000 Panax (Araliaceae) seedlings would, on average, be expected drop to just 15 individuals within 3 years before the simulated populations began to rise. Of course, many simulated runs ended with extirpation before any increase could begin. If the newly established populations are to have anything like the genetic diversity of the ones from which the founders were collected, expected losses during reintroduction must be accounted for in the original collection. These data are simulated results based on wild populations with positive growth rates. One assumption of these models is that outplanted individuals will behave demographically identically to naturally occurring plants, which
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is probably optimistic. Another assumption of the models is that the series of years for which data were gathered in the field accurately reflect what will happen during a reintroduction. Presumably there will be many stochastic environmental effects that cannot be anticipated but will affect establishment. In a series of experiments using similar techniques and comparable seed supplies, 27 field germination and seedling establishment trials of Erythronium elegans (Liliaceae) set out with fresh seed each year over a 5-year period spanned the range from 0 to 94 percent establishment (Guerrant 1999; Chapter 17, this volume). Clearly, attrition can be high and vary greatly between years. The implications for any collection effort to support even one reintroduction attempt are daunting. Expected losses of such magnitudes as observed in the stochastic modeling for Panax and Erythronium suggest that sample sizes might need to be one or more orders of magnitude greater than current guidelines suggest. Unfortunately, such collections may be too large for sampled populations to bear, prohibitively expensive in time and other resources needed to collect, store, and monitor, or simply not possible given the size of the extant population. In addition to increased sample sizes, other ways to compensate for potential losses associated with reintroduction must be explored. One potentially useful alternative is to use physically larger founding individuals, which might be expected to have greater survivorship than physically smaller founders. Of course, there are practical limits to using larger individuals as founders. Likewise, any postplanting care that can be provided to increase survivorship of the founding individuals should reduce sample size needs and predicted losses to attrition. Recommendation: To compensate for propagule mortality during reintroduction, start with an estimate of desired numbers of individuals surviving to reproduction in a new founding population. Then, account for expected losses during establishment. Some of these calculated losses can be mitigated by maintaining backup clonal material.
Effect of Collection on Extinction Risk The final question posed by the CPC genetic sampling guidelines concerns the level of collection that necessitates a multiyear collection strategy. Menges et al. (Chapter 15, this volume) used computer simulation to study the expected impact of collection on extinction risk for plants with a variety of life histories. Using empirical data as a basis for stochastic modeling, they
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looked at the impact on populations of different sizes (10, 50, 100, and 500 plants) with varying levels of intensity (10, 50, and 100 percent of seed production) and frequency of collection (10, 50, and 90 percent of years). They found that species differed in sensitivity to seed harvest, with long-lived species, especially woody plants, least sensitive. Populations of 500 or more generally were not harmed except by complete harvests for half or more of all years. Small populations of 10 were harmed by less complete harvesting, but sensitivity varied widely by species. Menges et al. (this volume) offered three seed harvest rules: • Harvesting 10 percent of seeds in 10 percent of years (or less) is
generally safe. • Harvesting 50 percent of seeds in 50 percent of years (or more) is
generally unsafe. • Less intense, frequent harvests are safer than more intense,
infrequent harvests. The first two essentially set quantitative brackets around the third, which is the qualitative heart of the recommendation. Truly safe levels and frequency of harvest depend very much on the population dynamics of a particular population at a particular time and place. We are not suggesting that a level of 10 percent of seeds in 10 percent of years should be considered standard. Rather, this represents a generally safe level and frequency in extreme situations, understanding that other factors elevating concern over total loss of fragile populations may override the general dictum to “Do no harm.” Recommendation: Less intense, frequent harvests are expected to have lower impact on sampled populations than more intense, infrequent harvests. To the degree possible, spread collection over 2 or more years, especially for small populations.
Extreme Situations As discussed at the beginning of these guidelines, it is wise not to collect material in volume before methods are available to use it judiciously, given the potential negative impact of collection on sampled populations. In practice, however, there are taxa and situations for which the threat of extirpation in the wild is so high that more extreme measures might be justified, situations in which it might be necessary to act sooner rather than later.
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Menges et al. (Chapter 15, this volume) modeled populations no smaller than 10 individuals. Part of their reasoning was the belief that populations this small and especially smaller are inherently threatened with extinction simply by chance. They noted that declining populations represent special cases in which considerations other than the harm done by collection itself might become important. For example, if a population is in decline and sliding toward extirpation anyway, collection affects only the timing, not the end result. In such cases, the potential benefits of collection must be weighed against the additional pressure of collection on extinction risk. Another area not covered directly in the models concerns very small and other populations for which the probability of extirpation in the foreseeable future is so high that rescue collections might be of conservation value and therefore worth the additional risk. The question then arises of what to do with very small populations that might be particularly susceptible to extirpation in the near to medium term (5–25 years). Although it is always best to keep in mind the dictum “Do no harm,” it may be necessary in some situations to collect so much material that collection itself becomes a serious threat to the sampled wild population, at least in the short term. There are cases in which future prospects are so bleak and successful reproduction so unlikely to happen in the wild that it may be prudent to harvest many if not all propagules for conservation use. The efforts to recover the California condor (Gymogyps californianus) and the black-footed ferret (Mustela nigripes) are two cases in point. When only a handful of individuals were believed to exist in the wild and even fewer in captivity, emergency efforts were made to capture all remaining individuals, thus driving the species to extinction in the wild, at least temporarily. In both cases, these animals were used in captive breeding programs whose primary goals are to release many more individuals into the wild than were removed, in more areas than just the original collection sites. Thus, we may ourselves be in the uncomfortable position of destroying a wild population in order to save it. Recommendation: For populations of species with extremely low overall numbers, particularly those that have 10 or fewer reproductive individuals and a poor history of recruitment or those that are known to be in precipitous decline, collect up to 100 percent of seed at the discretion of the permitted collector. Such collection levels assume that adequate facilities, procedures, and resources are available to care for the material and that such collections are part of a more inclusive strategy that is endorsed by the appropriate regulatory authorities.
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Worksheets A series of worksheets are provided to assist practitioners in estimating appropriate sample sizes (Figures A1.2–A1.4). Figure A1.2 can be used to summarize the status of a taxon in terms of how many populations of what size are known and of its biology and provides space to organize an initial needs assessment of what sample sizes are needed to supply anticipated purposes. The second and third worksheets bring the analysis to the population level. Figure A1.3 provides space to evaluate five populations for how many propagules each might supply to serve various networks. For taxa with more than five populations, multiple copies of this worksheet may be needed. The final worksheet (Figure A1.4) allows initial sample size estimates to be viewed in the context of existing collections and for final target collection sizes to be set. It may be necessary to cycle through the latter two worksheets several times during the evaluation process before a final target is decided.
Conclusions The basic structure of the collection guidelines offered in the original CPC guidelines is sound. However, this appendix recommends that the actual number of propagules should be revised upward, perhaps by one or more orders of magnitude in some instances. In the most recent and thorough statistical treatment of sampling strategy of which we are aware, Brown and Marshall (1995) have a benchmark target of 50 individuals per population in each of 50 populations per ecogeographic region per taxon. We endorse this strategy as a first approximation against which actual sample sizes are determined. All numbers are subject to change, and any collection strategy must be tempered with consideration for the purpose of collection, ability to maintain the samples in good condition off site, and damage to wild populations done by collecting itself. After all, off-site samples are part of a larger integrated conservation program, the ultimate purpose of which is to increase the long-term survival prospects of sampled populations in the wild. References Brown, A. H. D., and J. D. Briggs. 1991. Sampling strategies for genetic variation in ex situ collections of endangered species. Pages 75–86 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Brown, A. H. D., and D. R. Marshall. 1995. A basic sampling strategy: theory and practice. Pages 75–91 in L. Guarino, V. Ramanatha Rao, and R. Reid (eds.), Collecting
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Plant Genetic Diversity: Technical Guidelines. Wallingford, UK: CAB International. CPC (Center for Plant Conservation). 1991. Genetic sampling guidelines for conservation collections of endangered plants. Pages 225–238 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Falk, D. A. 1987. Integrated conservation strategies for endangered plants. Natural Areas Journal 7:118–123. Falk, D. A. 1990. Integrated strategies for conserving plant genetic diversity. Annals of the Missouri Botanical Garden 77:38–47. Falk, D. A., C. I. Millar, and M. Olwell (eds.). 1996. Restoring Diversity: Strategies for Reintroduction of Endangered Species. Washington, DC: Island Press. Guarino, L., V. Ramanatha Rao, and R. Reid (eds.). 1995. Collecting Plant Genetic Diversity: Technical Guidelines. Wallingford, UK: CAB International. Guerrant, E. O. Jr. 1996. Designing populations: demographic, genetic, and horticultural dimensions. Pages 171–207 in D. A. Falk, C. I. Millar, and M. Olwell (eds.), Restoring Diversity: Strategies for Reintroduction of Endangered Species. Washington, DC: Island Press. Guerrant, E. O. Jr. Comparative demography of Erythronium elegans in two populations: one thought to be in decline (Lost Prairie), and one presumably healthy (Mt. Hebo). Final report on five transitions, or six years of data. Unpublished report prepared for the USDI Bureau of Land Management and USDA Forest Service. Guerrant, E. O. Jr., and B. M. Pavlik. 1998. Reintroduction of rare plants: genetics, demography and the role of ex situ conservation methods. Pages 80–108 in P. L. Fiedler and P. Kareiva (eds.), Conservation Biology for the Coming Decade. 2nd edition. New York: Chapman & Hall. Lande, R. 1995. Mutation and conservation. Conservation Biology 9:782–791. Lynch, M., J. Conery, and R. Bürger. 1995. Mutation accumulation and the extinction of small populations. American Naturalist 146:489–518. Marshall, D. R., and A. H. D. Brown. 1975. Optimum sampling strategies in genetic conservation. Pages 53–80 in O. H. Frankel and J. G. Hawkes (eds.), Crop Genetic Resources for Today and Tomorrow. Cambridge, UK: Cambridge University Press. Marshall, D. R., and A. H. D. Brown. 1983. Theory of forage plant collection. Pages 135–148 in J. G. McIvor and R. A. Bray (eds.), Genetic Resources of Forage Plants. Melbourne: CSIRO. Menges, E. S. 1992. Stochastic modeling of extinctions in plant populations. Pages 253–275 in P. L. Fiedler and S. K. Jain (eds.), Conservation Biology: The Theory and Practice of Nature Conservation, Preservation and Management. New York: Chapman & Hall. Touchell, D. H., M. Richardson, and K. W. Dixon (eds.). 1997. Germplasm Conservation Guidelines for Australia: An Introduction to the Principles and Practices for Seed and Germplasm Banking of Australian Species. Canberra: Australian Network for Plant Conservation.
Taxon ____________________ Population number and sizes ____________________ Seed storage behavior: Orthodox, Intermediate, Recalcitrant, Unknown Life history: Annual, Short-lived perennial, Long-lived perennial, Woody, Herbaceous, Clonal Breeding system: Selfer, Outcrosser, Mixed mating system, Unknown Reproductive output: Seeds per fruit (one, few, many, or mean number if known) Fruits produced per plant per year ______________ Germination fraction ______________ Knowledge status: Germination ______ Propagation (standard horticulture) _____ Propagation (in vitro) ______ Seed storage behavior _______ Indicate sample sizes and source populations for each purpose and for each time period. Purpose of Collection
Near-Term Needs (1–3 years)
Medium-Term Needs (3–7 years)
Eventual Needs (7+ years)
To develop protocols Germination Propagation (standard horticulture) Propagation (in vitro) Seed storage behavior Ex situ storage Orthodox seed Attrition (rate) Recalcitrant seed Attrition (rate) Cryopreservation of tissue samples In vitro slow growth Attrition (rate) In cultivation Attrition (rate) Reintroduction Attrition rate (including demographic cost) Augmentation Attrition rate (including demographic cost) Other Figure A1.2 Genetic Sampling Guidelines Worksheet: Preliminary need assessment work-
sheet on which to summarize biological status and knowledge of a taxon and organize initial estimates of what sample sizes might be necessary to serve various potential needs for ex situ material. (For a printable version, go to http://www.islandpress.org/appendix/exsitu/ worksheetA1_2.html)
Taxon _________________ Page ___ of ___
For each population, indicate name and number of mature and juveniles above preliminary target numbers for collection. 1
2
Figure A1.3 Genetic Sampling Guidelines Worksheet: Preliminary sample size worksheet for estimates by population to serve various purposes. Sheet contains space for five populations, so multiple sheets may be needed for some taxa. Ind., individuals; Prop., propgaules; Mat., mature; Juv., juvenile. (For a printable version, go to http://www.islandpress.org/ appendix/exsitu/worksheetA1_3.html)
Mat.
Purpose of Collection To develop protocols Germination Propagation (standard horticulture) Propagation (in vitro) Seed storage behavior Ex situ storage Orthodox seed Attrition (rate) Recalcitrant seed Attrition (rate) Cryopreservation of tissue samples Attrition (rate) In vitro slow growth Attrition (rate) In cultivation Attrition (rate) Reintroduction Attrition rate pre-planting Attrition rate post-planting Other Project Attrition (rate) Is multiyear collection plan indicated?
Ind.
Props/ Ind.
Juv. Total
3
Mat. Ind.
Props/ Ind.
Juv. Total
4
Mat. Ind.
Props/ Ind.
Juv. Total
Mat. Ind.
Props/ Ind.
Juv. Total
Multiyear Collection Propagules per Individual Individual In Vitro Growing Plants Seeds Population Size Figure A1.4 Summary worksheet for revised sample size targets, by population, in which existing collections can be taken into account. (For a printable version, go to http://www.islandpress.org/appendix/exsitu/worksheetA1_4.html)
Population
Notes Final Targets for Collection Existing Collections Sampling Universe by Population
appendix 2
Guidelines for Seed Storage Christina Walters
Good storage practices slow down growth and deteriorative reactions in seeds by lowering the temperature or moisture level of tissues. The extent to which storage conditions can be manipulated and the shelf life that is achievable depend largely on the tolerance of seed cells to dehydration or chilling and the initial quality of the seed. The more tolerant the cells, the more moisture level and temperature can be reduced without damage, and the longer the seed will store. This appendix describes basic procedures to maintain seed viability. More detailed descriptions of seed aging and storage procedures are published in Principles and Practices of Seed Storage (Justice and Bass 1978) and review articles (Priestley 1986; Pritchard 1995; Hong and Ellis 1996; Walters 1998a, 1998b) (see also Chapters 6 and 7, this volume, for the scientific foundation on which these practical guidelines rest).
Classification of Seed Storage Behavior Mature seeds generally have higher tolerance of low moisture levels or temperatures than vegetative tissues. They acquire this tolerance during maturation on the parent plant, presumably as a strategy to survive winter or the dry season. The degree of tolerance varies between seed species, and three general categories are generally recognized (but see Chapter 7, this volume): orthodox, recalcitrant, and intermediate. Orthodox seeds, which are produced by most annual or biennial crop and horticultural species (grains, pulse crops, vegetables, florals, and temperate fruit trees), and many temperate forest tree and shrub species can survive complete water removal. These seeds can be easily stored for many years by drying and cooling. Recalcitrant seeds often are produced by herbaceous plants from aquatic habitats, 442
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perennials from tropical areas, and some deciduous forest trees from temperate areas. As the name implies, recalcitrant seeds are difficult to store and have shelf lives usually less than 1 year unless they are cryopreserved. Viability of these seeds can be maintained for one or more growing seasons by storing them at temperatures close to the minimum temperature for germination (Chapter 7, this volume; Pritchard et al. 1996; Tompsett and Pritchard 1998). Seeds with intermediate storage behavior are produced by perennials of tropical and subtropical origin (e.g., coffee, citrus, macadamia, papaya) and by some tree nut species (e.g., walnut, hickory, pecan, hazelnut, and pistachio; hazelnut and pistachio may belong to the recalcitrant category). With proper handling and excellent-quality seeds, viability of seeds with intermediate storage behavior can be maintained for a few years (Ellis et al. 1990, 1991; Hong and Ellis 1995). The best way to identify whether a seed is recalcitrant is to monitor survival while drying (Hong and Ellis 1996; Chapter 7, this volume). Seeds that are recalcitrant lose viability within minutes of drying below about 90 percent relative humidity (RH). Seeds that are intermediate lose viability within days of drying below about 20 percent RH. Seeds that are orthodox can survive many months or even years when dried to 5 percent RH. The ability to survive drying is acquired during development of orthodox seeds, so seeds harvested prematurely appear to be recalcitrant (Vertucci and Farrant 1995). Mature seeds must be dried sufficiently rapidly (within a few days) to avoid deterioration from time at intermediate water contents (Pammenter and Berjak 1999). Often drying experiments are cumbersome, and collectors resort to anecdotal information or the guidelines listed later in this chapter to identify the storage behavior of their seeds. Although guidelines exist to help identify seed storage behavior (Table A2.1; also see Pence 1995), there are no hard and fast rules. A compendium listing storage category of more than 7,000 species is now available (Hong et al. 1996; www.ipgri.cgiar.org/ themes/exsitu/seed_compedium.htm). Useful information on woody species grown in the United States can be gleaned from Schopmeyer (1974) or http://www.wpsm.net.
Seed Storage Behavior and General Storage Principles Because they are sensitive to drying, recalcitrant seeds must be stored under humid conditions (RH 92–98 percent). The temperature at which seeds are best stored depends on the chilling sensitivity of the species. Seeds from many tropical fruits (e.g., mango, avocado, cocoa, jackfruit) are sensitive to
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table a2.1 Guidelines to identify storage behavior of seeds. Trait
Guideline
Some Exceptions
Growth habit
Most herbaceous plants produce orthodox seeds. Many aquatic species, tropical rainforest species, and temperate climax forest species produce recalcitrant seeds. Most orthodox seeds dry naturally on the parent plant. Recalcitrant seeds often are large. Orthodox seeds can survive complete water loss; recalcitrant seeds cannot.
Aquatic species.
Habitat
Water content at harvest Seed size Desiccation sensitivity
Most native Hawaiian species, temperate conifers, some maples. All immature seeds, Solanaceae, Cucurbitaceae. Some aquatic species, Rutaceae, some Rubiaceae. Orthodox seeds dried very slowly (for more than 2 weeks) can be severely damaged.
chilling and should be stored at temperatures not lower than 15°C. These seeds have the shortest potential shelf life, remaining viable for 2 weeks to 3 months. Seeds produced from temperate species (oaks, buckeye, chestnut) can survive for 0.5 to possibly 2 years when stored at 2–5°C. A useful guideline is to store seeds close to the minimum temperature needed for germination (Chapter 7, this volume; Pritchard et al. 1996; Tompsett and Pritchard 1998). This procedure maximizes longevity by slowing germination and limiting chilling injury. Microbial contamination is always a problem at high humidities, and seeds often survive longer if a fungicide is applied. Often storage in damp peat moss is beneficial. Seeds with intermediate storage physiologies are more amenable to drying and consequently can be preserved for longer periods than recalcitrant seeds. RHs between 40 and 60 percent appear to provide maximum longevity (Eira et al. 1999; Sacandé et al. 2000). Intermediate seeds survive 1–6 months if stored at 25°C and 2–5 years if stored at 5°C. These seeds rapidly lose viability if placed in the freezer (Ellis et al. 1990, 1991; Hong and Ellis 1995). Seeds that exhibit orthodox behavior are easily stored by drying and cooling. The extent to which drying and cooling extend shelf life is best described by Harrington’s “Thumb Rules,” which state that the life of a seed is doubled for every 1 percent decrease in seed water content or every 5°C decrease in temperature (Justice and Bass 1978). The “100s Rule” states that adequate
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longevity for commercial seed storage purposes (<5 years) can be achieved by ensuring that the sum of the RH and temperature (in degrees Fahrenheit) of storage does not exceed 100. The “Thumb Rules” were developed as guidelines for seed storage for commercial purposes. They are not valid for extremely dry conditions, and their applicability at extremely cold conditions has not been adequately tested. A new rule is that seeds store optimally at RH between 15 and 25 percent (Walters 1998a, 1998b). Storing seeds at RH <15 percent will not increase shelf life, and research from many labs shows that it may actually accelerate deterioration (Walters 1998a). The water content achieved at 15–25 percent RH varies from about 2 percent to about 10 percent, depending on the lipid composition of the seed and the storage temperature. Research clearly shows that lowering the storage temperature to –18°C (temperature of standard freezer) increases seed shelf life. Whether this is a 16- to 32-fold increase from storage at 5°C, as Harrington’s rules predict, is debatable. Recent research results suggest that there may be a four- to fivefold longer shelf life in seeds stored at –18°C than in seeds stored at 5°C (Walters 1998b). Storage at –18°C is necessary if seeds are to be maintained for more than 15 years.
Seed Quality Factors Influencing Longevity of Orthodox Seeds Harrington’s rules are approximations and give relative, rather than absolute, longevities. Absolute longevities depend on the seed quality. Seed quality is controlled by genetic and environmental factors such as seed structure and composition, maturity, dormancy, purity, mechanical damage, and initial viability and vigor (Justice and Bass 1978; Ellis and Roberts 1980). The field conditions during seed development and harvest and postharvest treatments, such as drying temperature, cleaning procedures, and priming, affect overall seed quality and hence seed longevity (Walters 1998b). Infestations of storage fungi and insects significantly reduce seed quality and life spans. Numerous factors during seed development and maturation affect the potential longevity of seeds. Seeds harvested prematurely may have shorter shelf lives than seeds harvested when fully mature (fully mature is defined as having completed maturation drying). This may be because immature embryos are not fully tolerant of desiccation and so are damaged when dried or because maximum seed quality is acquired during the last stages of seed development on the parent plant. The high temperatures (>35°C) often used by producers to dry immature seeds rapidly can severely damage seeds. Some
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drying procedures call for several days of exposure to 35°C, and this treatment can initiate seed aging processes. Research over the last 10 years has shown that an initial treatment that dries immature seeds slowly enhances seed desiccation tolerance (Pammenter and Berjak 1999) and potentially seed longevity. If seeds must be harvested prematurely, holding them for 2–5 days in their fruiting structures can simulate this slow drying treatment. Once immature seeds receive a slight and slow desiccation treatment, they can be threshed with less damage. Although a brief period of slow drying (less than 5 days at 20°C) enhances the quality of prematurely harvested orthodox seeds, prolonged exposure of fully mature seeds to humidity more than 75 percent promotes deteriorative reactions, encourages microfloral infestations, and, at humidity greater than 95 percent, allows precocious germination (Pammenter and Berjak 1999). Seeds may be exposed to high humidities in the field if harvest is delayed by rain or after harvest if humidity controls are inadequate. Generally, exposure of seeds to high humidity results first in a reduction of seed vigor (slower germination) and then in a reduction in percentage germination. The higher the humidity or temperature, the faster the deterioration (refer to Harrington’s rules). Germination percentages of vegetables and grains may decline to 0 percent within 3–6 months of storage at 20°C and 90 percent RH. It is generally assumed that the reduction in seed quality leads to a reduction in shelf life of seeds once they are placed under more favorable storage conditions. Preharvest and postharvest treatments may explain the variation in longevity between seed lots of the same genetic background (Walters 1998b). Seed priming is a procedure used to accelerate germination rates of planted seeds. Seeds usually are held at temperatures less than 10°C and water potentials close to –1 MPa for a few days. Although this treatment gives the appearance of improving seed vigor (seeds germinate faster or more uniformly), it exposes mature seeds to high humidities and, as the previous paragraph describes, is likely to reduce the shelf life of seeds if they are redried and then stored (Walters 1998b). Anecdotal evidence suggests that seeds harvested from the wild have shorter shelf life than seeds harvested from cultivated plants. There are numerous possible explanations. Through agricultural practices, humans may have selected for seeds with superior quality. Seeds under cultivation may have received optimum growth conditions (irrigation, fertility, and pest management) and so are more robust. Seeds collected from the wild often have a high level of immaturity because maturation of uncultivated seeds is less uniform, and mature seeds often dehisce.
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Seed dormancy often is linked to seed longevity. Seeds with hard seedcoats store longer (Justice and Bass 1978), probably because the seedcoats limit the movement of air and water. Fabaceae and Malvaceae tend toward hard-seededness. The link between embryo dormancy and longevity is difficult to verify because seed dormancy confounds measurements of viability for longevity tests. Clearly, dormant seeds survive longer in the soil, but this is probably because they fail to germinate under moist conditions rather than because they have more efficient mechanisms to survive in the dry state. Given similar storage and harvest conditions, seeds from orthodox species exhibit different inherent longevities. For example, lettuce and onion seeds are fairly short lived, whereas tomato and barley seeds show substantial longevities. Generally, orthodox species from Fabaceae and Malvaceae are long lived, whereas many species in Asteraceae or Apiaceae produce shortlived seeds. Longevity of seeds from annual crops is better described than that from herbaceous or woody perennials. Sample longevities of different seed species are listed in Table A2.2. More exhaustive species lists and detailed storage conditions can be gleaned from Justice and Bass (1978), Priestley (1986), and Hong et al. (1996). Although species have characteristic longevities, variability between cultivars and between lots of the same cultivar may be so great as to preclude prediction of shelf life. For example, germination of sesame seeds with initial germination rates of 70–80 percent ranged from 80 percent to 0 percent after 18 years of storage at the National Seed Storage Laboratory. Similar results were obtained for potato, pepper, sorghum, onion, and tomato (initial germination 90–100 percent and germination after 18 years ranged from 100 to 0 percent). The source of this variability in shelf life is being researched; until it is understood, seed bank operators should monitor the viability of their seeds periodically to ensure quality.
Additional Storage Factors Affecting Shelf Life As stated previously, the RH and temperature of storage are the most important factors affecting shelf life of seeds. The gaseous environment may also be important, but evidence of a clear relationship is lacking. Most often cited is the beneficial effect of reduced oxygen tensions. Intuitively, this may make sense because aging reactions are believed to be oxidative. Seed storage under higher oxygen tensions causes more rapid deterioration (Ohlrogge and Kernan 1982), but the converse has not been demonstrated to date. Freeze drying or vacuum packaging seeds may be beneficial (Woodstock et
table a2.2 Approximate longevities for various seed storage behaviors and seed species. Storage Type
Species
Recalcitrant Recalcitrant Recalcitrant Recalcitrant Intermediate Intermediate Intermediate Intermediate Orthodox Orthodox Orthodox Orthodox Orthodox Orthodox Orthodox Orthodox Orthodox Orthodox
Tea (Camellia sinensis) Buckeye (Aesculus hippocastanum) Trifolate orange (Poncirus trifoliata) Wildrice (Zizania palustris) Coffee (Coffea arabica) Papaya (Carica papaya) Hickory (Carya spp.) Citrus (Citrus limon) Lettuce (Lactuca sativa) Onion (Allium cepa) Peanut (Arachis hypogeae) Soybean (Glycine max) Sunflower (Helianthus annuus) Grain corn (Zea mays) Chickpea (Cicer arietinum) Pea (Pisum sativum) Barley (Hordeum vulgare) Tomato (Lycopersicum esculentum)
Optimum Relative Humidity
Optimum Moisture Content (g H2O/g dw)a
95–98% 95–98% 90–95% 90–95% 40–60% 40–60% 80–90%
0.6–0.8 0.5–0.7 0.5–0.7 0.35–0.45 0.10–0.13 0.09–0.11
20% 20% 20% 20% 20% 20% 20% 20% 20% 20%
0.04–0.05 0.06–0.08 0.04–0.05 0.07–0.08 0.03–0.04 0.07–0.09 0.07–0.08 0.09–0.12 0.09–0.12 0.05–0.06
Time to 50% Loss of Viabilityb
2 weeks–2 months 6–8 months 6–14 months 9–18 months 2–4 years at 5C; damage at 0C 3–6 years at 5C; damage at 0C 3–5 years 6–18 months 4 years at 5C, 20 years at 18C 4 years at 5C, 20 years at 18C 4 years at 5C, 20 years at 18C 5 years at 5C, 20 years at 18C 6 years at 5C, 25 years at 18C 8 years at 5C, 25 years at 18C 8 years at 5C, 25 years at 18C 10 years at 5C, 25 years at 18C 10 years at 5C, 25 years at 18C 12 years at 5C, 25 years at 18C
a For orthodox seeds, the lower value in the water content range is more appropriate for 5C storage, and the higher value approximates optimum water contents for constant 18C storage. Water contents on a percentage dry weight or fresh weight basis often are used in the seed industry. To calculate percentage water content on a dry weight basis, simply multiply water contents given in the table by 100. b Values for approximate longevity are given for 5C storage for all species except Camellia sinensis, which is chilling-sensitive and should be stored at 15C.
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al. 1976; Justice and Bass 1978; Ellis et al. 1993), but it is not clear whether these treatments control moisture level or oxygen tension. Some have argued that respiratory processes cause seeds to age, and so reduced oxygen will slow the aging process. This reasoning is valid only for high-humidity storage (RH > 80 percent), where storage microflora can grow. Reducing storage RH is a far more effective method of limiting respiratory processes. The time and expense of substituting low-oxygen atmospheres probably are not worthwhile for short-term storage of orthodox seeds. Exposure to light—even low intensities—appears to damage dry seeds. Seeds intended for long-term storage should not be dried under the sun and should not be stored in a lighted room without a protective container.
Procedures to Dry Orthodox Seeds At storage temperatures greater than 0°C, seed water content or RH is the most important factor determining seed longevity. Seed producers must consider two factors when drying seeds: the rate at which seeds are dried and the level to which they are dried. With the exception of immature seeds that benefit from a short, slow drying period, orthodox seeds with high water contents (>18 percent water) should be dried to about 12 percent water content as rapidly as possible. Temperatures higher than 35°C can be damaging, so lower temperatures and high airflow are recommended. Seeds should be spread out in a thin layer to allow air to circulate through the seed mass and gently mixed daily to ensure even drying. Once seeds have been dried to a safe water content (<12 percent), optimum water contents for storage can be achieved if they are placed at a known RH. There are two strategies: packaging seeds in moisture-permeable bags and storing them in rooms where RH is mechanically controlled or equilibrating seeds to the desired RH and then packaging the seed in moisture-impermeable containers (glass, foil laminate, aluminum cans). In the moisture-proof containers, RH becomes a function of the water content of the seed and the eventual temperature at which the seed is stored. One of the questions most frequently asked by seed handlers is, What RH should I use to adjust seed water content? The answer is simple if seeds are to be stored in an RH-controlled room: 20–25 percent RH provides the optimum moisture level for seed storage of most orthodox species (Walters 1998a, 1998b; Chapter 6, this volume). The water contents achieved vary between seed species according to the lipid composition and the storage temperature: peanut
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table a2.3 Recommended drying conditions for seeds stored in moisture-proof containers at various temperatures. The given drying temperature and RH combinations give a storage RH of 20% at the indicated storage temperature. Drying Temperature (C)
Drying RH for Storage at 15C
Drying RH for Storage at 5C
Drying RH for Storage at 18C
25 15 5
28 20 14a
33 26 20
46 38 32
a Drying seeds at temperatures less than the storage temperature is not cost-effective and therefore strongly discouraged: dehumidification is more difficult at lower temperatures, and the refrigeration costs used during drying might be more effectively spent during storage.
seeds (45 percent lipid) contain 2–5 percent water, whereas pea seeds (2 percent lipid) contain 8–12 percent water (Chapter 6, this volume). When RH is controlled, water content need not be monitored unless the RH sensor is suspected of malfunctioning. The simplicity of this method may cause some producers to use it exclusively; however, constant dehumidification, especially in conjunction with refrigeration, may be prohibitively expensive. Depending on the volume of seeds handled and the risk of mechanical failure, it may be safer and more cost-effective to store seeds in moisture-proof containers. The complexity of adjusting water contents increases when seeds are stored in moisture-proof bags, but the flexibility in choosing drying and storage conditions also increases. The interested reader may refer to more indepth literature on water sorption isotherms of seeds (Walters 1998a, 1998b; Chapter 6, this volume). The guiding principle is that 20–25 percent RH provides optimum storage. But because RH is a function of temperature, both the drying and storage temperatures must be considered (Figure A2.1). Generally, seeds are dried at ambient temperatures or slightly higher to speed up the drying process and reduce refrigeration costs. The greater the temperature difference between the drying and storage temperatures, the higher the allowable RH for drying (Figure A2.1). For instance, if seeds are dried at 25°C and stored at 15, 5, or –18°C, they should be equilibrated to about 28, 33, or 46 percent RH, respectively. Table A2.3 gives a brief summary of recommended drying conditions for specific storage temperatures, and Figure A2.1 shows drying RH and temperature combinations that give 20 percent RH at storage temperatures between –20 and 40°C. To date, these guidelines have proven reliable for all species tested.
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Figure A2.1 Drying relative humidity and temperature combinations needed
to achieve 20 percent relative humidity at storage temperatures ranging from –20 to 40°C. Data are taken from water sorption isotherms constructed for pea at a range of temperatures (Chapter 6, this volume). The lower portion of the graphed surface (drying RH ≤ 20 percent) depicts conditions used exclusively for research because storing seeds at higher temperatures than they are dried would be inefficient. The upper portion of the graphed surface represents conditions in which the drying temperature is higher than the storage temperature (most common practice) and shows the flexibility available to gene bank operators in choosing conditions to optimize the moisture level of their seeds. The graph clearly shows that drying RH must be chosen in the context of the drying temperature and storage temperature.
Assessing Changes in Seed Quality No storage procedure guarantees that seeds will remain viable forever. Seeds eventually lose vigor and then viability with time. The extent to which aging occurs can be monitored with initial and subsequent germination assays. Different seed species have different germination needs that are cataloged in rule books published by the Association of Official Seed Analysts (AOSA 1999). Assessments of seed vigor are more difficult than assessments of percentage germination but usually provide an early warning of deterioration.
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References AOSA (Association of Official Seed Analysts). 1999. Rules for Testing Seeds. Lincoln, NE: Association of Official Seed Analysts. Eira, M. T. S., C. Walters, L. S. Caldas, L. C. Fazuoli, J. B. Sampaio, and M. C. L. L. Dias. 1999. Tolerance of Coffea spp. seeds to desiccation and low temperature. Revista Brasileira de Fisiologia Vegetal 11:97–105. Ellis, R. H., T. D. Hong, M. C. Martin, F. Pérez-García, and C. Gomez-Campo. 1993. The long term storage of seeds of seventeen crucifers at very low moisture contents. Plant Varieties and Seeds 6:75–81. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1990. An intermediate category of seed storage behaviour. I. Coffee. Journal of Experimental Botany 41:1167–1174. Ellis, R. H., T. D. Hong, and E. H. Roberts. 1991. Effect of storage temperature and moisture on the germination of papaya seeds. Seed Science Research 1:69–72. Ellis, R. H., and E. H. Roberts. 1980. Improved equations for the prediction of seed longevity. Annals of Botany 45:13–30. Hong, T. D., and R. H. Ellis. 1995. Interspecific variation in seed storage behaviour within two genera: Coffea and Citrus. Seed Science and Technology 23:165–181. Hong, T. D., and R. H. Ellis. 1996. A Protocol to Determine Seed Storage Behaviour. IPGRI Technical Bulletin #1. Rome: International Plant Genetic Resources Institute. Hong, T. D., S. Linnington, and R. H. Ellis. 1996. Seed Storage Behaviour: A Compendium. Handbooks for Genebanks no. 4. Rome: International Plant Genetic Resources Institute. Justice, O. L., and L. N. Bass. 1978. Principles and Practices of Seed Storage. Agriculture Handbook no. 506. Washington, DC: USDA. Ohlrogge, J. B., and T. P. Kernan. 1982. Oxygen-dependent aging of seeds. Plant Physiology 70:791–794. Pammenter, N. W., and P. Berjak. 1999. A review of recalcitrant seed physiology in relation to desiccation-tolerance mechanisms. Seed Science Research 9:13–37. Pence, V. C. 1995. Cryopreservation of recalcitrant seeds. Pages 29–50 in Y. P. S. Bajaj (ed.), Biotechnology in Agriculture and Forestry 32: Cryopreservation of Plant Germplasm 1. New York: Springer-Verlag. Priestley, D. A. 1986. Seed Aging, Implications for Seed Storage and Persistence in the Soil. Ithaca, NY: Cornell University Press. Pritchard, H. W. 1995. Cryopreservation of seeds. Pages 133–144 in J. G. Day and M. R. McLellan (eds.), Methods in Molecular Biology, Vol. 38: Cryopreservation and Freeze-Drying Protocols. Totowa, NY: Humana. Pritchard, H. W., P. B. Tompsett, and K. Manger. 1996. Development of a thermal time model for the quantification of dormancy loss in Aesculus hippocastanum seeds. Seed Science Research 6:127–135. Sacandé, M., J. Buitink, and F. A. Hoekstra. 2000. A study of water relations in neem seed that is characterised by complex storage behaviour. Journal of Experimental Botany 51:635–643. Schopmeyer, C. S. 1974. Seeds of Woody Plants in the United States. Agriculture Handbook no. 450. Washington, DC: USDA. Tompsett, P. B., and H. W. Pritchard. 1998. The effect of chilling and moisture status on the germination, desiccation tolerance and longevity of Aesculus hippocastanum L. seeds. Annals of Botany 82:249–261.
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Vertucci, C. W., and J. M. Farrant. 1995. Acquisition and loss of desiccation tolerance. Pages 237–271 in J. Kigel and G. Galili (eds.), Seed Development and Germination. New York: Marcel Dekker. Walters, C. 1998a. Ultra-dry seed storage. Seed Science Research 8(suppl. 1). Walters, C. 1998b. Understanding the mechanisms and kinetics of seed aging. Seed Science Research 8:223–244. Woodstock, L. W., J. Simkin, and E. Schroeder. 1976. Freeze drying to improve seed storability. Seed Science and Technology 4:301–311.
appendix 3
Guidelines for Ex Situ Conservation Collection Management: Minimizing Risks Kayri Havens, Edward O. Guerrant Jr., Mike Maunder, and Pati Vitt
The most desirable strategy for long-term conservation of plant species as evolutionary units is the protection of viable natural habitats (Pavlik 1994). However, in many parts of the world in situ conservation alone is no longer sufficient because of development and other human impacts. These pressures on native habitats have resulted in the increased application of ex situ conservation of plant species in botanical gardens and seed banks (Given 1984; Heywood 1993; Maunder 1994). Ex situ collections are intended to reduce population extinction risk, preserve genetic diversity, and provide propagules for restoration and recovery programs (Falk and Holsinger 1991; Maunder 1992; Bowles and Whelan 1994; Falk et al. 1996). Given the seriousness of these challenges, ex situ collections should be maintained to preserve representative levels of genetic diversity, to minimize deleterious genetic change, and to reduce genetic risks. Genetic guidelines for seed collection and reintroduction are explored in Chapter 17 and Appendix 1. In this appendix, we focus on the genetic management of living collections, both as growing plants and as dormant seed. Ex situ living collections face many of the same genetic hazards as small populations in the wild, as well as some hazards unique to the ex situ environment. There are three major ways in which the genetic makeup of ex situ populations can shift relative to wild populations: genetic drift, artificial selection, and mutation accumulation. Ex situ populations may also face risks of inbreeding depression, outbreeding depression, hybridization, and problems with pests and pathogens. We will briefly review these phenomena (see also Chapters 12, 13, and 16) and focus on management strategies to reduce their 454
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effects on ex situ collections and present some sample management strategies (Box A3.1). Although genetic risks are inherent in all ex situ activities, different management tools carry different risks (Tables A3.1 and A3.2), and all techniques can be carried out in ways to minimize the risks.
Genetic Drift Genetic drift causes the random loss of alleles or change of allele frequencies. As alleles are lost, homozygosity (having identical alleles at a given locus) increases. Eventually, the loss of alleles results in fixation; some loci become monomorphic. Mean population fitness is likely to decline as deleterious alleles are fixed at some loci. The magnitude of genetic drift increases as population size decreases. It may be exacerbated in ex situ collections if the living collection is maintained at a smaller effective population size (number of individuals that are assumed to contribute genes equally to the next generation) than the wild population. A decrease in effective population size, or Ne, increases the rate of genetic drift, which results in the fixation of alleles and consequently the loss of genetic variation (Ellstrand and Elam 1993). Effective population sizes are difficult to measure in nature but are usually less than the census population size because of variations in family size, unequal sex ratios, fluctuations in population size, and overlapping generations. In most instances the ratio of effective size to census size is between 0.25 and 1.0, but in some organisms, such as captive populations of Drosophila and cichlid fish, it has been found to be less than 0.05. Effective size can be driven to values greater than the census size by genetic management techniques such as equalizing family sizes, sex ratios, and numbers in different generations and removing dominant males from breeding populations (Briscoe et al. 1992; Fiumera et al. 2000). The influence of effective population size on extinction probability has been demonstrated empirically by Newman and Pilson (1997) in Clarkia pulchella (Onagraceae). They varied effective population size while holding census size constant and measured mean population fitness and survival probability. They found that decreased genetic effective population size, which exacerbates drift as well as inbreeding, increased the probability of population extinction. Loss of genetic variation, presumably caused in large part by drift, has been demonstrated empirically in Drosophila. Genetic variation, estimated with both quantitative and molecular techniques, declined with time in captivity, often quite drastically. For example, allozyme heterozygosity was reduced by
box a3.1 Sample Ex Situ Management Plans for Taxa of Conservation Concern For these plans, it is assumed that reintroduction is planned, or a future possibility, so maintaining genetic integrity is a priority. If material is destined for different purposes, such as research or educational needs, managing for genetic integrity may be less important. Taxa with Orthodox Seed • Seed storage is the most cost-effective method of long-term conser-
vation for plant taxa with orthodox seed. • Collect large seed accessions (e.g., at least 1,000 seeds/population)
•
•
•
•
whenever possible, but do not harm the wild population (Appendix 1 and Chapter 15, this volume). Always keep maternal lines separate. When possible, return seed directly to the reintroduction site, avoiding a grow-out at the ex situ facility. If a grow-out is necessary to increase numbers before reintroduction, equalize family size (put out equal numbers of seeds from each maternal line). Subject plants to conditions similar to those they will encounter in the wild. Multiple regeneration cycles should be avoided. If necessary, provide some immigration (five individuals/generation) from wild source population. During a grow-out, keep plants isolated (by bagging or by distance) from congeners and conspecifics from different populations to avoid hybridization and outbreeding depression. To produce seeds, cross plants within the population (or conservation management unit) to minimize outbreeding depression but between maternal lines to minimize inbreeding depression, unless there is a compelling reason to cross between populations (e.g., a population consists of a single self-incompatible clone). Observe phytosanitary precautions and maintain impeccable records.
Taxa with Recalcitrant or Intermediate Seed • Cryopreservation of large numbers of individuals is the long-term con-
servation method of choice if resources (staff, facilities, maintenance costs) are not prohibitive. • Follow same grow-out recommendations as above. • If resources are limited, consider inter-situ methods or field gene banks. These techniques may have a higher risk of maladaptation and other genetic problems but are less costly.
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Critically Endangered Taxa (e.g., fewer than 50 individuals remaining) • In these cases, maintaining every remaining genetic individual (every
potential founder for the next generation) is vital. Through vegetative propagation, one or more replicate ex situ populations can be created and managed for seed production for recovery. Whenever possible, the wild population should be managed for growth. • Whenever possible, conserve a clonal line of every individual through vegetative propagation methods including tissue culture. Cryopreserve or use slow-growth tissue culture to maintain individuals. • Consider inter-situ methods or field gene banks to produce seed from vegetatively propagated individuals. A carefully planned crossing design can increase the effective population size of the off-site populations. Store seed accessions from each clonal line. To minimize population damage, collect seeds judiciously, unless population destruction is imminent. • Consider horticultural management of the wild population to increase population size.
86 percent in one population and 62 percent in another in 60–90 generations despite the large census size of the captive populations. Although the populations appeared large (1,000–3,500), their effective sizes were quite small, as low as 15 in one case (Briscoe et al. 1992). Lacy (1987) has used computer simulations to model the effect of drift on small, managed populations. His data suggest that drift can cause significant losses in heterozygosity, on the order of 60 percent over 100 generations in populations with an effective size of 120. Greater and more rapid losses were seen in smaller populations. He noted that if large wild populations still exist, periodic immigration could drastically reduce drift of the captive population away from the wild population. In his simulation, five or more migrants per generation virtually halted the effects of drift. The benefits of periodic immigration have also been demonstrated in empirical studies. Immigration rates as low as one individual per generation increased fitness and reduced extinction probabilities in captive populations of the housefly Musca domestica (Bryant et al. 1999). Newman and Tallmon (2001) found that 1 or 2.5 migrants per generation increased fitness in recently isolated populations of the plant Brassica campestris (Brassicaceae).
Managing Genetic Drift There are several tools collection managers can use to counter the effects of drift in ex situ populations. The first is to increase the effective population
table a3.1 Descriptions and best applications of the range of ex situ plant conservation techniques (for diagrammatic representation, see Figure 1.1, Chapter 1). Technique
Description
Cryopreservation
Long-term storage of seed, plantlets, apices, pollen, and somatic and zygotic embryos in liquid nitrogen.
Seed banking
Tissue culture (storage)
Tissue culture (propagation)
Cultivation and management in dedicated conservation facility Specialist cultivation in controlled environment
Best Application
Storage of genetically representative samples and specimens (founders or clones) of nonorthodox seeds and vegetative samples from high-priority taxa. Recommended in preference to maintaining cultivated collections over extended periods of time. Long-term storage of seeds in Storage of genetically conditions of low moisture representative samples and temperature. from orthodox-seeded taxa. Recommended in preference to maintaining cultivated collections over extended periods of time. Short-term storage (up to 1 year) Storage of specimens (founders, clones) of highof plantlets and somatic tissue priority taxa when stored under conditions of cryopreservation and seed slow growth (modified light, banking are not feasible. nutrients, temperature). Propagation of somatic tissue, Storage or propagation of embryos, or seeds. specimens (founders, clones) for recovery activities when seed storage and traditional horticultural propagation are not feasible. Storage or propagation of Plants cultivated under taxonspecimens (founders, specific horticultural regime clones) for recovery to serve needs of recovery activities of taxa needing planning. specific or specialized horticultural management. Plants cultivated under artificial Recommended for display and educational material environment, (e.g., tropical only unless carefully species in heated glasshouses managed for genetic in temperate regions). integrity.
table a3.1 (continued) Descriptions and best applications of the range of ex situ plant conservation techniques (for diagrammatic representation, see Figure 1.1, Chapter 1). Technique
Description
Best Application
Cultivation in mixed display or reference collections
Standard method for most threatened species in botanic and horticultural institutions (plants not subjected to genetic management, monitoring, or phytosanitary quarantine procedures). Cultivation standards influenced by institutional curatorial, research, and display policies. Open-air extensive planting to maintain genetic diversity (species or population level). Sustained seed or propagule production. Usually lowintensity nonspecialized management. Cultivated as a profit-generating horticultural or agricultural crop. Management dictated by commercial pressures, with few opportunities for genetic management or control of provenances. Cultivation in village-or community-maintained plots. Management dictated by local community needs and resources, with few opportunities for genetic management or control of provenances. Plants cultivated in horticulturally maintained seminatural conditions.
Recommended for display and educational material only.
Field gene bank
Commercial cultivation
Community garden
Inter situ
Cultivation of long-lived woody taxa where there is a sustained need for seed and propagules supporting species reintroduction and when seed banking is not feasible. Cultivation of valued species and selections to reduce wild collection of threatened taxa or to establish public interest in a local or regional flora. Cultivation of locally valued species and selections of plants (e.g., medicinals) to increase supply and reduce wild collection of threatened taxa and promote local stewardship. Cultivation of horticulturally demanding taxa that need seminatural conditions for regeneration. Suitable for maintaining large numbers of threatened species in seminatural conditions in which surviving but nonviable populations are scattered through degraded habitats.
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table a3.1 (continued) Descriptions and best applications of the range of ex situ plant conservation techniques (for diagrammatic representation, see Figure 1.1, Chapter 1). Technique
Description
Best Application
In situ, horticulturally managed wild
Wild plants subject to horticultural management (e.g., pollination, seed or fruit protection, pest management). Management dictated by recovery needs of the taxon. Wild plants growing in managed habitat and subject to community-level management. Management dictated by restoration or management needs of the community or ecosystem. Wild plants subject to prevailing ecological processes.
Recovery of high-priority taxa that need horticultural intervention to recover a viable in situ population.
In situ, managed wild
In situ, no management
Restoration of threatened habitats and communities and associated recovery of constituent taxa.
Conservation of intact and extensive ecosystems, habitats, and species not threatened by degradation (e.g., invasives).
size (Ne). Recall that Ne can be increased by equalizing family sizes (putting out equal numbers of propagules from each maternal line), sex ratios (important in dioecious plants), and numbers in different generations and by removing dominant males (for plants, equalize pollen donation as much as possible). The debate regarding what effective population size is large enough has not been resolved. Early attempts to answer this question led to the 50/500 rule; estimates of Ne = 50 were widely asserted to be large enough to avoid inbreeding depression and 500 large enough for retention of evolutionary potential and avoidance of drift (Franklin 1980). Now this appears to be an underestimate, perhaps by an order of magnitude or more, for many taxa (Lande 1995). However, in some taxa populations have existed at low census sizes for many generations. Unfortunately, there is not a single quantitative rule that will serve all taxa, but in general larger is better. Additionally, because regeneration reduces genetic diversity and effective population size, sample sizes should increase substantially with each generation. Schoen and Brown (2001) note that for a regenerated sample to conserve with a 95 percent probability all the allelic diversity in the original sample, it must be
table a3.2 Relative strengths and limitations of the range of ex situ plant conservation techniques. Solid circles () indicate high cost or risk, half-filled circles ( ) indicate moderate cost or risk, and open circles () indicate low cost or risk.
Technique
Cryopreservation Seed banking Tissue culture (storage) Tissue culture (propagation) Cultivation and management in dedicated conservation facility Specialist cultivation in controlled environment Cultivation in mixed display or reference collections Field gene bank Commercial cultivation Community garden Inter situ In situ, horticulturally managed wild In situ, managed wild In situ, no management
Long-Term Staff Damaging Inbreeding Disease Capital Maintenance Time Low Controlled Vulnerability Genetic or or Cost per Cost per per Public Seed to Disaster Change Hybridization Outbreeding Pathogens Facility Taxon Taxon Visibility Propagation
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at least three times as large each regeneration cycle. Lastly, periodic immigration from wild populations can alleviate the effects of drift (Lacy 1987; Briscoe et al. 1992). Recommendation: Whenever possible, maintain large effective population sizes. If a population must be maintained off site for a number of generations, provide for periodic immigration from the wild source population of approximately five migrants per generation and increase (triple, if possible) sample size each generation. When possible, use cryostorage or vegetative propagation to lengthen generation time, thus minimizing number of generations in captivity.
Selection Plants adapt to novel environments through natural selection, and this can lead to an adaptation to cultivation. Although data specifically addressing the adaptation of plants to ex situ garden conditions are scarce, there is abundant evidence regarding local adaptation in wild populations and adaptation to captivity in animals. In wild populations, there is good evidence that environmental variation produces genetic heterogeneity (reviewed in Linhart and Grant 1996). Some of the best evidence for fine-scale local adaptation is seen in mine-tailing soils containing high concentrations of heavy metals. Metal tolerance can evolve over just a few generations in a wide variety of plant taxa (Antonovics 1971). Adaptation to heavy metals comes with a tradeoff: metal-tolerant individuals generally do not compete well in nontoxic soils (Cook et al. 1971). The rate of adaptation is a function of the strength of selection and the presence of additive genetic variance for the trait under selection. Adaptation to heavy metals is an extreme example in which selection pressure is intense, but there is also evidence that adaptation to the types of conditions commonly encountered in gardens occurs. Dependence on or tolerance to high levels of fertilizers can evolve rapidly, even after one or two generations in Anthoxanthum odoratum (Poaceae) (Snaydon and Davies 1976, 1982). Herbicide resistance has evolved in numerous taxa (Lebaron 1991). Genetic differentiation between morphotypes of plants living at different elevations on a single hillside has been demonstrated in Avena barbata (Poaceae), presumably because of differences in moisture and temperature regimes (Hamrick and Holden 1979). In Plantago lanceolata (Plantaginaceae) there are genetically differentiated sun and shade populations (Teramura and Strain
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1979). Competition-induced selection has been shown in Veronica peregrina (Scrophulariaceae) (Linhart 1988). Subpopulations exposed to intraspecific or interspecific competition showed adaptive differentiation in seed size, germination timing, growth rate, branching patterns, and overall plant size. Plants in gardens will experience different soils, along with their associated communities, than those experienced under natural conditions. Many species have obligate or facultative associations with mycorrhizal fungi, host plants, or other mutualists (Fiedler and Laven 1996), which can be disrupted in the ex situ environment. Similarly, plants may be removed from their pollinator or exposed to different pollinators in the garden environment, and pollinators can exert strong selective pressure on floral morphology (Galen 1989). Evidence from zoo research about adaptation to captivity is equally compelling and is especially problematic in organisms that have large family sizes. Rapid changes in response to selection have been documented in many species of fish (Harada et al. 1998) and insects, including Drosophila (Frankham and Loebel 1992) and flour beetles (Enfield 1980). Spurway (1955) points out that in zoos, captive breeding practices can select for animals that are less disturbed by transportation and humans, less exacting in their environmental requirements, and less discriminating in mate choice. These traits are expected to prove maladaptive when animals are reintroduced. In ex situ plant conservation programs, unintentional selection can occur at almost every step of the process. Selection occurs for seeds that are mature on the particular day, in the particular season when collecting takes place. Selection occurs for seeds that can withstand ex situ storage conditions and germinate under typical propagation regimes. For example, garden-grown seeds of temperate species typically experience stable stratification temperatures rather than variable winter cold conditions. Numerous species have polymorphic seeds, that is, two or more different types of seeds with different dormancies or germination requirements (Baskin and Baskin 1998). If the different types of dormancies are genetically based, growing out only the early germinants (a common greenhouse practice) would quickly select against deep dormancy types. For example, smaller Cirsium pitcheri seeds tend to germinate in the first growing season, whereas larger seeds germinate the second year (Bowles and McBride, unpublished data, 1998), but it is not yet known whether these differences are genetically based. Selection also occurs for individuals that survive under greenhouse or garden conditions that may be strikingly different from selection pressures faced in the wild. Plants grown ex situ may face less interspecific and more intraspecific
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competition. They may also be isolated from their typical mutualists, including pollinators, seed dispersers, mycorrhizae, and nurse plants, and probably will encounter novel ones. The local adaptation literature suggests that changes in nutrients, water, light, temperature, and competitive regimes can lead to adaptive changes. It is reasonable to expect that many of these adaptive changes may compromise a plant’s ability to survive when reintroduced. Although published examples of reintroduction failures caused by maladaptation are rare, there is at least one example in which cultivated stocks were unable to survive when reintroduced, presumably because they adapted to cultivation. A legume native to the Canary Islands, Lotus berthelotii (Fabaceae), has been grown in European botanic gardens for many generations and was extinct in the wild. Conservationists decided to reintroduce the species back to its native Tenerife. All repatriated plants died in the nursery on Gran Canaria, perhaps as a result of the higher temperatures compared with the European gardens (D. Bramwell, pers. comm., 2002).
Managing Artificial Selection The most effective ways to eliminate the effect of artificial selection are to equalize family size (e.g., put out equal numbers from several maternal lines in the new population), minimize number of generations in captivity, and provide periodic immigration of propagules from wild populations (Allendorf 1993). Recommendation: When creating an ex situ population from wild founders or when reintroducing a species to the wild using ex situ stock, equalize family size. Maintain maternal lines separately to allow family size equalization in future generations. Minimize generations in captivity or, preferably, avoid grow-outs of seed accessions altogether until needed for reintroduction activities.
Deleterious Mutations Although there is much evidence that plants in ex situ environments may adapt to novel garden conditions, a secondary, related problem is the relaxation of natural selection that can occur in cultivation. Plants in gardens generally experience fewer, or different, stresses than their wild counterparts. Resources are rarely limiting; water and nutrients are provided in adequate amounts. Competition is minimized, and pests are controlled. Greenhouse germination results in greater survivorship of seedlings in many taxa, including Cirsium pitcheri (Bowles and McBride 1996) and Asclepias meadii (Asclepiadaceae) (Bowles et
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al. 1998), and this may permit survivorship of less fit individuals. Selection is also relaxed via the practice of equalizing family sizes. One of the consequences of relaxing selection is the accumulation of deleterious mutations. Several authors have noted that the accumulation of mildly deleterious mutations can threaten small populations by decreasing viability and increasing extinction risk (Lande 1994; Lynch et al. 1995a, 1995b). Specifically, in populations with an effective population size less than 100 or a census size less than 1,000, extinction is likely in approximately 100 generations. This is based on the assumption that the organisms in question are under selection. The problem is made worse when selection is relaxed, as it is in many ex situ programs, because mutations accumulate more rapidly when they are sheltered from selection. In the absence of selection, mutations will accumulate at the genomic mutation rate per generation, independent of population size, but the probability of fixation of these mutations by drift depends on the effective population size, and ex situ populations tend to be small. Fitness declines may not become apparent until the population is returned to a selective arena, that is, reintroduced to the wild (Lynch et al. 1995a). Bryant and Reed (1999) found nearly immediate declines in fecundity of 0.5 percent per generation in captive populations of Drosophila. Similarly, Shabalina et al. (1997) found declines in competitive ability in Drosophila on the order of 1.6 percent per generation under relaxed selection. Schoen et al. (1998), modeled mutation accumulation in ex situ plant collections. They concluded that significant fitness declines would be seen in 25–50 generations when sample sizes are less than 75 or when family size is equalized.
Managing Mutation Accumulation Paradoxically, equalizing family size, a practice that slows loss of genetic variation by drift and slows adaptation to captivity, accelerates mutation accumulation by relaxing selection. Several steps can be taken to slow mutation accumulation in our living collections, including reducing the frequency of regeneration of stored seeds. Best practice methods for seed storage can increase the time interval between regenerations. Some storage technologies, such as liquid nitrogen, may also slow or stop the mutational clock. Subjecting plants to naturalistic selective forces such as stress or competition during grow-outs will also help slow the accumulation of mutations. Recommendation: Minimize generations in cultivation and investigate cryogenic storage options. If it is necessary to maintain a living
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collection, consider doing so in conditions that mimic natural selection regimes.
Inbreeding Inbreeding, through self-fertilization or crossing between close relatives, can lead to inbreeding depression, a reduction in fitness in selfed offspring as compared with outcrossed offspring. Inbreeding depression has been documented in numerous plant species (reviewed by Charlesworth and Charlesworth 1987), especially in small populations. The genetic basis of inbreeding depression is not yet clearly understood. The two main hypotheses (which are not mutually exclusive) are known as dominance and overdominance. Under the dominance model, inbreeding depression is caused by the presence of recessive lethal or deleterious alleles. If this is the case, repeated selfing, under conditions that ensure selection against these alleles, should purge the population of its genetic load unless the mutations have only a mild deleterious effect (Charlesworth and Charlesworth 1987; Hedrick 1994). In the overdominance model, inbreeding depression is caused by heterozygote superiority; heterozygous genotypes are more fit than homozygous genotypes (Charlesworth and Charlesworth 1987). For threatened species, it has been suggested that deliberate inbreeding of captive individuals to purge populations of deleterious alleles may decrease inbreeding depression in future generations (Templeton and Read 1983). However, purging genetic load can cause loss of variation at other loci or the fixation of some deleterious alleles, both of which may increase the probability of extinction (Hedrick 1994; Hedrick and Kalinowski 2000). In addition, a recent review by Byers and Waller (1999) indicates that purging may not be effective in the majority of plant species.
Avoiding Inbreeding Depression The key to avoiding inbreeding depression is to maintain large effective population sizes ex situ. This is especially important in taxa that are primarily outcrossing in natural populations. Equalization of family sizes can increase effective population size, resulting in lower levels of inbreeding and a greater retention of genetic diversity (Borlase et al. 1993). Schaal and Leverich (Chapter 13, this volume) point out that the production of individuals of low fitness by inbreeding or outbreeding depression might be accommodated by excess reproductive capacity. A population may be able to tolerate moderate amounts
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of inbreeding or outbreeding depression without harming overall population fitness because plants exhibit a range of fitnesses. If the population is large enough and producing more offspring than the habitat can support, the less fit individuals will be selected out with little net effect on the population. Recommendation: Maintain large effective population sizes. In the ex situ living collection, mimic natural pollen flow patterns. Because pollinators may not be present, use hand pollination to promote outcrossing.
Outbreeding Outcrossing between plants that are very different genetically, such as plants from distant populations, can lead to a reduction in fitness known as outbreeding depression. Outbreeding depression may be a consequence of local adaptation in the sense that crosses between plants adapted to different local conditions may yield offspring that are poorly adapted to both of the parental environments (Price and Waser 1979; Schierup and Christiansen 1996). Another cause of outbreeding depression is the disruption of coadapted gene complexes. When this occurs, first-generation offspring may demonstrate hybrid vigor, and the major effect of outbreeding depression is not observed until the second or later generations (Templeton 1986; Schierup and Christiansen 1996). Outbreeding depression in plants has not been studied to the extent of inbreeding depression, although it has been demonstrated in a number of species (Waser and Price 1989, 1991; Parker 1992; Fenster and Galloway 2000; Montalvo and Ellstrand 2001). The spatial scale over which outbreeding depression can occur is highly variable. Outbreeding depression has been detected in plants separated by 100 m (Waser and Price 1991), but other studies have not detected it unless plants were separated by 1,000 km (Fenster and Galloway 2000). It is possible that outbreeding depression may be most likely to occur when individuals from different evolutionary lineages are crossed.
Hybridization Hybridization is, in a sense, the most extreme example of outbreeding, and as such can lead to outbreeding depression or genetic assimilation of a population. As discussed in Chapter 16, spontaneous hybridization in ex situ facilities can undermine the genetic integrity of collections and potentially
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contaminate open-pollinated seed or seedlings destined for reintroduction. Similarly, hybridization can occur between plants in ex situ collections and adjacent wild populations, leading to contamination of the ex situ collection or contributing to genetic assimilation and extinction of the natural population. Hybridization may occur directly between wild and cultivated material or indirectly via naturalized escapes that can provide important stepping-stones between ex situ collections and surrounding natural and seminatural habitats. Hybridization, which is promoted by habitat disturbance, unspecialized pollinators, and weak crossing barriers (Levin et al. 1996), can be especially prevalent in gardens where numerous taxa are brought into artificial sympatry.
Managing Outbreeding Depression and Hybridization In nursery situations, these related phenomena are avoided simply by preventing crossing between individuals from different populations or different species. Whenever possible, keep conservation and display collections well separated. Assess origins of all founder stocks, and pay particular attention to seed-derived founders from unmanaged ex situ collections where hybridization may have occurred (this can be determined by genetic screening if necessary). Avoid artificial sympatry (i.e., growing closely related species or individuals from different populations of the same species at the same time and in close proximity). If it is necessary to have simultaneous grow-outs, bag and hand-pollinate flowers to produce seed. In botanic gardens that also manage natural or seminatural areas, care should be taken to avoid hybridization between cultivars in display areas and native species, particularly those of conservation concern, in natural areas. Recommendation: Prevent crossing between plants of different species and, in most cases, between individuals from different populations, unless there is a compelling reason to do so (e.g., a population is reduced to a single incompatibility genotype and augmenting diversity will restore reproductive viability, or a population is suffering from severe inbreeding depression).
Additional Best Practices for Collection Maintenance It is recommended that basic phytosanitary precautions be followed at all times. Keep taxa of conservation concern as isolated as possible from the
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remainder of the living collection. Because any plant entering the collection is a potential source of disease, quarantine new arrivals to the permanent collection and avoid contact with rapidly changing display collections where quarantine is not practical. Minimize the chance of diseases moving between plants by cleaning equipment, such as pruners, and keeping leaf-chewing insects in check. Reliable provenance information is of utmost importance for conservation collections. Maintain impeccable records, carefully and securely label living plants (e.g., put one label in pot or soil and another attached to plant), record horticultural protocols, and follow industry standards for collection documentation. Recommendation: Follow best management practices to reduce the probability of disease spread and to document the collection. Track each individual in propagation breeding programs.
Conclusions The management decisions faced by the curator of an ex situ living collection can be daunting. How large should an ex situ population be? How long can one expect to maintain plants ex situ? Is it more important to equalize family size and relax selection in order to increase population size to combat drift and artificial selection, or is it more important to place plants in a selective arena to eliminate mutations? The first step to resolving these questions is to consider the goals of a given ex situ program and the timescale over which each genetic risk can cause significant fitness declines. Both drift and mutation accumulation depend on effective population size and become most problematic after 25–100 generations, according to models. Depending on its strength, selection can act much more quickly, in just a few generations. For programs with a primary goal of producing plants or seeds for reintroduction, the effects of drift and mutation accumulation can be virtually eliminated by maintaining large effective population sizes and keeping accessions ex situ for only one or a few generations. Minimizing generations in cultivation, providing periodic immigration from wild populations and naturalistic grow-out conditions, can help prevent artificial selection. Some of the inter-situ techniques (high-diversity reserves and naturalistic cultivation) may also prove particularly valuable in minimizing artificial selection. If similar management techniques are used, we can also expect to bank seeds for short- to mediumterm storage without major genetic risks. As seed storage technologies
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improve, the length of regeneration will increase, allowing storage to potentially extend for several hundred years. For programs that have a goal of long-term maintenance, or maintenance for perpetuity, one cannot realistically expect to maintain a population of a wild taxon ex situ for 25, 50, or 100 generations without genetic change. However, the risk of genetic change may be acceptable given the alternatives. For the most threatened taxa, long-term ex situ maintenance may be the only way to prevent extinction, even though it may be an indefinite, perhaps permanent refuge. Some authors have suggested that because genetic changes in ex situ collections are likely after several generations, ex situ methods should be used cautiously, or not at all (Hamilton 1994; Schoen et al. 1998). We interpret these same data more optimistically and are encouraged that plants can be kept ex situ for a few generations without severe dysgenic effects. Ex situ programs represent an additional cadre of tools that conservationists can use to promote species recovery. The garden can serve as short- to medium-term shelter where numbers of individuals can be increased before release into appropriate habitats. It may also have to serve as a last refugium for the most highly threatened species.
Acknowledgments The authors thank Marlin Bowles, R. Stephen Howard, Joyce Maschinski, and Stuart Wagenius for their thoughtful reviews of various drafts of this manuscript. We also thank the staff of the Institute for Plant Conservation at the Chicago Botanic Garden and the Conservation Officers and Scientific Advisory Groups of the Center for Plant Conservation for productive and helpful discussions. References Allendorf, F. W. 1993. Delay of adaptation to captive breeding by equalizing family size. Conservation Biology 7:416–419. Antonovics, J. 1971. The effects of a heterogeneous environment on the genetics of natural populations. American Scientist 59:593–599. Baskin, C. C., and J. M. Baskin. 1998. Seeds: Ecology, Biogeography, and Evolution of Dormancy and Germination. New York: Academic Press. Borlase, S. C., D. A. Loebel, R. Frankham, R. K. Nurthen, D. A. Briscoe, and G. E. Daggard. 1993. Modeling problems in conservation genetics using captive Drosophila populations: consequences of equalization of family sizes. Conservation Biology 7:122–131.
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Galen, C. 1989. Measuring pollinator-mediated selection on morphometric floral traits: bumblebees and the alpine sky pilot, Polemonium viscosum. Evolution 43:882–890. Given, D. R. 1984. Principles and Practice of Plant Conservation. Portland, OR: Timber Press. Hamilton, M. B. 1994. Ex situ conservation of wild plant species: time to reassess the genetic assumptions and implications of seed banks. Conservation Biology 8:39–49. Hamrick, J. L., and L. R. Holden. 1979. Influence of microhabitat heterogeneity on gene frequency distribution and gametic phase disequilibrium in Avena barbata. Evolution 33:521–533. Harada, Y., M. Yokota, and M. Iizuka. 1998. Genetic risk of domestication in artificial fish stocking and its possible reduction. Research in Population Ecology 40:311–324. Hedrick, P. W. 1994. Purging inbreeding depression and the probability of extinction: full-sib mating. Heredity 73:363–372. Hedrick, P. W., and S. T. Kalinowski. 2000. Inbreeding depression in conservation biology. Annual Review of Ecology and Systematics 31:139–162. Heywood, V. 1993. The role of botanic gardens and arboreta in the ex-situ conservation of wild plants. Opera Botanica 121:309–312. Lacy, R. C. 1987. Loss of genetic diversity from managed populations: interacting effects of drift, mutation, immigration, selection, and population subdivision. Conservation Biology 1:143–158. Lande, R. 1994. Risk of population extinction from fixation of new deleterious mutations. Evolution 48:1460–1469. Lande, R. 1995. Mutation and conservation. Conservation Biology 9:782–791. LeBaron, H. M. 1991. Distribution and seriousness of herbicide-resistant weed infestations world-wide. Pages 27–43 in J. C. Casley, G. W. Cussans, and A. R. Atkin (eds.), Herbicide Resistance in Weeds and Crops. Oxford, UK: Butterworth-Heinemann. Levin, D. A., J. Francisco-Ortega, and R. K. Jansen. 1996. Hybridization and the extinction of rare plant species. Conservation Biology 10:10–16. Linhart, Y. B. 1988. Intra-populational differentiation in annual plants. III. The contrasting effects of intra- and inter-specific competition. Evolution 42:1047–1064. Linhart, Y. B., and M. C. Grant. 1996. Evolutionary significance of local genetic differentiation in plants. Annual Review of Ecology and Systematics 27:237–277. Lynch, M., J. Conery, and R. Bürger. 1995a. Mutation accumulation and the extinction of small populations. American Naturalist 146:489–518. Lynch, M., J. Conery, and R. Bürger. 1995b. Mutational meltdowns in sexual populations. Evolution 49:1067–1080. Maunder, M. 1992. Plant reintroduction: an overview. Biodiversity and Conservation 1:51–61. Maunder, M. 1994. Botanic gardens: future challenges and responsibilities. Biodiversity and Conservation 3:97–103. Montalvo, A. M., and N. C. Ellstrand. 2001. Nonlocal transplantation and outbreeding depression in the subshrub Lotus scoparius (Fabaceae). American Journal of Botany 88:258–269. Newman, D., and D. Pilson. 1997. Increased probability of extinction due to decreased genetic effective population size: experimental populations of Clarkia pulchella. Evolution 51:354–362.
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Newman, D., and D. A. Tallmon. 2001. Experimental evidence for beneficial fitness effects of gene flow in recently isolated populations. Conservation Biology 15:1054–1063. Parker, M. A. 1992. Outbreeding depression in a selfing annual. Evolution 46:837–841. Pavlik, B. M. 1994. Demographic monitoring and the recovery of endangered plants. Pages 322–350 in M. L. Bowles and C. J. Whelan (eds.), Restoration of Endangered Species: Conceptual Issues, Planning, and Implementation. New York: Cambridge University Press. Price, M. V., and N. M. Waser. 1979. Pollen dispersal and optimal outcrossing in Delphinium nelsonii. Nature 277:294–297. Schierup, M. H., and F. B. Christiansen. 1996. Inbreeding depression and outbreeding depression in plants. Heredity 77:461–468. Schoen, D. J., and A. H. D. Brown. 2001. The conservation of wild plant species in seed banks. BioScience 51:960–966. Schoen, D. J., J. L. David, and T. M. Bataillon. 1998. Deleterious mutation accumulation and the regeneration of genetic resources. Proceedings of the National Academy of Sciences of the United States of America 95:394–399. Shabalina, S. A., L. Yampolsky, and A. S. Kondrashov. 1997. Rapid decline of fitness in panmictic Drosophila populations under relaxed selection. Proceedings of the National Academy of Sciences of the United States of America 94:13034–13039. Snaydon, R. W., and M. S. Davies. 1976. Rapid population differentiation in a mosaic environment. II. Morphological variation in Anthoxanthum oderatum. Evolution 36:289–297. Snaydon, R. W., and M. S. Davies. 1982. Rapid divergence of plant populations in response to recent changes in soil conditions. Evolution 36:289–297. Spurway, H. 1955. The causes of domestication: an attempt to integrate some ideas of Konrad Lorenz with evolutionary theory. Journal of Genetics 53:325–362. Templeton, A. R. 1986. Coadaptation and outbreeding depression. Pages 105–116 in M. E. Soulé (ed.), Conservation Biology: The Science of Scarcity and Diversity. Sunderland, MA: Sinauer. Templeton, A. R., and B. Read. 1983. The elimination of inbreeding depression in a captive herd of Speke’s gazelle. Pages 241–261 in C. M. Schonewald, S. M. Chambers, B. MacBryde, and L. Thomas (eds.), Genetics and Conservation. Menlo Park, CA: Benjamin-Cummings. Teramura, A. H., and B. R. Strain. 1979. Localized populational differences in the photosynthetic response to temperature and irradiance in Plantago lanceolata. Canadian Journal of Botany 57:2559–2563. Waser, N. M., and M. V. Price. 1989. Optimal outcrossing in Ipomopsis aggregata: seed set and offspring fitness. Evolution 43:1097–1109. Waser, N. M., and M. V. Price. 1991. Outcrossing distance effects in Delphinium nelsonii: pollen loads, pollen tubes, and seed set. Ecology 72:171–179.
appendix 4
Ex Situ Plant Conservation Organizations and Networks Kevin M. James
Off-site germplasm storage can provide a vital buffer against the loss of biological diversity in wild populations. A wide range of agencies and institutions across the globe are committed to integrated conservation plans that include off-site, or ex situ, components. Whether ex situ strategies focus on conserving crops, ornamentals, plants with cultural or historical significance, or threatened native plants, the collections are much more valuable than mere static archives. Germplasm acquisition and storage can offer security against extinction and genetic erosion. In addition it can provide material for recovery activities including reintroduction, where appropriate, and can support the protection of natural populations. Furthermore, integrated conservation programs allow partnerships and networks to develop, which leads to an effective division of labor between institutions. The wide-ranging threats facing the world’s flora are reflected in the varied missions, goals, and programs of the groups dedicated to plant conservation. Groups are involved in ex situ conservation for economic, academic, cultural, or educational reasons, but they are all bound by the desire to preserve the heritage and integrity of the world’s flora. The contributions that agencies and organizations make with ex situ programs range from preserving a single individual or genotype to entire floras. Although ex situ collections are established for different reasons, they all revolve around a set of fundamental challenges: effectively collecting germplasm from the wild, maintaining the necessary levels of genetic diversity ex situ, and making the collections available for conservation use. Meeting these challenges entails different techniques for storing seeds, tissue cultures, live plants, and pollen (see Chapter 1, this volume). For some agencies and organizations, advancing these methods is the main focus of research, whereas for others it is only a step toward achieving a particular conservation goal. 474
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International Ex Situ Plant Conservation Organizations Numerous organizations around the globe are involved in ex situ plant conservation. The following list is far from comprehensive, but it illustrates the diversity of types, sizes, and missions of ex situ plant conservation organizations.
Botanic Gardens Conservation International Botanic Gardens Conservation International (BGCI) was founded in 1987, but its roots date back to the 1976 Threatened Plants Committee conference. This conference requested that a Botanic Gardens Conservation Coordinating Body be established; over the ensuing years, conferences, and acronyms, BGCI emerged with the mission to create a global network of botanic gardens working toward effective plant conservation. Of the 1,600 botanic gardens in the world, more than 500, across 100 countries, are members of BGCI. Together, these members work to implement the Botanic Gardens Conservation Strategy for Plant Conservation and the Global Strategy for Plant Conservation. BGCI places great emphasis on the continued professional development of ex situ conservation for threatened plants and further development of professional capacity in underresourced areas (BGCI 2001).
International Plant Genetic Resources Institute Implementing ex situ conservation strategies and technologies for developing nations is the primary goal of the International Plant Genetic Resources Institute (IPGRI). Founded in 1974, IPGRI grew out of the Consultative Group on International Agricultural Research (CIGAR) and later the International Board for Plant Genetic Resources (IBPGR). Today IPGRI has a staff of more than 170 and offices in 15 countries and is the largest international institute dedicated solely to the conservation and use of plant genetic resources. Through worldwide collaborations with developing nations, plant genetic resource work is undertaken to help secure plant genetic resources, eliminate poverty, increase food security, and protect the environment. The IPGRI is funded primarily by developed-country donor and development agencies, along with a growing number of less developed countries. Through research, publications, training, and assistance on policy and legal matters, IPGRI conducts and directs ex situ conservation on a global scale. It has supported more than 550 germplasm collection trips in 136 countries, has trained more than 2,000 national scientists, and supports the
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establishment of numerous national gene banks. In 1996, IPGRI collaborated with the United Nations Food and Agriculture Organization (FAO) to produce a report on the state of the world’s plant genetic resources for the International Technical Conference on Plant Genetic Resources. At this conference, a Global Plan of Action was adopted along with the Leipzig Declaration. The adoption of the Leipzig Declaration commits countries to implement the priority activities of the Global Plan of Action.
World Conservation Union The World Conservation Union (International Union for the Conservation of Nature, IUCN: http://www.iucn.org/about/index.htm) “seeks to influence, encourage and assist societies throughout the world to conserve the integrity and diversity of nature and to ensure that any use of natural resources is equitable and ecologically sustainable.” A central secretariat coordinates the IUCN program, and through its six commissions, IUCN mobilizes more than 10,000 volunteers who focus on species and biodiversity conservation and the management of habitats and natural resources. Formed more than 50 years ago, the Species Survival Commission (SSC) continues to play a fundamental role in ex situ conservation. The SSC is the largest of the six IUCN commissions and advises the IUCN on all aspects of species conservation. The SSC is a network of nearly 7,000 volunteer members from all over the world. Within the SSC there are 120 specialist groups and task forces that focus on a particular group of at-risk organisms or regions. Moreover, the SSC produces the IUCN Red List of Threatened Species, along with action plans, policy guidelines, and newsletters regarding species conservation. This commission promotes an integrated plant conservation philosophy that brings together in situ and ex situ conservation organizations under a global network. The Reintroduction Specialist Group and the Conservation Breeding Specialist Group have played a major role in developing SSC ex situ policy and guidelines for reintroduction and ex situ conservation (Box A4.1).
International Seed Banks and Ex Situ Projects Many gene banks began seed-banking programs to offset the loss of genetic diversity in crops and crop relatives. The methods and techniques developed in these programs have been used to establish ex situ collections of threatened wild plants as a backup to on-site conservation.
box a4.1 IUCN World Conservation Union Guidelines for the Management of Ex Situ Populations • The basis for responsible ex situ population management in support
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of conservation is founded on benefits for both threatened taxa and associated habitats. The primary objective of maintaining ex situ populations is to help support the conservation of a threatened taxon, its genetic diversity, and its habitat. Ex situ programs should give added value to other complementary programs for conservation. Although there will be taxon-specific exceptions because of unique life histories, the decision to initiate ex situ programs should be based on one or more of the appropriate IUCN Red List criteria, including (1) when the taxon or population is prone to effects of human activities or stochastic events or (2) when the taxon or population is likely to become Critically Endangered, Extinct in the Wild, or Extinct in a very short time. Additional criteria may need to be considered when taxa or populations of cultural importance and significant economic or scientific importance are threatened. All Critically Endangered and Extinct in the Wild taxa should be subject to ex situ management to ensure recovery of wild populations. Ex situ conservation should be initiated only when an understanding of the target taxon’s biology and ex situ management and storage needs is such that there is a reasonable probability of enhancing species conservation or when the development of such protocols could be achieved within the time frame of the taxon’s needed conservation management, ideally before the taxon becomes threatened in the wild. Ex situ institutions are strongly urged to develop ex situ protocols before any forthcoming ex situ management. Consideration must be given to institutional viability before a long-term ex situ project is initiated. For threatened taxa for which husbandry or cultivation protocols do not exist, surrogates of closely related taxa can serve important functions, such as in research and protocol development, conservation biology research, staff training, public education, and fundraising. Although some ex situ populations may have been established before the ratification of the Convention on Biological Diversity (CBD), all ex situ and in situ populations should be managed in an integrated, multidisciplinary manner and, where possible, in accordance with the principles and provisions of the CBD.
• Extreme and desperate situations, in which taxa or populations are in
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imminent risk of extinction, must be dealt with on an emergency basis. This action must be implemented with the full consent and support of the range nation. All ex situ populations must be managed so as to reduce risk of loss through natural catastrophe, disease, or political upheaval. Safeguards include effective quarantine procedures, disease and pathogen monitoring, duplication of stored germplasm samples in different locations, and provision of emergency power supplies to support collection needs (e.g., climate control for long-term germplasm repositories). All ex situ populations should be managed so as to reduce the risk of invasive escape from propagation, display, and research facilities. Taxa should be assessed as to their invasive potential and appropriate controls taken to avoid escape and subsequent naturalization. The management of ex situ populations must minimize any deleterious effects of ex situ management, such as loss of genetic diversity, artificial selection, pathogen transfer, and hybridization, in the interest of maintaining the genetic integrity and viability of such material. Particular attention should be paid to initial sampling techniques, which should be designed to capture as much wild genetic variability as practicable. Ex situ practitioners should adhere to, and further develop, any taxon- or region-specific recordkeeping and genetic management guidelines produced by ex situ management agencies. Those responsible for managing ex situ populations and facilities should seek to increase public awareness, concern, and support for biodiversity and to support the implementation of conservation management through education, fundraising, professional capacitybuilding programs, and support for direct action in situ. Where appropriate, data and the results of research derived from ex situ collections and ex situ methods should be made freely available to ongoing in-country management programs concerned with supporting conservation of in situ populations, their habitats, and the ecosystems and landscapes in which they occur.
Note: Ex situ conservation is defined here, as in the CBD, as “the conservation of components of biological diversity outside their natural habitats.” Ex situ collections include whole plant or animal collections, zoological parks and botanic gardens, wildlife research facilities, and germplasm collections of wild and domesticated taxa (zygotes, gametes, and somatic tissue).
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Millennium Seed Bank Project The Millennium Seed Bank Project (MSBP) was developed to safeguard 10 percent (about 24,000) of the world’s dryland seed-bearing flora by 2010, with an additional emphasis on the native flowering plants of the United Kingdom. In addition to these two principal goals, the project seeks to improve all aspects of seed conservation, including making seed available for research and reintroduction and furthering seed conservation technology. The MSBP is part of the Royal Botanical Garden, Kew, and is housed at Wakehurst Place in Sussex. The project expanded on the Kew Seed Bank, founded in 1974, and continues to conduct research on seed storage and germination. Building on the spirit of the 1992 Convention on Biological Diversity (CBD), MSBP promotes access and exchange of seed conservation technology. This technology is used to secure biological diversity and promote sustainable development.
National Center for Genetic Resources Preservation In the United States, the National Seed Storage Laboratory (NSSL) and National Plant Germplasm System (NPGS) conduct ex situ conservation of crop germplasm. Both groups are part of the Department of Agriculture, Agriculture Research Services. The National Center for Genetic Resources Preservation (NCGRP) oversees the NSSL, and both are located in Fort Collins, Colorado. The NSSL is the largest single gene bank in the world, storing more than 8,000 species and conserving millions of individual plants via seed. Within NCGRP, ex situ plant conservation is divided between the Seed Viability and Storage Research Unit and the Plant Germplasm Preservation Research Unit. These two units are charged with long-term storage of seeds and maintenance of a secure supply of plant germplasm, respectively, and with furthering ex situ technology and methods. Preserving the genetic diversity of agronomically important plants is the sole mission of the NPGS. This system is a cooperative effort by state and federal bodies and private organizations. It is designed to give scientists access to crop material for research by acquiring, preserving, evaluating, documenting, and distributing genetically diverse germplasm. Designed to counter the genetic uniformity that has resulted from modern agriculture and marketing, NPGS collects genetically diverse crop germplasm from around the world and maintains the stored material at more than 30 germplasm repositories across the United States. This collection of nearly 500,000 accessions is overseen by the Crop Germplasm Committee
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(CGC) and is made available for research purposes. Base seed collections of NPGS are stored at the NSSL. The NPGS via a memorandum of understanding with the Center for Plant Conservation (CPC) holds seed samples of threatened wild plant species on behalf of CPC gardens.
National and Regional Ex Situ Plant Conservation Networks Many organizations that are involved in plant conservation, including botanic gardens and arboreta, have formed networks to facilitate integration between in situ and ex situ practitioners, fundraising, education, and outreach programs and to avoid duplication of efforts. The following are some of the larger plant conservation networks from around the world.
Australian Network for Plant Conservation Starting as a grassroots organization in 1991, the Australian Network for Plant Conservation (ANPC) has as its major aim the integration of all approaches to plant conservation. Recognizing the need for coordination between in situ and ex situ programs, the ANPC joins together more than 400 members from state and federal agencies, botanic gardens, zoos, native plant enthusiasts, and industry groups. The ANPC promotes and develops germplasm conservation through training, conferences, publications, and the National Endangered Flora Collection (NEFC). This collection not only catalogs the threatened flora being held as seed, whole plants, or other types of germplasm by each member but also is an important source of propagation material for research, education, and recovery projects. Using propagated material from the collection helps reduce the amount of plant material collected from the wild. To this end, collection maintenance guidelines have been published along with a catalog of NEFC listings.
Canadian Botanical Conservation Network Dedicated to preserving Canada’s biological diversity of rare and endangered native plant species, wild habitats, and ecosystems, the Canadian Botanical Conservation Network (CBCN) assists the professional botany community with all aspects of plant conservation in Canada. In addition to leading cooperative research projects, CBCN develops educational material to raise public awareness about the threats to native plants and their habitats. Like most
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current integrated plant conservation programs, the CBCN benefited greatly from the Convention on Biological Diversity. Furthering Canada’s commitment to the CBD process, the Canadian Biodiversity Strategy (CBS) was established. Out of this strategy emerged CBCN as a national system to organize and unite ex situ conservation programs across botanic gardens and arboreta. The network operates as a clearinghouse for resource material and information to assist ex situ collections and research.
Center for Plant Conservation One of the first networks created with the sole purpose of conserving and saving rare and threatened plants is the CPC. Established in 1984, the CPC is the leader in promoting conservation science in botanical gardens in the United States. Today, the CPC is composed of 33 participating botanic gardens and arboreta and is a nonprofit organization with a national office hosted by the Missouri Botanical Garden in St. Louis, Missouri. The national office coordinates conservation programs, sets policy, raises funds, and informs government officials and the general public about plant conservation priorities. The National Collection of Rare and Endangered Plants, launched by the CPC, is one of the largest living collections of threatened plants in the world. This collection is a collaboration of ex situ methods across participating institutions and state and federal government agencies. The collection is composed of federally listed threatened and endangered plant species stored as seeds or maintained as whole plants. Through a memorandum of understanding, most seeds are stored at the NSSL. Additional seeds are banked at participating institutions and used for research and restoration projects. From 1992 to 2000, the CPC operated a satellite field office in Honolulu, Hawaii. This office was designed to coordinate efforts of the five CPC institutions in the state. Benefiting from the support of the Bishop Museum, CPC-Hawaii’s chief mission was to bring the most threatened species into the National Collection of Endangered Plants quickly and to help coordinate the activities of Hawaii plant conservation statewide.
Native Seeds/SEARCH Long-term seed storage is not limited to just scientific or government agencies but extends to nonprofit groups preserving both crop and cultural diversity. Based in Tucson, Arizona, Native Seeds/Southwestern Endangered Aridlands Resources Clearinghouse (Native Seeds/SEARCH, or NS/S) is a
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nonprofit organization preserving Native American crop seeds and their wild relatives in the southwestern United States and northwestern Mexico. NS/S was founded in 1983 as a result of requests from Native Americans on the Tohono O’odham reservation near Tucson who could not locate seeds of traditional crops. NS/S stores and distributes seeds while promoting their uses in traditional communities. Educational outreach and training sessions are conducted to teach the history of these crops and safeguard their cultural heritage along with the seeds. The main benefit of NS/S is the conservation of locally adapted seed types. Preserving this level of fine-scale genetic diversity helps offset the genetic uniformity found in modern agriculture and ensures material for traditional farming at the local level.
International Ex Situ Policy: The Global Strategy for Plant Conservation In early 2002, delegates at the sixth Conference to the Parties to the CBD (COP6-CBD) adopted the Global Strategy for Plant Conservation. The decision is a landmark in that it is the first time that plant conservation issues have received such detailed scrutiny by the governments of the world (183 countries are parties to the CBD). It is also the first time that targets, albeit voluntary, have been set to guide and monitor progress by the CBD (Box A4.2). If the strategy is to reach its targets by 2010, every country will need to make major commitments to plant conservation. If necessary, the targets will be reviewed and amended at future conferences. Sustainable use, capacity building, education and awareness, increased networking, and community involvement are vital elements of the strategy if it is to achieve lasting results.
Conclusions One very important aspect of ex situ plant conservation that is carried out by every agency or organization described is the management and dissemination of data and information. The maintenance of scientific collections extends beyond the physical collection, storage, and propagation of material to include the tracking of information associated with each accession. Each group mentioned in this appendix is actively involved in maintaining a searchable database that, in some cases, can be accessed remotely via the Internet. Working within and across networks to achieve a common goal greatly benefits from the efficient exchange of current and accurate infor-
box a4.2 Global Strategy for Plant Conservation Targets Adopted by the Sixth Conference to the Parties to the Convention on Biological Diversity Target 1: A widely accessible working list of known plant species, as a step towards a complete world flora. Target 2: A preliminary assessment of the conservation status of all known plant species, at national, regional and international levels. Target 3: Development of models with protocols for plant conservation and sustainable use, based on research and practical experience. Target 4: At least 10 percent of each of the world’s ecological regions effectively conserved. Target 5: Protection of 50 percent of the most important areas for plant diversity assured. Target 6: At least 30 percent of production lands managed consistent with the conservation of plant diversity. Target 7: 60 percent of the world’s threatened species conserved in situ. Target 8: 60 percent of threatened plant species in accessible ex situ collections, preferably in the country of origin, and 10 percent of them included in recovery and restoration programmes. Target 9: 70 percent of the genetic diversity of crops and other major socio-economically valuable plant species conserved, and associated local and indigenous knowledge maintained. Target 10: Management plans in place for at least 100 major alien species that threaten plants, plant communities and associated habitats and ecosystems. Target 11: No species of wild flora endangered by international trade. Target 12: 30 percent of plant-based products derived from sources that are sustainably managed. Target 13: The decline of plant resources, and associated local and indigenous knowledge innovations and practices, that support sustainable livelihoods, local food security and health care, halted. Target 14: The importance of plant diversity and the need for its conservation incorporated into communication, education and public awareness programmes. Target 15: The number of trained people working with appropriate facilities in plant conservation increased, according to national needs, to achieve the targets of this strategy. Target 16: Networks for plant conservation activities established or strengthened at national, regional and international levels.
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mation. Up-to-date accession information reduces duplication of collection and research efforts while maximizing the utility of the entire collection. Ex situ conservation programs face similar obstacles and share many successes. They are designed to preserve a specific botanical target over a particular time frame. Whatever the species under consideration, the accepted methods, or the new technologies a strategic plan entails, all programs rely on each other. It is apparent from these brief overviews of organizations involved in ex situ plant conservation that networks and collaborations are the key to successfully integrated plant conservation strategies at any scale.
Acknowledgments The authors thank Kayri Havens, Mike Maunder, and Ed Guerrant for reviews and Vickie Caraway (Department of Land and Natural Resources, Hawaii), Ehsan Dulloo (IPGRI-Rome), Florent Engelmann (IPGRI-Rome), David Galbraith (CBCN), Peter Wyse Jackson (BGCI), Jeanette Mill (ANPC), Suzanne Nelson (Native Seeds/SEARCH), Roger Smith (Royal Botanic Gardens, Kew), and Christina Walters (U.S. Department of Agriculture) for their contributions to this appendix and symposium presentations. Reference BGCI (Botanic Gardens Conservation International). 2001. Botanic Garden Agenda for Conservation. London: Botanic Gardens Conservation International.
about the contributors
Volume Editors Edward O. Guerrant Jr. holds a B.S. in botany and zoology from the University of Washington, an M.A. in biology from Sonoma State College, and a Ph.D. in botany from the University of California at Berkeley, where he studied evolutionary ecology. Since 1989, he has been conservation director at the Berry Botanic Garden, where he oversees an active program that includes the Seed Bank for Rare and Endangered Plants of the Pacific Northwest. His research interests focus on demography and reintroduction of rare and endangered plant species. Kayri Havens holds a B.S. and an M.A. in botany from Southern Illinois University and a Ph.D. in biology from Indiana University, where she studied reproductive success in a rare evening primrose (Oenothera organensis) from New Mexico. She spent 3 years as the conservation biologist at Missouri Botanical Garden before joining the Chicago Botanic Garden in April 1997. She is currently the garden’s director of the Institute for Plant Conservation. Her research interests include restoration genetics and the biology of plant rarity and invasiveness. She also collaborates with a variety of academic institutions and stewardship organizations to help improve conservation efforts for rare plants. Mike Maunder holds a diploma in horticulture from the Royal Botanic Gardens, Kew, and an M.Sc. and Ph.D. from the University of Reading, where he studied the ex situ management of threatened plant species. He was head of the Conservation Projects Development Unit at the Royal Botanic Gardens, Kew, before joining the National Tropical Botanical Gar485
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den in Hawaii as director of conservation in 2000. He is currently director of the Fairchild Tropical Garden, Miami. Mike is an active member of the World Conservation Union Species Survival Commission (IUCN/SSC) Plant Conservation Committee and Deputy Chair of the SSC Reintroduction Specialist Group.
Chapter Authors Carol C. Baskin is a professor of biology and agronomy at the University of Kentucky. Her research, in collaboration with Jerry M. Baskin, focuses on ecological life history of plant species, with particular reference to seed germination. They have studied the germination ecology and ecophysiology of species from different habitats and biomes and with different life cycles, life forms, and phylogenetic relationships. Jerry M. Baskin is a professor of biology at the University of Kentucky and conducts a joint research program with Carol C. Baskin. Lesley G. Campbell completed her M.Sc. at the University of Guelph, Canada, and is currently a doctoral candidate in the Department of Biology at Ohio State University, studying the effects of crop gene introgression on the evolution of weeds. Anne Cochrane is a research scientist with the Science Division of the Department of Conservation and Land Management, Western Australia. She manages the Threatened Flora Seed Centre, a long-term storage facility for seed of rare and threatened species of Western Australia. She is also involved in research into new techniques for the collection and storage of seed and plant-animal interactions related to seed set, seed predation, and the productivity of the species. Alastair Culham is a lecturer at the School of Plant Sciences, University of Reading, with a special focus on evolutionary studies of vascular plants (Actaea, Cyclamen, Drosera, and Perlagonium), conservation biology, and horticultural taxonomy. Kingsley W. Dixon is the director of science at Kings Park and Botanic Garden in Perth, Western Australia. His research interests include conservation genetics and tissue culture propagation of native and rare Australian plants.
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Donald A. Falk studies forest ecology in the Department of Ecology and Evolutionary Biology and the Laboratory of Tree-Ring Research at the University of Arizona. Don served as executive director of the Society for Ecological Restoration (SER) from 1993 to 2000. From 1984 to 1993, Falk cofounded and directed the Center for Plant Conservation. He currently serves as associate editor of the SER–Island Press book series Science and Practice of Ecological Restoration. Peggy L. Fiedler was a professor and director of the graduate program in conservation biology at San Francisco State University. She currently is a consultant with L. C. Lee and Associates in California. Samara Hamzé was a research assistant at Archbold Biological Station in Florida. Currently she is an instructor in the Department of Biology and Microbiology at the University of Wisconsin at Oshkosh. Julie A. Hawkins is a lecturer in botany at the Department of Plant Sciences at the University of Reading, where she studies the systematics of flowering plants, principally Leguminosae, species relationships using morphological and molecular data sets, and molecular, ecological, and geographic investigation of hybridization. Colin Hughes is a researcher at the School of Plant Sciences, University of Oxford, and the Royal Society University Research Fellow, specializing in the systematics of the Leguminosae, particularly the genera Leucaena and Lupinus. Brian C. Husband is an associate professor of botany at the University of Guelph, Canada. He is interested in the effects of plant reproductive systems on population persistence, adaptation, and speciation. Kevin M. James was the conservation programs information coordinator for the Center for Plant Conservation in St. Louis, Missouri, and is currently a graduate student. Kathryn L. Kennedy is the president and executive director of the Center for Plant Conservation in St. Louis, Missouri. She is committed to raising awareness about rare plants and improving ex situ and in situ plant conservation practices. Charles Lamoureux (deceased) was the director of the Harold L. Lyon Arboretum and professor of botany at the University of Hawaii at Manoa.
488
about the contributors
He received his B.S. from the University of Rhode Island, his M.S. from the University of Hawaii, and his Ph.D. from the University of California, Davis. Although his professional specialty was plant morphology, he had been interested in conservation of Hawaiian plants since he was a graduate student, and in recent years most of his professional activities were in conservation biology. Wesley J. Leverich is an associate professor of biology at St. Louis University. His research interests include evolutionary biology of plant populations, plant reproductive ecology, fitness measures, and organization of genetic variation in plant populations. Eric S. Menges is senior research biologist at Archbold Biological Station in Florida and a member of the Center for Plant Conservation’s Scientific Advisory Committee. His current research focuses on the demography and life history of plant populations, effects of habitat fragmentation on genetic structure and ecological traits, modeling extinction probability and population viability, life history adaptations of scrub plants to fire, and fire effects on plant population dynamics. Valerie C. Pence is director of plant research at the Cincinnati Zoo and Botanical Garden. Her research interests include in vitro collecting, tissue culture propagation, and cryopreservation of endangered plants, focusing on species that have challenged traditional methods of plant preservation and propagation. She is also a member of the Center for Plant Conservation’s Scientific Advisory Committee. Ghillean T. Prance is scientific director of the Eden Project in Cornwall and visiting professor at Reading University. He was director of the Royal Botanic Gardens, Kew, from 1988 to 1999. His exploration of Amazonia included 15 expeditions in which he collected more than 350 new plant species. He is author of 17 books and has published more than 400 scientific and general papers in taxonomy, ethnobotany, economic botany, conservation, and ecology. He has received numerous awards and honors, including election as a Fellow of the Royal Society, and was knighted in 1995. Hugh W. Pritchard is the head of research at the Seed Conservation Department, Royal Botanic Gardens, Kew. Current research interests include ecological correlates of seed desiccation sensitivity, especially in
about the contributors
489
the palms; seed dormancy models, particularly in relation to heat shock; and the biochemistry and biophysics of seed aging. He is a founding trustee of the International Society for Seed Science and serves on committees for the International Seed Testing Association and the World Conservation Union. Peter H. Raven is director of the Missouri Botanical Garden and George Engelmann Professor of Botany at Washington University in St. Louis. He currently serves as president-elect of Sigma Xi, chair of the Division of Earth and Life Studies in the National Research Council, and chair of the National Geographic Society’s Committee for Research and Exploration. He is a member of more than 20 academies of science in countries around the world and a recipient of the National Medal of Science. He devotes much of his time to advocacy for conservation and sustainable development. Holmes Rolston III is a university distinguished professor and professor of philosophy at Colorado State University. A founder of environmental ethics as a philosophical discipline, he is past and founding president of the International Society for Environmental Ethics. Barbara Schaal is the Spencer T. Olin Professor in the Department of Biology at Washington University, St. Louis, Missouri. She uses molecular genetics techniques to study evolution of plants, especially native species. She serves as chair of the Scientific Advisory Council of the Center for Plant Conservation and as a member of the board of trustees, Missouri chapter of the Nature Conservancy. Pritpal S. Soorae is senior conservation officer at the Environmental Research & Wildlife Development Agency, Abu Dhabi, United Arab Emirates, and coordinator for the IUCN/SSC Reintroduction Specialist Group. Mark R. Stanley Price is the former director of Africa operations for the African Wildlife Foundation and chair of the IUCN/SSC Reintroduction Specialist Group. Mark is currently director of the Durrell Wildlife Conservation Trust and active in developing policy on the strategic use of ex situ conservation. Nellie Sugii received her B.S. and M.S. degrees in horticulture science from the University of Hawaii at Manoa. In 1998, Nellie became a junior researcher at the Lyon Arboretum, where she manages the Micropropa-
490
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gation Laboratory and the Hawaiian Rare Plant Project and is the arboretum’s curator of Hawaiian plants. Leigh E. Towill is a plant physiologist with the National Center for Genetic Resources Preservation (formerly National Seed Storage Laboratory) in Fort Collins, Colorado. His research interests include cryopreservation, clonal germplasm, plant tissue culture, seed and pollen storage, and vitrification. Pati Vitt is a conservation scientist at the Chicago Botanic Garden. Her conservation interests include the effects of reproductive biology on population growth and stability, evolutionary ecology of plant invasions, and the effects of community-level management on populations of rare plant species. Christina Walters is research leader of plant germplasm preservation research at the U.S. Department of Agriculture Agricultural Research Service (ARS) National Center for Genetic Resources Preservation (formerly USDA National Seed Storage Laboratory) in Fort Collins, Colorado. Her research interests include seed development and quality, with particular focus on acquisition of desiccation tolerance and longevity, seed storage behavior among plant species, life under extremely dry or cold conditions, and biophysical aspects of water on biological surfaces. She is also a member of the Center for Plant Conservation’s Scientific Advisory Committee.
index
Abscisic acid (ABA), 215–18 Acaena argentea, 349 Accession sizes, 113–14 Adalberto, Dr., xxvii Adaptation/adaptive variation, 255–56, 462 see also Evolutionary principles, identifying genetic/demographic factors by looking at Adiantum tenerum, 218 Aeonium, 338 Aesculis hippocastanum, 151, 152 Aethusa cynapium, 170 Afghanistan, xxvi Africa, 91, 92–93, 103, 395–96 Agave arizonica, 335 Agenda 21, 413 Agriculture and difference between spontaneous nature and deliberated culture, 23 Agriculture Department, U.S. (USDA), 77, 78, 114, 410 Agropyron, 163 Agrostis, 249 American Zoo and Aquarium Association (AZA), 74, 87, 93 Amherstia nobilis, 7 Amorphophallus titanium, 181 Amphibians, 93 Angiosperms, 180, 190 Anthoxanthum, 240, 249, 462 Aquariums/aquaria, 87, 91 Arabis koehleri, 379
Araucaria, 148, 150–52 Argyranthemum, 338, 346 Argyroxiphium, 337 Aristotle, 29 Ark paradigm, 8, 85 Artemia, 120 Asclepias meadii, 270–72, 380 Association of Official Seed Analysts (AOSA), 451 Aster, 343 Astragalus, 407 Attrition, sources of, 432 Australia, 11, 394–95, 404 see also Western Australia’s ex situ program Australian Network for Plant Conservation (ANPC), 9, 47, 404, 408, 480 Auxin, 198 Avena barbata, 462 Ayurvedic medicines, xxvi, xxvii Azadirachta indica, 148 Bananas, 190 Beach habitat, 71 Bees, 342 Bencomia, 338 Berry Botanic Garden (BBG), xviii, 410, 411 Birds, 92, 93 Bishop Museum, 481 Bison, American, 85 Black River George National Park, 97 Bok Tower Gardens, 71
491
492
index
Bolivia, 97 Botanic gardens: concentrated site of species richness, 3 Hawaii, xiv hybridization, 346 International Agenda for Botanic Gardens in Conservation, 405 need, are out-of-country facilities managing species most in, 93 publications dedicated to, lack of, 87 roles of, various, xiv statistics on, 10–11 what and where are the main facilities, 91–92 see also Zoos/botanic gardens, lessons from and opportunities for; individual subject headings Botanic Gardens and Park Authority (BGPA), 49, 55 Botanic Gardens Conservation International (BGCI), 2 Convention on Biological Diversity, 90 global plant conservation, realizing the full potential of, xxv, 405 International Agenda, 87–88, 405 overview, 475 shared activities between botanic gardens, 406 Brassica campestris, 457 Brassolaeliocattleya, 344 Brazil, xxvii, 6, 102, 401, 404 Brighamia insignis, xxiv–xxv, 181 Bristol Zoo, 103 British Red Data Book, 335 Broad-sense heritability, 290–91 Bryophytes/pteridophytes, ex situ conservation methods for: conclusions, 220 field collecting, 218–19 horticultural collections, 207–8 overview, 206–7 spore banking, 208–10 tissues, vegetative banking, 215–18 bryophyte gametophytes, 215–17 pteridophyte gametophytes, 217–18 pteridophyte sporophytes, 218 in vitro ex situ collections, 214, 215 in vitro propagation, 210–14 Bureau of Land Management (BLM), 77, 298, 410, 411
Bushmallow, San Clemente Island, 29–30 Bushmeat, 103 Cactus, 349 California, 342, 394, 399, 410 Cameroon, 103 Canadian Botanical Conservation Network (CBCN), 480–81 Canary Islands, 346, 464 Capsicum, 412 Captive plants and ethical/philosophical concerns, 28–31 Caralluma, 371 Cardamine, 176, 177, 339 Carduus, 253 Caring for the Earth, 88 Ceanothus sanguineus, 164 Center for Plant Conservation (CPC), xxv effective ex situ methods, 387–88 evolution of, 11 Fish and Wildlife Service, U.S., 73–74 genetics, integrating quantitative, 294, 297–300 global plant conservation, realizing the full potential of, 409 Guidelines for Conservation Collections, 419–24 Hawaii, 395 micropropagation, 189–90 National Collection of Endangered Plants, 14–15 overview, 481 propagation of species, revised policy for controlled, 75 sample decline/survival, 365 successful model, xiv–xv, xvii–xix, 8–9 threatened plants per facility, 401 Central America, 91 Ceratiola ericoides, 175 Chaerophyllum tainturieri, 170, 171 Chamerion angustifolium, 250 Chicago Botanic Garden, xviii, 86, 349, 406 China, 16, 91, 404 Chlorophyllous spores, 209 Chrysanthemum leucanthemum, 163 Cirsium pitcheri, 380, 463 Clarkia pulchella, 248, 455 Climate change, 44, 68, 399 Climax forests, 269
index Clones/clonal material, 192–93, 382 see also Genetic listings Coffea, 141, 143, 148, 150 Coffee, 148 Collection, management guidelines for seed, 327–29 see also Guidelines listings; Sample decline during storage/reintroduction; Seed listings Colonialism, conservation, 97 Commercial cultivation, 13, 391, 459, 461 Communal ownership, biodiversity held under, 393 Community garden, 13, 459, 461 Competitive natural selection pressures, 30 Compromise, analyzing the process/pitfalls of, 31–38 Condor, California, 85, 436 Congo, Democratic Republic of, 96–97, 100 Contamination, genetic, 347–49 see also Sample decline during storage/reintroduction Convention on Biological Diversity (CBD): Botanic Gardens Conservation International, 90 colonialism, conservation, 97 global plant conservation, realizing the full potential of, 405 Global Strategy for Plant Conservation, 483, 482 hybridization, 344 legitimization of ex situ conservation, 1, 9, 10 in situ conservation, xxiii–xxiv, 10, 67 zoos, 89–90 Corydalis flavula, 170 Cotinus, 163 Crop Germplasm Committee (CGC), 479–80 Crop plants, framework/guidelines for ex situ conservation of, 233, 237 Cryopreservation, 12, 132, 140, 215–17, 390, 458, 461 Culture, differences between spontaneous nature and deliberated, 21–23 Cupressocyparis leylandii, 345 Cyanea, xxiv, 200 Cycads, 180 Cypripedium, 343 Cytokinin, 198
493
Dactyloleje-unea acanthifolia, 206 Daisy, Lakeside, 181, 292 Danthonia sericea, 316 Dasypyrum villosum, 299 Deer, Père David’s, 85 Defense Department, U.S., 77 Dehydration, seeds’ response to, see Seed listings Delphinium tricorne, 171 Desiccation-tolerant seeds, 115 see also Seed listings Deteriorative reactions, underlying precept of controlling, 114–17 see also Seed listings Developing countries, 15–16, 92, 392–99, 401, 406 Dieback disease, 43 Diplazium laffanianum, 206 Disney Animal Kingdom, 86 DNA (deoxyribonucleic acid), 276, 334, 372–73 see also Genetic listings Dormancy-breaking and germination requirements from fewest seeds: breaking physiological and morphophysiological dormancy, 166–72 conclusions, 177 guidelines for seed storage, 447 identifying dormancy types, 163 family-level dormancy patterns, 163–66 key to general types of seed dormancy, 166 morphological dormancy, 164–65 morphophysiological dormancy, 165–66 physical dormancy, 163–64 physiological dormancy, 165 move-along experiment, 172–77 multiple cues to stimulate germination, 47–48, 51, 55 overview, 162 Dracaena ombet, 8 Dragonflies, 104 Drift, genetic, 268–69, 455, 457, 462 Drosophila, 455 Dune habitat, 71 Durrell Wildlife Conservation Trust, 97, 103, 409 Ebony, Saint Helena, xxviii Echinacea, 349
494
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Echinocactus grusonii, 14 Echium, 338, 343, 346 Ecological constraints and populations’ response to new environment, 251–54 Ecological processes, changes in, 68 Economic use of wild species, linking ex situ conservation with the, 102–3 Ecosystem restoration, 295–97 Ecotypic differentiation, 272–76 Eden Project, xxviii Eichhornia paniculata, 250, 252 Embryos and micropropagation, 197 Embryos and seed dormancy, 168–72, 175 Encapsulation dehydration technique, 216 Encephalartos woodii, 337 Erythrina, 338, 347 Erythronium elegans, 377, 381 Ethical/philosophical concerns: captive plants, 28–31 complementing or undercutting in situ conservation, 31–38 ecosystems, plants in, 26–28 values, plants and intrinsic, 23–26 Ethical/philosophical concerns: natural and the artificial, the, 21–23 Eucalyptus, 338 Eupatorium resinosum, 316 Euphorbia, 338, 343, 345 Europe, 15, 16, 91, 342 European Association of Zoos and Aquaria (EAZA), 92, 103 European Committee for Conservation of Bryophytes, 207 Evolutionary principles, identifying genetic/demographic factors by looking at: adaptive response, constraints to the ecological constraints, 252–54 genetic constraints, 246–51 maladaptation, 252–53 mutations, deleterious, 249–51 physiological acclimation, 254 size and life history, population, 253–254 variance, genetic, 246–49 biology, evolutionary, 232–37 conclusions, 258–60 implications for ex situ conservation, 254–58
new environments, response to long-term population growth and persistence, 241, 244 maladaptation, 237–43 short-term changes in survival and reproduction, 241, 244 overview, 231 Ex situ conservation: concerns about, initial, 68–69 Convention on Biological Diversity, 90 ecological and evolutionary context of, 229–30 evolution of, 7–10 funding issues, 70 genetically representative collections providing benefits, 70–71 legitimization of, xxiii–xxiv, 10 means not an end, conservation as a, 37–38 place for, the, 68–71 public support for, 69–70 risk assessment, 70 scope/potential of, 1–2 structure of, xvii–xviii tools for, 111–12 see also Storage types, classification of seed; Dormancy-breaking and germination requirements from fewest seeds; Pollen storage as a conservation tool; Micropropagation as a tool for genetic conservation; Bryophytes/pteridophytes, ex situ conservation methods for; Germplasm in gene banks, principles for preserving urgent need for, 71–72 see also Guidelines listings; Methods, ex situ; individual subject headings Extinction/threatened species, xiii, 3, 6–7, 42, 45–48, 316, 335, 400–401, 457 see also Guidelines for conservation collections of rare/endangered species, revised genetic; Seed collection and extinction risk of perennial plants F. Proserpinacoides, 177 Fairchild Tropical Garden, 181 Faked nature, 34–35 Fallopia, 351, 352 Federal guidance, role of:
index agencies as leaders/partners in conservation, 76–78 Endangered Species Act of 1973, 72–74 Fish and Wildlife Service, U.S., 73–76 Policy and Guidelines for Planning and Coordinating Recovery of Endangered and Threatened Species, 76 propagation of species, revised policy for controlled, 74–76 state and federal partnerships, 79–81 Ferns, 209, 214 Ferret, black-footed, 85, 436 Fish, 93 Fish and Wildlife Service, U.S. (FWS), 73–76, 410 Flagship/totem species, 100 Floerkea proserpinacoides, 176 Florida, 410 Food and Agriculture Organization (FAO), 413, 476 Forest Service, U.S. (USFS), 76, 410, 411 Foundation for the Revitalization of Local Health Traditions (FRLHT), 412 France, 16, 91 Franklinia, xxviii, 7, 207, 357, 411 Freezing stress and deterioration in biological materials, 115–16 see also Germplasm in gene banks, principles for preserving; Storage types, classification of seed Fritillaria, 343 Fumana procumbens, 316 Funding issues, 70, 408–9 Gaza-Kruger-Gonarezhou Transfrontier Conservation Area, 102 Gazelle, Mhorr, 95 Geese, 22, 23 Genetic issues, population: adaptive response, constraints on the mutations, deleterious, 249–51 variance, genetic, 246–49 conclusions, 281–82 drift, genetic, 268–69 ex situ populations, genetics of, 276–77 gene banks, 13, 69, 391, 459, 461 genetic diversity and assessing effectiveness of ex situ techniques, 14–15
495
geographic provenance of seed sources, 272–76 hard or soft selection, 279–81 inbreeding depression: Lupinus texensis, 277–79 plasticity, 275 small population size: Asclepias meadii, 270–72 source populations, the genetics of, 267–70 see also Guidelines listings; Hybridization Genetics, integrating quantitative: Center for Plant Conservation guidelines for seed collections, 297–300 conclusions, 300 ex situ conservation, 297 heritability of quantitative traits, 288–91 overview, 286–87 quantitative traits, 287–89 redefining genetic diversity and conservation, 291–94 reintroductions/ecosystem restoration, 295–97 Genetic Safety Net Listing (GSNL), 189, 191, 202 Genetic Sampling Guidelines for Conservation Collections of Endangered Plants, 423 Genetics and Conservation of Rare Plants (Falk & Holsinger), xvii, xviii, 286 Geographic provenance of seed sources, 272–76 Germany, 16, 91 Germination requirements, see Dormancybreaking and germination requirements from fewest seeds; Sample decline during storage/reintroduction Germplasm in gene banks, principles for preserving: accession sizes, 113–14 conclusions, 134 deteriorative reactions, underlying precept of controlling, 114–17 goals of gene banks, 113 reaction kinetics cellular constituents, 117–18 temperature, 118–19 water, 119–24
496
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Germplasm in gene banks, (continued) recalcitrant and intermediate seeds, 131–33 temperature’s role revisited, 128–31 water contents for storage, optimum, 124–28 see also Storage types, classification of seed Gesneriads, 180 Gilman International Conservation, 96–97 Ginkgo biloba, 7, 14 Global Diversity Strategy, 88 Global plant conservation, realizing the full potential of: biodiversity-rich regions, 392–99 Botanic Gardens Conservation International, xxv conclusions, 413 how many species can ex situ facilities manage, 400–404 integrated approach, toward a more, 389, 392 recommendations, practical fundraising and outreach, 408–9 in situ management, support to, 407–8 institutional continuity, 411–12 liabilities, managing ex situ, 411 overview, 404 partnerships, 409–11 policy, ex situ, 405 professional capacity, improving, 405–6 selection, species, 407 sustainable use, contribution to, 412–13 techniques for ex situ management of wild plant diversity, 390–91 Global Strategy for Plant Conservation, 400, 405, 406, 483, 482 Goose, Hawaiian, xxiv Gordonia lasianthus, 337 Gould, Stephen J., xxiii Greece, 16 Greening Australia, 50 Grevillea scapigera, 58–59 Groundwater, 43 Guidelines for collection management minimizing risks: artificial selection, 464 best practices for collection maintenance, 468–69
conclusions, 469–70 drift, genetic, 455, 457, 462 hybridization, 467–68 inbreeding, 466–67 mutations, deleterious, 464–66 outbreeding, 467, 468 overview, 454–55 sample management plan, 456–57 selection, 462–64 techniques, description/applications/strengths and weakness of, 458–61 Guidelines for conservation collections of rare/endangered species, revised genetic: Center for Plant Conservation, 423–24 conclusions, 437 cost of reintroduction, demographic, 433–36 evaluation cycle, 432–33 history and direction of guidelines, 419–22 inputs choice of taxa, 426 number and size of populations, 426–27 storage options: seeds or growing plants, 427–28 purposes of a conservation collection, 428–32 size, appropriate sample, 424–25 worksheets, 437, 439–41 Guidelines for seed storage: assessing changes in seed quality, 451 classification of seed storage behavior, 442–44 dry orthodox seeds, procedures to, 449–51 principles, general storage, 443–45 quality influencing longevity, seed, 445–48 shelf life, additional factors affecting, 447, 449 Guidelines for the Maintenance of Orthodox Seeds, 367 Gunnera, 349 Guy’s Hospital, xxvi Habitat loss/degradation, xxiv, 6, 43, 68–69, 268
index Hamamelis, 344 Harvest matrices/rules, seed, 310–11, 320–21 Hawaii, xxiv–xxv, 189–91, 342, 395, 399, 410 see also Micropropagation as a tool for plant genetic conservation Hawaii Rare Plant Restoration Group (HRPRG), 191, 202 Hawtin, Geoffrey, xxvi Helianthus, 275, 339 Heracleum, 171, 350 Herbicides, 252 Heritability of quantitative traits, 288–91, 293 Hevea brasiliensis, 372 Hibiscadelphus, xxiv, 337 Historical continuity in natural systems, 36 Holland, 371 Horticulture and difference between spontaneous nature and deliberated culture, 23 Hotspots, biodiversity, xiii–xiv, 98–99, 104–5, 346, 392–96 Housefly, 457 Human Wellbeing Index, 413 Humidity and deterioration in biological materials, 44, 126–28, 449–51 Hybridization: collection management guidelines, 327–29 conclusions, 353–54 contamination, genetic, 347–49 crossability and sympatry, 338–42 frequency and importance of, 333, 334 guidelines for collection management minimizing risks, 467–68 integrity, loss of genetic, 344–47 invasives, 349–53 management of natural/artificial hybrids conservation tool, hybridization as a, 337 hybrids within populations of a threatened species, 336 overview, 334–35 swarms, hybrid, 336 threatened taxa of hybrid origin, 335 overview, 325–26 pathways and conservation, 342–44 reasons to consider hybrids, five, 326–27, 330
497
sympatry, natural/artificial, 331–32 what is a hybrid, 333 Hyophorbe lagenicaulis, 14 Iguana group, West Indian, 96 Impatiens gordonii, xxviii Inbreeding, 248, 249–51, 257, 277–80, 466–67 In-country conservation, 100, 101 India, 16 Indices Seminum, 350 Indonesia, 401, 404 Inga punctata, 152 Insensitive cases, seed harvests and, 314–15 In situ conservation: benefits of, 67–68 Convention on Biological Diversity, xxiii–xxiv, 10, 67 culture, differences between spontaneous nature and deliberate, 21–23 definitions and usage, 13, 88, 391 description/best applications, 459–60 future difficulties for, 68 global plant conservation, realizing the full potential of, 407–8 reproductivity in the garden and, differences in, 30 spontaneity of wild nature vs. evolutionary history, 36 strengths/weaknesses, 461 Institute for Plant Conservation, 406 Integrated conservation strategies, xiv–xv, xvii, xxvii–xxviii, 8–9, 389, 392 see also Western Australia’s ex situ program; individual subject headings Intermediate seeds, 140, 456 International Agenda for Botanic Gardens in Conservation, 87–88, 405 International Center for Agricultural Research in the Dry Areas (ICARDA), xxvi International Crane Foundation, 91 International Development Targets (IDT), 413 International Plant Genetic Resources Institute (IPGRI), xxv, 2, 112, 125, 475–76 International Technical Conference on Plant Genetic Resources, 476 International treaties and legislation, 9, 10
498
index
International Union for the Conservation of Nature (IUCN), xxv, 2 bryophytes/pteridophytes, 206–7 crisis, plant diversity, 6, 42 global plant conservation, realizing the full potential of, 405 guidelines for management of ex situ populations, 477–78 overview, 476 Reintroduction Specialist Group (RSG), 92 zoos, 89 International Zoo Year Book, 87 Internet, the, 97 Intrinsic values, plants and, 23–26 Introgression, 334, 347 Invasive species, xxiv, 32, 68, 327, 330, 349–53 Invertebrates, 93 In vitro collecting (IVC), 214, 215, 219 In vitro propagation, bryophytes/pteridophytes and, 210–14 Isopyrum biternatum, 171 Isotherm shape and deterioration in biological materials, 122, 123, 126–28 Italy, 16, 371–72 Ituri Forest, 96–97 IUCN Plant Red Data Book (Lucas & Synge), 8 IUCN Red List of Threatened Species, 476 Janet Meakin Poor Research Symposium Series, xviii Japan, 91 Jardin Conservatoire, 8 Jeffersonia diphylla, 176 Juan Fernandez Islands, 349 Kaa-Iya National Park, 97 Kanaloa kahoolawensis, xxiv Kenya, 395–96 Kings Park and Botanic Garden (KPBG), 86, 394–95, 406 Knotweed, Japanese, 351 Kokia cookei, 201 Labordia, 338 Lake Wales Ridge Habitat, 71 Lambertia orbifolia, 53–54 Lamoureux, Charles, 199
Lantana camara, 350 Latin America, 6 Lau, Joel, 199 Legislation: Australia Western Australia Wildlife Conservation Act of 1950, 42 United States Clean Water Act, 77 Defense Appropriations Act of 1991, 76–77 Endangered Species Act (ESA) of 1973, 1, 72–74, 348, 410 Federal Land Management Act, 77 Food, Agriculture, Conservation and Trade Act of 1990, 77 National Environmental Policy Act, 77 National Fish and Wildlife Coordination Act, 77 National Forest Management Act of 1976, 76 National Park Service Organic Act of 1916, 77 Sikes Act of 1960, 77 Legumes, hard-seeded, 147 Lemur, white ruffled, 95 Leucaena, 338, 340, 347 Liabilities, managing ex situ, 411 Lilium, 180, 345 Lily, western, 71 Limbe Botanic Garden, 406 Limonium, 338, 346 Lipid-rich seeds, 122, 126 Liverwort, see Bryophytes/pteridophytes, ex situ conservation methods for Living Pharmacies, xxvii Lobelia monostachya, 199 Lotus berthelotii, 464 Lupinus, 248, 277–80 Lyon Arboretum Micropropagation Laboratory, 189–91, 195 Madagascar, 97, 401, 404 Magnolia, 344 Maladaptation, 237–43, 251–53, 255 Malaysia, 401, 404 Mammals, 92–93 Management, see Guidelines listings Manual of the Flowering Plants of Hawaii, 199
index Marchantia polymorpha, 216 Margyricarpus digynus, 349 Masoala National Park, 97 Mattos, J., xxvii Mauritius, 97 McMillan, Calvin, 273 Mead, Samuel, 270 Meconopsis, 345 Media formulations commonly used in micropropagation, 197–98 Medicinal plants becoming extinct through overuse, xxvi–xxvii Mediterranean hotspots, 393–94 Melbourne Botanic Gardens, 340 Memoranda of understanding (MOU), 409–10 Meristems, apical/axillary/lateral, 198 Metals, heavy, 252 Metaphysical naturalists, 21 Methods, ex situ: conclusions, 16–17 crisis, plant diversity, 6–7 effective, 387–88 see also Guidelines listings; Global plant conservation, realizing the full potential of; Organizations/networks evolution of, 7–10 reluctance to use, 4–5 samples collected, types of, 3–5 tools and facilities, 10–13 see also tools for under Ex situ conservation Mexico, 401, 482 Micromeria, 338 Micropropagation as a tool for plant genetic conservation: clones, 192–93 conclusions, 201–3 decontamination of plant propagules, 195–97 explant collection for propagation, 194 health, plants that are in good, 194–95 indispensable tool, 190, 191–92 juvenile plant tissues, 194 media formulations commonly used, 197–98 meristems, apical/axillary/lateral, 198 original plant genotype, preserving integrity of, 193 overview, 189–90
499
postharvest handling of plant propagules, 195 selection of suitable plant material, 193–94 single plant cell, entire plants produced from a, 192 species that have benefited, 199–201 tissue culture, four basic type of, 197–98 variation in tissue culture, genetic, 193 weaning of micropropagated plants to the greenhouse, 198–99 Milium, 373 Milkweed, Mead’s, 270–72 Millennium Seed Bank, 86, 97, 406, 479 Mimulus, 249 Missouri Botanical Garden, 481 Molecular markers in plant/animal conservation biology, xxvii, 276, 286, 292, 294 see also Genetics, integrating quantitative Monitoring survival rates of stored seed, 432–33 Monogenic traits, 288 Morphological dormancy, 164–65 Morphophysiological dormancy, 162, 165–66 Mortality in seed collections, see Sample decline during storage/reintroduction Morus rubra, 240 Moss, see Bryophytes/pteridophytes, ex situ conservation methods for Mozambique, 102 Munroidendron racemosum, 71 Mutations, 249–51, 256–57, 464–66 National Center for Genetic Resources Preservation (NCGRP), 78, 114, 479–80 National Center for Plant Conservation, 480 National Collection of Rare and Endangered Plants, 14–15, 410, 481 National Council for the Conservation of Plants and Gardens (NCCPG), 11, 340 National Endangered Flora Collection (NEFC), 480 National Fish and Wildlife Foundation, 77 National Genetic Resources Program, 77, 78
500
index
National Park Foundation, 78 National Park Service (NPS), 77 National Plant Germplasm System (NPGS), 78, 114, 479 National Seed Storage Laboratory (NSSL), 479, 481 National Tropical Botanical Garden, 181 National Wildlife Refuge System, 73 Native Plant Conservation Initiative, 78 Native Seeds/SEARCH, 409, 412, 481, 482 Natural and the artificial, the, 21–23 Neem, 150 Neomacounia nitida, 206 Nepenthes, 371 Netherlands, 16, 101 New Caledonia, 399, 404 New England Wildflower Society, 381 New/novel environments, population responses to, see Evolutionary principles, identifying genetic/demographic factors by looking at North America, 91 Northern Hemisphere model, 399 Oceania, 91 Oenothera, 348–49 Oilseeds, 147 Okapi, 96–97 Onopordum, 351 Orchids, 147, 148, 150, 180, 346–47, 396 Organ culture and micropropagation, 198 Organisation for Economic Co-operation and Development, 413 Organizations/networks: conclusions, 482, 484 Global Strategy for Plant Conservation, 483, 482 international plant conservation organizations Botanic Gardens Conservation International, 475 International Plant Genetic Resources Institute, 475–76 International Union for the Conservation of Nature, 476–78 international seed banks and ex situ projects Millennium Seed Bank Project, 479 National Center for Genetic Resources Preservation, 479–80
overview, 476 national and regional networks Australian Network for Plant Conservation, 480 Canadian Botanical Conservation Network, 480–81 Center for Plant Conservation, 481 Native Seeds/SEARCH, 481, 482 overview, 474 Orthodox seeds, 115, 449–51, 456 see also Seed listings Orthotrichum truncato-dentatum, 206 Oryx, Arabian, 85 Osmorhiza longistylis, 172 Outbreeding, 467, 468 Out-of-country conservation, 88, 100 Overuse, plant extinction through, xxvi–xxvii Oxford Botanic Garden, 335 Ozaki, Earl, 199 Paintbrush, yellow-flowered golden, 71 Palms, 148, 180, 349, 371–72 Panax, 172, 375–77 Papaver, 170 Papaya, 148–50 Partnerships, 79–81, 406, 409–11 Peas seeds, 125–26 Peccary, Chaco, 97 Pedicularis furbishiae, 310–11, 313, 316–17 Pennantia baylissiana, 337 Pennisetum, 185 Penstemon eatonii, 366–67 Peregrine Fund, 91 Phaseolus vulgaris, 339 Philippines, 401 Phylogenetic position, 407 Physical dormancy, 163–64 Physiological dormancy, 162, 165 Phytophthora, 52, 394 Picea glauca, 293 Pietermaritzburg Botanic Garden, 104 Pigmentation and heritability of quantitative traits, 288–89 Pine, Torrey, 34, 35 Pinus, 291–93 Pittosporum halophilum, 181 Plantago lanceolata, 462 Plant Conservation Alliance of the United States, 9, 78
index Plant Preservative Mixture (PPM), 196 Plasticity, 254, 275 Plum, Pyne’s ground, 71 Policy and Guidelines for Planning and Coordinating Recovery of Endangered and Threatened Species, 76 Political instability and realizing full potential of global plant conservation, 399 Pollen storage as a conservation tool: angiosperms, 180 collection, 182 conclusions, 186–87 historic uses of stored/transported pollen, 180–81 longevities, 180 pollen-pistil interaction, 339 retrieval/distribution/use, 186 scenarios, conservation, 181 storage conditions, 184 desiccation-sensitive pollen, 185–86 desiccation-tolerant pollen, 184–85 developmental stage, 183–84 viability and vigor, 182–83 Polygenic traits, 288 POPPROJ3, 311 Potentilla robbinsiana, 381 Prairies, 269 Priming, sed, 446 Primula, 343, 345, 371 Principles and Practices of Seed Storage, 442 Propagation of species, revised federal policy for controlled, 74–76 Pteridophytes, see Bryophytes/pteridophytes, ex situ conservation methods for Public support for ex situ conservation, 69–70 Pulsatilla, 343 Purshia tridentata, 296
501
Quality of life, plants holding key to enhanced, xiii Quantitative traits, 287–89 see also Genetics, integrating quantitative Quercus robur, 152, 154
Rancho Santo Ana Botanic Garden, 394 Random amplified polymorphic DNA (RAPD), 270, 272, 276, 286, 292 Rare Hawaiian Plant Program (RHPP), 189, 192, 202 Reaction kinetics, see under Germplasm in gene banks, principles for preserving Recalcitrant seeds, 44, 115, 131–33, 140, 141, 150–56, 193, 456 see also Seed listings Redwood, St. Helena, 337 Reference collections, cultivation in mixed display or, 12–13, 391, 459, 461 Reintroductions/translocations, plant: Bushmallow, San Clemente, 29–30 faked nature, 34–35 genetics, integrating quantitative, 296–97 Genetics and Conservation of Rare Plants, xviii guidelines for conservation collections of rare/endangered species, revised genetic, 433–36 productivity in gardens vs. the wild, 30 zoos, 94–96 see also Sample decline during storage/reintroduction; Western Australia’s ex situ program Reptiles, 93 Research, Western Australia’s ex situ program and biology, 50–52 Restoration ecology, 35–36, 68, 295–97 Restoring Diversity: Strategies for the Reintroduction of Endangered Plants (Falk), xvii Rhino, Sumatran, 95 Rhododendron, 180, 340, 344, 351, 353 Rhus, 163, 164 Risk assessment, 70 see also Seed collection and extinction risk of perennial plants Root rot fungus, 43 Rosa, 344 Royal Botanic Gardens, xviii, xxvi, 86, 97, 101, 207, 298, 371, 406, 409 Rulingia, 55–56 Rumex obtusifolius, 163
R. fluitans, 215, 216 Ragwort, British, 355 RAMAS/Stage, 373
Safety concerns about ex situ conservation, 69 Saintpaulia, 345
502
index
Salix, 338, 347 Sample decline during storage/reintroduction: conclusions, 382–83 cost of reintroduction, estimating demographic, 373–78 minimizing losses, 373 overview, 365–66 plants, ex situ storage as growing, 370–73 reducing the demographic cost of reintroduction, 378–82 seed, ex situ storage as, 366–70 see also Guidelines listings San Clemente Island, 29–30 Santa Barbara Botanic Garden, 394 Saxifraga, 292, 343 Scabiosa, 293 Schiedea adamantis, 199–200 Seed banks: concentrated site of species richness, 3 description/best application, 458 overview, 390 sample decline/survival, 366–70 spore banking similar to, 208 strengths/weaknesses, 461 see also Dormancy-breaking and germination requirements from fewest seeds; Germplasm in gene banks, principles for preserving; Guidelines listings; Storage types, classification of seed; Western Australia’s ex situ program Seed collection and extinction risk of perennial plants: assumptions, some, 312 conclusions, 319–22 demographic models, 306–11 frequent vs. intense harvests, 318–19 modeling protocol, 311 overview, 305–6 responses, classifications of population extinction-prone cases, 316 insensitive cases, 314–15 sensitive cases, 316 response to harvest: and example, 312–14 size of populations and seed harvesting patterns, 317–18 variation within a species, 316–17 Seed Science Research, 125 Selaginella, 211 Selection:
elimination natural, 30 genetic issues, population: soft or hard selection, 279–81 global plant conservation, realizing the full potential of, 407 guidelines for collection management minimizing risks, 462–64 micropropagation, 193–94 Senecio, 335, 339, 373 Sensitive cases, seed harvests and, 316 Sharjah Desert Park, 101 Simplicia, 346 Singapore Botanic Garden, 372 Smoke-stimulated germination, 51, 55 Sodium hypochlorite, 196 Sonchus, 338 Sonora Desert Museum, 409 Sophora, 11, 207, 340 South Africa, xxviii, 102, 394 South America, 91 Spartina, 350 Specialist cultivation in a controlled environment, 12, 390, 458, 461 Species Survival Commission (SSC), xxv, 102, 476 Spontaneous hybridization, 330 Spore banking, 208–10 State and federal partnerships, 79–81 Stochastic simulations examining intensity/frequency of seed harvests, 306–7, 310 Storage types, classification of seed: conclusions, 157 moisture content, seed types in relation to, 141–44 overview, 139–40 phenology/temperature conditions/prestorage moisture status type I, 144–47 type II, 145, 147–50 type III, 150–56 shortcut to types, 156–57 see also Germplasm in gene banks, principles for preserving; Guidelines listings; Pollen storage as a conservation tool; Sample decline during storage/reintroduction Succession, plant, 31, 36 Sustainable use and realizing full potential of global plant conservation, 412–13
index Sympatry, natural/artificial, 331–32, 338–42 Syringa, 344 Tamarin, golden lion, 85, 100, 102 Tanzania, 395–96 Taro, 190 Taxon-specific horticultural regime, plants cultivated under, 12 Tecophilaea cyanocrocus, xxviii Temperature and deterioration in biological materials, 115, 118–19, 128–31, 185, 209, 445–46 see also Germplasm in gene banks, principles for preserving; Pollen storage as a conservation tool; Storage types, classification of seed Tetraplasandra flynnii, 200–201 Texas, 410 Thelypteris altissima, 206 Thistle, Pitcher’s, 71 Threatened Flora Seed Centre (TFSC), 47, 49, 56 see also Extinction/threatened species “Thumb Rules,” Harrington’s, 444–45 Tissue culture storage/propagation, 12, 390, 458, 461 see also Bryophytes/pteridophytes, ex situ conservation methods for; Micropropagation as a tool for plant genetic conservation Tohono O’odham reservation, 482 Tresor Project, 100–101 Trifolium hirtum, 248 Trillium grandiflorum, 171 Triticum, 185, 186, 288 Trochetiopsis ebenus, xxviii Tulipa, 14, 343 Turkey, 16 Tween 20, 196 Typha latifolia, 146 Ultradry storage, 128–30 United Arab Emirates, 101 United Kingdom: botanic gardens, 16 candidate species, identifying, 101 hybridization, 335, 342, 346, 351 national collections of garden plants, 11
503
sample decline/survival, 371 zoos, 91, 103 United Nations, 413, 476 University of California at Berkeley, 394 U.S. Geological Services Biological Resources Division, 78 Utrecht Botanical Garden, 101, 409 Values, plants and intrinsic, 23–26 Variance, genetic, 245–49, 455, 457 Veronica peregrina, 462 Verticordia fimbrilepis, 56–58 Vetch, milk, 71 Vigor/viability, pollen, 182–83 Violets, African, 100 Wahlenbergia, 345 Warburgia salutaris, 149 Water and deterioration in biological materials, see Germplasm in gene banks, principles for preserving; Storage types, classification of seed Western Australia’s ex situ program case studies, 53–59 clearing and associated degradation of vegetation, 42–43 climate change, 44 collaborating agencies/groups/individuals, 49–50 conclusions, 59–60 diversity of species and high rate of endemism, 40–41 germination, seeds that need multiple cues to stimulate, 47–48 groundwater hydrology, 43 habitat loss/degradation, 43 invasive weeds, 32 legislative protection, threatened species listing and, 41–42 recovery: integrated conservation, 44–46 reintroductions/translocations, 52–53 research, conservation biology, 50–52 root rot fungus, 43 South West Botanical Province, 40 threatened species program, 45–48 Wetlands, 71 Wild, conserving biodiversity in the, 67 see also In situ conservation Wildlife Conservation Society (WCS), 97 Wild Screen Project, 100
504
index
Wood, Ken, 200 World Conservation Monitoring Center (WCMC), 6 World Wide Fund for Nature Australia (WWF), 50 World Zoo Conservation Strategy (WZCS), 88–89 Yellowstone-to-Yukon Project, 102 Yew seeds, 125–26 Zea, 185 Zigadenus, 175 Zimbabwe, 102 Zizania palustris, 148–49 Ziziphus, Florida, 71 Zoo Biology, 87 Zoos/botanic gardens, lessons from and opportunities for: advertising/retail sales, commercial approach to, 86–87 changes in world zoos, three remarkable, 84 conclusions candidate species, identifying, 101 economic use of wild species, linking ex situ conservation with the, 102–3
expanding roles in conservation efforts, 104–5 hotspots, developing in-country facilities for biodiversity, 98–99, 104–5 in-country/out-of-country facilities, linking, 100, 101 regional programs, incorporating species conservation with, 101–2 conservation role for zoos since midtwentieth century, 84–85, 88–90 definitions and usage, 88 entertainment value/drawing power of species, 93 future directions, lesson learned and, 96–101 guidelines for collection management minimizing risks, 463 impacts of out-of-country facilities, 94 need, are out-of-country facilities managing species most in, 92–93 plant and animal collections, different nature of, 86 reintroductions and repatriations, 94–96 research and conservation activities, 87 what and where are the main facilities for ex situ conservation, 91–92
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