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.
18
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Ellis, S., and U. S. Seal. 1995. Tools of the trade to aid decision making for species survival. Biodiversity and Conservation 4(6):553–572. 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. Falk, D. A., and P. Olwell. 1992. Scientific and policy considerations in the restoration and reintroduction of endangered species. Rhodora 94:287–315. Feldman, W. R. 1996. Botanical gardens in the 21st century: an endangered species. Public Garden 11(1):6–7. Given, D. R. 1987. What the conservationist requires of ex situ collections. 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. Greuter, W. 1994. Extinctions in Mediterranean areas. Philosophical Transactions of the Royal Society of London B 344:41–46. 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. New York: Chapman & Hall. Hannah, L., D. Lohse, C. Hutchinson, J. L. Carr, and A. Lankerani. 1994. A preliminary inventory of human disturbance of world ecosystems. Ambio 23:246–250. Hernández-Bermejo, J. E., and M. Clemente-Muñoz. 1994. Protección de la Flora en Andalucía. Cordoba, Spain: Agencia de Medio Ambiente, Junta de Andalucia. Heslop-Harrison, J. 1974. Postscript: the Threatened Plants Committee. Succulents in peril, supplement to International Organization for the Study of Succulent Plants. IOS Bulletin 3(3):30–32. Hughes, J. B., G. C. Daily, and P. R. Ehrlich. 1997. Population diversity: its extent and extinction. Science 278:689–692. IUCN/UNEP/WWF. 1980. World Conservation Strategy: Living Resource Conservation for Sustainable Development. Gland, Switzerland: IUCN. IUCN/UNEP/WWF. 1991. Caring for the Earth: A Strategy for Sustainable Living. Gland, Switzerland: IUCN. Kennedy, K. 2002. The Center for Plant Conservation. Endangered Species Bulletin 27(3):5–7. Koopowitz, H., A. Thornhill, and M. Anderson. 1994. A general model for the prediction of biodiversity losses based on habitat conversion. Conservation Biology 8(2):425–438. Lavranos, J. 1974. Letter to IOS Conference. Succulents in peril. Supplement to International Organization for Succulent Plant Study Bulletin 3(3):22–24. Le Maitre, D., C. Gelderblom, L. Maphasa, S. Yssel, M. Van den Belt, and Y. Manuel. 1997. Communicating the value of fynbos: results of a survey of stakeholders. Ecological Economics 22:105–121.
1. Ex Situ Methods: A Vital but Underused Set of Conservation Resources
19
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 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
Series B
30/15 ↓ 20/10 ↓ 15/6b ↓ 5 ↓ 15/6b ↓ 20/10 ↓ 30/15 ↓ 20/10 ↓ 15/6b ↓ 5
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
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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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Bertrand, A. M., M. A. Albuerne, H. Fernandez, A. Gonzalez, and R. Sanchez-Tames. 1999. In vitro organogenesis of Polypodium cambricum. Plant Cell Tissue and Organ Culture 57:65–69. Bold, H. C., C. J. Alexopoulos, and T. Delevoryas. 1980. Morphology of Plants and Fungi. New York: Harper & Row. Bonner, F. 1986. Technologies to maintain tree germplasm diversity. Pages 630–672 in Technologies to Maintain Biological Diversity, Vol. 2. Part D. Washington, DC: U.S. Office of Technology Assessment. Borgan, A. K., and S. K. Naess. 1987. Homogeneity and plant quality of in vitro propagated Nephrolepis exaltata Bostoniensis. Acta Horticulturae 212:433–438. Breznovits, A., and J. Mohay. 1987. In vitro problems related to propagation of different fern species. Acta Horticulturae 212:427–431. Burch, J., and T. Wilkinson. 2002. Cryopreservation of protonemata of Ditrichium cornubicum (Paton) comparing the effectiveness of four cryoprotectant pretreatments. CryoLetters 23:197–208. Camloh, M. 1999. Spore age and sterilization affects germination and early gametophyte development of Platycerium bifurcatum. American Fern Journal 89:124–132. Camloh, M., and N. Gogala. 1991. Platycerium bifurcatum: adventitious bud and root formation without growth regulators in vitro. Acta Horticulturae 289:89–90. Camloh, M., N. Gogala, and J. Rode. 1994. Plant regeneration from leaf explants of the fern Platycerium bifurcatum in vitro. Scientia Horticulturae 56:257–266. Camloh, M., B. Vilhar, J. Zel, and M. Ravnikar. 1999. Jasmonic acid stimulates development of rhizoids and shoots in fern leaf culture. Journal of Plant Physiology 155:798–801. Cheema, H. K., and M. B. Sharma. 1993. In vitro induction and differentiation of gametophytic callus in Ceratopteris thalictoroides. Indian Fern Journal 8:74–77. Cheema, H. K., and M. B. Sharma. 1994. Induction of multiple shoots from adventitious buds and leaf callus in Ceratopteris thalictroides. Indian Fern Journal 11:63–67. Christianson, M. L. 1998. A simple protocol for cryopreservation of mosses. The Bryologist 101:32–35. Clark, J., and V. C. Pence. 1999. In vitro propagation of Lobelia boykinii, a rare wetland species. In Vitro Cellular and Developmental Biology 35:64A. Committee on Managing Global Genetic Resources. 1991. The U.S. National Plant Germplasm System. Washington, DC: National Academy Press. Cooke, R. C. 1979. Homogenisation as an aid in tissue culture propagation of Platycerium and Duvallia. HortScience 14:21–22. De Garcia, E., and L. Furelli. 1987. Clonal mass propagation of the fern Cyrtonium falcatum. Acta Horticulturae 212:655–660. De Maggio, A. E., and R. H. Wetmore. 1961. Growth of fern embryos in culture. Nature 191:94–95. Duckett, J. G., P. Fletcher, R. Francis, H. W. Matcham, D. J. Read, and A. J. Russell. In press. In vitro cultivation of bryophytes: practicalities, promise, progress and problems. In A. Jackson (ed.), Developing an Ex Situ Strategy for the Conservation of Bryophytes in the UK. Richmond, UK: Royal Botanic Gardens, Kew. Dyer, A. F. 1979. The culture of fern gametophytes for experimental investigation. Pages 253–305 in A. F. Dyer (ed.), The Experimental Biology of Ferns. New York: Academic Press.
222
tools of the trade
Dyer, A. F., and S. Lindsay. 1996. Soil spore banks: a new resource for conservation. Pages 153–160 in J. M. Camus, M. Gibby, and R. J. Johns (eds.), Pteridology in Perspective. Richmond, UK: Royal Botanic Gardens, Kew. Dykeman, B. W., and B. G. Cumming. 1985. In vitro propagation of the ostrich fern (Matteuccia struthiopteris). Canadian Journal of Plant Science 65:1025–1032. Ellis, R. H., and E. H. Roberts. 1980. Improved equations for the prediction of seed longevity. Annals of Botany 45:13–30. Fabre, J., and J. Dereuddre. 1990. Encapsulation dehydration: a new approach to cryopreservation of Solanum shoot tips. CryoLetters 11:413–426. Farrar, D. R. 1967. Gametophytes of four tropical fern genera reproducing independently of their sporophytes in the southern Appalachians. Science 155:1266–1267. Fay, M. F. 1992. Conservation of rare and endangered plants using in vitro methods. In Vitro Cellular and Developmental Biology: Plant 28:1–4. Fernández, H., A. Bertrand, and R. Sanchez-Tames. 1991. Micropropagation of Asplenium nidus-avis. Acta Horticulturae 289:113–114. Fernández, H., A. Bertrand, and R. Sanchez-Tames. 1993. In vitro regeneration of Asplenium nidus L. from gametophytic and sporophytic tissue. Scientia Horticulturae 56:71–77. Fernández, H., A. Bertrand, and R. Sanchez-Tames. 1996. Micropropagation and phase change in Blechnum spicant and Pteris ensiformis. Plant Cell Tissue and Organ Culture 44:261–265. Fernández, H., A. Bertrand, and R. Sanchez-Tames. 1999. Biological and nutritional aspects involved in fern multiplication. Plant Cell Tissue and Organ Culture 56:211–214. Fernández, H., and M. A. Revilla. 2003. In vitro culture of ornamental ferns. Plant Cell Tissue and Organ Culture 73:1–13. Ford, M. V. 1992. Growing ferns from spores in sterile culture. Pages 295–297 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). Ford, M. V., and M. F. Fay. 1990. Growth of ferns from spores in axenic culture. Pages 171–180 in J. W. Pollard and J. M. Walker (eds.), Plant Cell and Tissue Culture. Methods in Molecular Biology. Vol. 6. Clifton, NJ: Humana. Fulford, M. 1956. The young stages of the leafy Hepaticae: a résumé. Phytomorphology 6:199–235. George, E. F. 1996. Plant Propagation by Tissue Culture. Part 2: In Practice. Eversley, UK: Exegetics, Ltd. Given, D. R. 1987. What the conservationist requires of ex situ collections. Pages 103–116 in D. Bramwell, O. Hamann, V. Heywood, and H. Synge (eds.), Botanic Gardens and the World Conservation Strategy. London: Academic Press. Goller, K., and J. J. Rybczynski. 1995. In vitro culture used for woody fern Cyathea australis (R. Br.) Domin. vegetative propagation. Acta Societatis Botanicorum Poloniae 64:13–17. Grimsley, N. H., and L. A. Withers. 1983. Cryopreservation of cultures of the moss Physcomitrella patens. CryoLetters 4:251–258. Hallingbäck, T., and B. C. Tan. 1996. Towards a global action plan for endangered bryophytes. Anales del Instituto de Biologia Universidad Nacional Autonoma de Mexico. Serie Botanica 67:213–221.
11. Ex Situ Conservation Methods for Bryophytes and Pteridophytes
223
Hellwege, E. M., K. J. Dietz, and W. Hartung. 1996. Abscisic acid causes changes in gene expression involved in the induction of the landform of the liverwort Riccia fluitans L. Planta 198:423–432. Hennen, G. R., and T. J. Sheehan. 1978. In vitro propagation of Platycerium stemaria. HortScience 13:245. Henson, N. 1979. The aseptic culture of ferns at Kew. The Plantsman 1:74–80. Hicks, G., and P. von Aderkas. 1986. A tissue culture of the ostrich fern Matteuccia struthiopteris (L.) Todaro. Plant Cell Tissue and Organ Culture 5:199–204. Higuchi, H., and W. Amaki. 1989. Effects of 6-benzylaminopurine on the organogenesis of Asplenium nidus L. through in vitro propagation. Scientia Horticulturae 37:351–359. Higuchi, H., W. Amaki, and S. Suzuki. 1987. In vitro propagation of Nephrolepis cordifolia Presl. Scientia Horticulturae 32:105–113. Hirsch, A. M. 1975. The effect of sucrose on the differentiation of excised fern leaf tissues into either gametophytes or sporophytes. Plant Physiology 56:390–393. Hvoslef-Eide, A. K. 1990. In vitro storage as influenced by growth conditions prior to storage. Pages 124–131 in F. De Fong (ed.), Integration of In Vitro Techniques in Ornamental Plant Breeding. Wageningen, the Netherlands: Eucarpia Symposium. Jackson, A. (ed.). In press. Developing an Ex Situ Strategy for the Conservation of Bryophytes in the U.K. Richmond, UK: Royal Botanic Gardens, Kew. Jacobsen, H. J., and G. Dohmen. 1990. Modern plant biotechnology as a tool for the reestablishment of genetic variability in Sophora toromiro. Courier Forschungs Institut Semckenberg 125:233–237. Janssens, J., and M. Sepelie. 1989. In vitro multiplication of Blechnum spp. and Pelaea rotundifolia (Forst.) Hook., by homogenisation. Scientia Horticulturae 38:161–164. Karp, A. 1994. Origins, causes and uses of variation in plant tissue cultures. Pages 139–151 in I. K. Vasil and T. A. Thorpe (eds.), Plant Cell and Tissue Culture. Dordrecht, the Netherlands: Kluwer. Kawakami, S. M., K. Kondo, and S. Kawakami. 1999. Analysis of nucleolar organizer constitution by fluorescent in situ hybridization (FISH) in diploid and artificially produced haploid sporophytes of the fern Osmunda japonica (Osmundaceae). Plant Systematics and Evolution 216:325–331. Kim, J.-K., I. T. Hwang, and H. Kim. 1996. Micropropagation by in vitro spore culture and potting medium in Platycerium bifurcatum [in Korean]. Journal of Agricultural Science Biotechnology 38:267–272. Kiss, H. G., and J. Z. Kiss. 1998. Spore germination in populations of Schizaea pusilla from New Jersey and Nova Scotia. International Journal of Plant Sciences 159:848–852. Knauss, J. F. 1976. A partial tissue culture method of pathogen-free propagation of selected ferns from spores. Proceedings of the Florida State Horticultural Society 89:363–365. Kshirsagar, M. K., and A. R. Mehta. 1978. In vitro studies in ferns: growth and differentiation in rhizome callus of Pteris vittata. Phytomorphology 28:50–58. Kwa, S. H., Y. C. Wee, and P. P. Kumar. 1995. Role of ethylene in the production of sporophytes from Platycerium coronarium (Koenig) Desv. Frond and rhizome pieces cultured in vitro. Journal of Plant Growth Regulation 14:183–189. Kwa, S. H., Y. C. Wee, T. M. Lim, and P. P. Kumar. 1997. Morphogenetic plasticity of callus reinitiated from cell suspension cultures of the fern Platycerium coronarium. Plant Cell Tissue and Organ Culture 48:37–44.
224
tools of the trade
Lal, M. 1984. The culture of bryophytes, including apogamy, apospory, parthenogenesis and protoplasts. Pages 97–115 in A. F. Dyer and J. G. Duckett (eds.), The Experimental Biology of Bryophytes. London: Academic Press. Laliberté, B. 1997. Botanic garden seed banks/gene banks worldwide, their facilities, collections and network. Botanic Gardens Conservation News 2:18–23. Lebkuecher, J. G. 1997. Desiccation-time limits of photosynthetic recovery in Equisetum hyemale (Equisetaceae) spores. American Journal of Botany 84:792–797. Leffring, L., and A. C. Soede. 1982. Influence of salt concentration on the homogeneity of Nephrolepis plants from tissue culture. Pages 499–500 in Plant Tissue Culture 1982. Proceedings of the 5th International Congress of Plant Tissue and Cell Culture. Tokyo: Japanese Association of Plant Tissue Culture. Lloyd, R. M., and E. J. Klekowski Jr. 1970. Spore germination and viability in Pteridophyta: evolutionary significance of chlorophyllous spores. Biotropica 2:129–137. Loescher, W. H., and C. N. Albrecht. 1978. Development in vitro of Nephrolepis exaltata cv. Bostoniensis runner tissues. Physiologia Plantarum 47:250–254. Longton, R. E. 1981. Inter-population variation in morphology and physiology in the cosmopolitan moss Bryum argenteum Hedw. Journal of Bryology 11:501–520. Longton, R. E. In press. Ex vitro cultivation of bryophytes and re-establishment in the field. In A. Jackson (ed.), Developing an Ex Situ Strategy for the Conservation of Bryophytes in the UK. Richmond, UK: Royal Botanic Gardens, Kew. Longton, R. E., and M. W. Holdgate. 1967. Temperature relationships of Antarctic vegetation. Philosophical Transactions of the Royal Society B252:237–250. Lucas, G., and H. Synge. 1978. The IUCN Plant Red Data Book. Morges, Switzerland: IUCN. Materi, D. M., L. A. Dionne, and B. G. Cumming. 1995. In vitro polyploidization of mature ostrich fern sporophytes through rejuvenation. Plant Cell Tissue and Organ Culture 42:27–31. 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 14:1341–1350. Mehra, P. N., and H. K. Palta. 1971. In vitro–controlled differentiation of the root callus of Cyclosorus dentatus. Phytomorphology 29:286–298. Mehra, P. N., and D. S. Sulklyan. 1969. In vitro studies on apogamy, apospory and controlled differentiation of rhizome segments of the fern, Ampelopteris prolifera (Retz.) Copel. Botanical Journal of the Linnean Society 62:431–443. Miller, J. H. 1968. Fern gametophytes as experimental material. Botanical Review 34:361–367. Morgensen, G. S. 1978. Spore development and germination in Cinclidium (Mniaceae, Bryophyta), with special reference to spore mortality and false anisospory. Canadian Journal of Botany 56:1032–1060. Morini, S. 2000. In vitro culture of Osmunda regalis fern. Journal of Horticultural Science & Biotechnology 75:31–34. Mottier, D. M. 1914. Resistance of certain fern prothalli to extreme desiccation. Science 39:295. Murashige, T. 1974. Plant propagation through tissue cultures. Annual Review of Plant Physiology 25:135–166. Murashige, T., and F. Skoog. 1962. A revised medium for rapid growth and bioassays with tobacco tissue cultures. Physiologia Plantarum 15:473–497.
11. Ex Situ Conservation Methods for Bryophytes and Pteridophytes
225
Nakamura, M., and M. Maeda. 1994. Isolation and culture of protoplasts from young sporophytes of Salvinia natans aseptically obtained by co-culture of female and male gametophytes. Plant Cell Tissue and Organ Culture 36:237–242. Oliver, M. J. 1996. Desiccation tolerance in vegetative plant cells. Physiologia Plantarum 97:779–787. Padhya, M. A. 1987. Mass propagation of ferns through tissue culture. Acta Horticulturae 212:645–649. Padhya, M. A. 1995. Rapid propagation of Adiantum trapeziformae L. through tissue culture. Indian Fern Journal 12:20–23. Paek, K. Y., C.-H. Lee, J. K. Choi, and B. H. Kwack. 1984. Mass propagation of Nephrolepis exaltata by runner tips in vitro. Journal of the Korean Horticultural Society 25:313–321. Page, C. N. 1979. Experimental aspects of fern ecology. Pages 552–589 in A. F. Dyer (ed.), The Experimental Biology of Ferns. London: Academic Press. Page, C. N., A. F. Dyer, S. Lindsay, and D. G. Mann. 1992. Conservation of pteridophytes: the ex situ approach. Pages 269–278 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). Pais, M. S. S., and M. Casal. 1987. Propagation of the fern Adiantum copillas-veneris through tissue culture of the circinate part of young leaves. Acta Horticulturae 212:651–654. Pattison, G. A. 1992. National plant collections: ex situ conservation. Pages 305–308 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). Pence, V. C. 1996. In vitro collection (IVC) method. Pages 181–190 in M. N. Normah, M. K. Narimah, and M. M. Clyde (eds.), In Vitro Conservation of Plant Genetic Resources. Kuala Lumpur: Universiti Kebangsaan Malaysia. Pence, V. C. 1998. Cryopreservation of bryophytes: the effects of abscisic acid and encapsulation dehydration. The Bryologist 101:278–281. Pence, V. C. 1999. The application of biotechnology for the conservation of endangered plants. Pages 227–250 in E. E. Benson (ed.), Plant Conservation Biotechnology. London: Taylor & Francis. Pence, V. C. 2000a. Cryopreservation of in vitro grown fern gametophytes. American Fern Journal 90:16–23. Pence, V. C. 2000b. Survival of chlorophyllous and nonchlorophyllous fern spores through exposure to liquid nitrogen. American Fern Journal 90:119–126. Pence, V. C. 2001. Cryopreservation of shoot tips of Selaginella uncinata. American Fern Journal 91:37–40. Pence, V. C., and M. L. Christianson. In press. Cryopreservation as a tool for ex situ preservation of bryophytes. In A. Jackson (ed.), Developing an Ex Situ Strategy for the Conservation of Bryophytes in the UK. Richmond, UK: Royal Botanic Gardens, Kew. Pence, V. C., J. R. Clark, and B. Plair. 2002. Wild and endangered species. Pages 76–82 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.
226
tools of the trade
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.
11. Ex Situ Conservation Methods for Bryophytes and Pteridophytes
227
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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ecological/evolutionary context of ex situ conservation
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
238
ecological/evolutionary context of ex situ conservation
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
240
ecological/evolutionary context of ex situ conservation
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.
12. Population Responses to Novel Environments
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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
244
ecological/evolutionary context of ex situ conservation
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-
12. Population Responses to Novel Environments
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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
246
ecological/evolutionary context of ex situ conservation
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
12. Population Responses to Novel Environments
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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-
248
ecological/evolutionary context of ex situ conservation
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
12. Population Responses to Novel Environments
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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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ecological/evolutionary context of ex situ conservation
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.
Acknowledgments We thank two reviewers for comments and the Natural Sciences and Engineering Research Council of Canada for financial support to B.C.H. References Anonymous. 1988. Wire-lettuce progress. Oryx 22:148. Anonymous. 1989. Rare sunflower planted on South Carolina power line. The Center for Plant Conservation 4:2. Anonymous. 1993. Plant reintroductions in the Mascarenes. Oryx 27:75. Antonovics, J. 1976. The nature of limits to natural selection. Annals of the Missouri Botanical Garden 63:224–247. Barrett, S. C. H., and B. C. Husband. 1997. Ecology and genetics of ephemeral plant populations: Eichhornia paniculata (Pontederiaceae) in northeast Brazil. Journal of Heredity 88:277–284. Barrett, S. C. H., and J. R. Kohn. 1991. Genetic and evolutionary consequences of small population size in plants: implications for conservation. Pages 1–30 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Barrett, S. C. H., M. T. Morgan, and B. C. Husband. 1989. The dissolution of a complex genetic polymorphism: the evolution of self-fertilization in tristylous Eichhornia paniculata (Pontederiaceae). Evolution 43:1398–1416. Bataillon, T., and M. Kirkpatrick. 2000. Inbreeding depression due to mildly deleterious mutations in finite populations: size does matter. Genetic Research 75:75–81. Bowles, M., and R. Flakne. 1993. Recovery planning and reintroduction of the federally threatened Pitcher’s thistle (Cirsium pitcheri) in Illinois. Natural Areas Journal 13:164–176. Bradshaw, A. D. 1984. Ecological significance of genetic variation between populations. Pages 213–228 in R. Dirzo and J. Sarukhán (eds.), Perspectives on Plant Population Ecology. Sunderland, MA: Sinauer.
12. Population Responses to Novel Environments
261
Bradshaw, A. D. 1991. Genostasis and the limits to evolution. The Croonian Lecture. Philosophical Transactions of the Royal Society of London B 333:289–305. Bradshaw, A. D., and K. Hardwick. 1989. Evolution and stress: genotypic and phenotypic components. Biological Journal of the Linnean Society 37:137–155. Brown, A. H. D., and J. D. Briggs. 1991. Sampling strategies for genetic variation in ex situ collections of endangered plant species. Pages 99–122 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. L. Marshall. 1995. A basic sampling strategy: theory and practice. Pages 75–92 in L. Guarino, V. Ramanatha Rao, and R. Reid (eds.), Collecting Plant Diversity: Technical Guidelines. Wallingford, UK: CAB International. Brown, R. W., and R. S. Johnston. 1976. Revegetation of an alpine mine disturbance: Beartooth Plateau, Montana. USDA Forest Service Research Note INT-206. Bürger, R., and M. Lynch. 1995. Evolution and extinction in a changing environment: a quantitative-genetic analysis. Evolution 49:15–163. Bürger, R., G. P. Wagner, and F. Stettingner. 1989. How much heritable variation can be maintained in finite populations by a mutation-selection balance? Evolution 43:1748–1766. Byers, D. L., and D. M. 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 30:479–514. Center for Plant Conservation. 1991. Appendix: genetic sampling guidelines for conservation collections of endangered plants. Pages 225–248 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Chapin, F. S. III, K. Autumn, and F. Pugnaire. 1993. Evolution of suites of traits in response to environmental stress. American Naturalist 142:S78–S92. Charlesworth, D., and B. Charlesworth. 1987. Inbreeding depression and its evolutionary consequences. Annual Review of Ecology and Systematics 18:237–268. Cox, P. 1990. Reintroduction of the Texas snowball (Styrax texana). Endangered Species Update 8:64–65. Cranston, D. M., and D. H. Valentine. 1983. Transplant experiments on rare plant species from Upper Teesdale. Biological Conservation 26:175–191. Davies, M. S., and R. W. Snaydon. 1976. Rapid population differentiation in a mosaic environment. III. Measures of selection pressures. Heredity 36:59–66. De Mauro, M. M. 1994. Relationship of breeding system to rarity in the Lakeside daisy (Hymenoxys acaulis var. glabra). Conservation Biology 7:542–550. Dennison, W. C., and R. S. Alberte. 1986. Photoadaptation and growth of Zostera marina L. (eelgrass) transplants along a depth gradient. Journal of Experimental Marine Ecology 98:265–282. Drake, J. W., B. Charlesworth, D. Charlesworth, and J. F. Crow. 1998. Rates of spontaneous mutation. Genetics 148:1667–1686. Drayton, B., and R. B. Primack. 2000. Rates of success in the reintroduction by four methods of several perennial plant species in eastern Massachusetts. Rhodora 102:299–331. Dudash, M. R. 1990. Relative fitness of selfed and outcrossed progeny in a selfcompatible, protandrous species, Sabatia angularis L. (Gentianaceae): a comparison of three environments. Evolution 44:1129–1139. Edmunson, G., F. Chan, and N. Crane. 1984. The re-establishment of Lupinus sericatus. Restoration and Management Notes 2:85.
262
ecological/evolutionary context of ex situ conservation
Falconer, D. S. 1989. Introduction to Quantitative Genetics. New York: Ronald. Falk, D. A. 1987. Integrated conservation strategies for endangered plants. Natural Areas Journal 7:118–123. Falk, D. A. 1991. Joining biological and economic models for conserving plant genetic diversity. Pages 209–238 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Falk, D. A., and K. E. Holsinger (eds.). 1991. Genetics and Conservation of Rare Plants. New York: Oxford University Press. Fischer, M., and D. Matthies. 1998a. Effects of population size on performance in the rare plant Gentianella germanica. Journal of Ecology 86:195–204. Fischer, M., and D. Matthies. 1998b. RAPD variation in relation to population size and plant fitness in the rare Gentianella germanica (Gentianaceae). American Journal of Botany 85:811–819. Frankel, O. H., and A. H. D. Brown. 1984. Current plant genetic resources: a critical appraisal. Pages 1–11 in V. Chopra (ed.), Genetics: New Frontiers, Vol. 4. New Delhi: India Book House. Frankel, O. H., A. H. D. Brown, and J. J. Burdon. 1995. The Conservation of Plant Biodiversity. Cambridge, UK: Cambridge University Press. Frankham, R. 1996. Relationship of genetic variation to population size in wildlife. Conservation Biology 10:1500–1508. Franklin, I. R. 1980. Evolutionary change in small populations. Pages 135–149 in M. E. Soulé and B. A. Wilcox (eds.), Conservation Biology: An EvolutionaryEcological Perspective. Sunderland, MA: Sinauer. Fritsche, F., and O. Kaltz. 2000. Is the Prunella (Lamiaceae) hybrid zone structured by an environmental gradient? Evidence from a reciprocal transplant experiment. American Journal of Botany 87:995–1003. Galen, C., J. S. Shore, and H. Deyoe. 1991. Ecotypic divergence in alpine Polemonium viscosum: genetic structure, quantitative variation, and local adaptation. Evolution 45:1218–1228. Gomulkiewicz, R., and R. D. Holt. 1995. When does evolution by natural selection prevent extinction? Evolution 49:201–207. Goodman, D. 1987. The demography of chance extinction. Pages 11–34 in M. E. Soulé (ed.), Viable Populations for Conservation. Cambridge, UK: Cambridge University Press. Grant, M. C. 1974. Genetic Properties of Ecologically Marginal Populations of Anthoxanthum odoratum. Ph.D. thesis, Duke University, Durham, NC. Grell, E., N. F. Haas-von Schmude, A. Lamb, and A. Bacon. 1988. Re-introducing Paphiopedilum rothschildianum to Sabh, North Borneo. American Orchid Society Bulletin 57:1138–1246. Guerrant, E. O. Jr. 1990. Transplantation of an otherwise doomed population of Barrett’s penstemon, Penstemon barrettiae. Endangered Species Update 8:66–67. Guerrant, E. O. Jr. 1992. Genetic and demographic considerations in the sampling and reintroduction of rare plants. Pages 321–344 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. 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 Plants. Washington, DC: Island Press. Haldane, J. B. S. 1957. The cost of natural selection. Journal of Genetics 55:511–524.
12. Population Responses to Novel Environments
263
Hawkes, J. G. 1987. A strategy for seed banking in botanic gardens. Pages 131–149 in D. Bramwell, O. Hamman, V. Heywood, and H. Synge (eds.), Botanic Gardens and the World Conservation Strategy. London: Academic Press. Helenurm, K. 1998. Outplanting and differential source population success in Lupinus guadalupensis. Conservation Biology 12:118–127. Helenurm, K., and L. S. Parsons. 1997. Genetic variation and reintroduction of Cordylanthus maritimus ssp. maritimus to Sweetwater Marsh, California. Restoration Ecology 5:236–244. Hoffmann, A. A., and M. W. Blows. 1994. Species borders: ecological and evolutionary perspectives. Trends in Ecology and Evolution 9:223–227. Holtsford, T. P., and N. C. Ellstrand. 1990. Inbreeding effects in Clarkia tembloriensis (Onagraceae) populations with different natural outcrossing rates. Evolution 44:2031–2046. Huenneke, L. F. 1991. Ecological implications of genetic variation in plant populations. Pages 31–44 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Husband, B. C., and S. C. H. Barrett. 1992a. Effective population size and genetic drift in tristylous Eichhornia paniculata (Pontederiaceae). Evolution 46:1875–1890. Husband, B. C., and S. C. H. Barrett. 1992b. Genetic drift and the maintenance of the style-length polymorphism in tristylous Eichhornia paniculata. Heredity 69:440–449. Husband, B. C., and D. W. Schemske. 1995. Magnitude and timing of inbreeding depression in a diploid population of Epilobium angustifolium (Onagraceae). Heredity 75:206–215. Husband, B. C., and D. W. Schemske. 1996. Evolution of the magnitude and timing of inbreeding depression in plants. Evolution 50:54–70. Husband, B. C., and D. W. Schemske. 1997. The effect of inbreeding in diploid and tetraploid populations of Epilobium angustifolium (Onagraceae): implications for the genetic basis of inbreeding depression. Evolution 51:737–746. Jain, S. K., and A. D. Bradshaw. 1966. Evolutionary divergence among adjacent plant populations. I. Evidence and its theoretical analysis. Heredity 21:407–441. Jules, E. S., and A. J. Shaw. 1994. Adaptation to metal-contaminated soils in populations of the moss Ceratodon purpureus: vegetative growth and reproductive expression. American Journal of Botany 81:791–797. Jusaitis, M. 1991. Experimental translocations of endangered Phebalium spp. (Rutaceae) in South Australia: an update. Re-Introduction News 4:1–2. Keightley, P. D., and W. G. Hill. 1988. Quantitative genetic variation maintained by mutation-stabilizing selection balance in finite populations. Genetical Research 52:33–43. Kimura, M., and J. F. Crow. 1963. The measurement of effective population numbers. Evolution 17:279–288. Kirkpatrick, M., and N. H. Barton. 1997. Evolution of a species range. American Naturalist 150:1–23. 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. 1979. Quantitative genetic analysis of multivariate evolution, applied to brain:body size allometry. Evolution 33:402–416. Lande, R. 1982. A quantitative theory of life history evolution. Ecology 63:607–615.
264
ecological/evolutionary context of ex situ conservation
Lande, R. 1988. Genetics and demography in biological conservation. Science 241:1455–1460. Lande, R. 1993. Risks of population extinction from demographic and environmental stochasticity and random catastrophes. American Naturalist 142:911–927. 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. Lande, R., and S. J. Arnold. 1983. The measurement of selection on correlated characters. Evolution 37:1210–1226. Lande, R., and D. W. Schemske. 1985. The evolution of self-fertilization and inbreeding depression in plants. I. Genetic models. Evolution 39:24–40. Lande, R., and S. Shannon. 1996. The role of genetic variation in adaptation and population persistence in a changing environment. Evolution 50:434–437. Lesouef, J. Y. 1988. The rescue of Ruizia cordata and the possible extinction of Astiria rosea. Botanic Gardens Conservation News 1:35–39. Levin, D. A. 1984. Immigration in plants: an exercise in the subjunctive. Pages 242–260 in R. Dirzo and J. Sarukhán (eds.), Perspectives on Plant Population Ecology. Sunderland, MA: Sinauer. Lynch, M., J. Conery, and R. Bürger. 1995. Mutation accumulation and the extinction of small populations. American Naturalist 146:489–518. Lynch, M., and W. Gabriel. 1990. Mutation load and the survival of small populations. Evolution 44:1725–1737. Lynch, M., W. Gabriel, and A. M. Wood. 1991. Adaptive and demographic responses of plankton populations to environmental change. Limnology and Oceanography 36:1301–1312. Lynch, M., and R. Lande. 1993. Evolution and extinction in response to environmental change. Pages 23–250 in P. M. Kareiva, J. G. Kingsolver, and R. B. Huey (eds.), Biotic Interactions and Global Climate Change. Sunderland, MA: Sinauer. Martins, P. S., and S. K. Jain. 1979. Role of genetic variation in the colonizing ability of rose clover (Trifolium hirtum All.). American Naturalist 114:591–595. Maruyama, T., and P. A. Fuerst. 1985a. Population bottlenecks and non-equilibrium models in population genetics. II. Number of alleles in a small population derived from a large steady-state population by means of a bottleneck. Genetics 111:675–689. Maruyama, T., and P. A. Fuerst. 1985b. Population bottlenecks and non-equilibrium models in population genetics. III. Genic homozygosity in populations which experience periodic bottlenecks. Genetics 111:691–703. Maunder, M., A. Culham, A. Bordeu, J. Allainguillaume, and M. Wilkinson. 1999. Genetic diversity and pedigree for Sophora toromiro (Leguminosae): a tree extinct in the wild. Molecular Ecology 8:725–738. May, D. E., P. J. Webber, and T. A. May. 1982. Success of transplanted alpine tundra plants on Niwot Ridge, Colorado. Journal of Applied Ecology 19:965–976. McGraw, J. B., and J. Antonovics. 1983. Experimental ecology of Dryas octopetala ecotypes. I. Ecotypic differentiation and life-cycle stages of selection. Journal of Ecology 71:879–897. McNair, M. R. 1993. Tansley review no. 49. The genetics of metal tolerance in vascular plants. New Phytologist 124:541–559. McNaughton, S. J., T. C. Folsom, T. Lee, F. Park, C. Price, D. Roeder, J. Schmitz, and C. Stockwell. 1974. Heavy metal tolerance in Typha latifolia without the evolution of tolerant races. Ecology 55:1163–1165.
12. Population Responses to Novel Environments
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McNeilly, T., and A. D. Bradshaw. 1968. Evolutionary processes in populations of copper tolerant Agrostis tenuis. Evolution 22:108–118. Muller, H. J. 1964. The relation of recombination to mutational advance. Mutation Research 1:2–9. Nei, M., T. Maruyama, and R. Chakraborty. 1975. The bottleneck effect and genetic variability in populations. Evolution 29:1–10. 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. Obee, E. M., and R. J. Cartica. 1997. Propagation and reintroduction of the endangered hemiparasite Schwalbea americana (Schrophulariaceae). Rhodora 99:134–147. Olivieri, I., D. Couvet, and P. H. Gouyon. 1990. The genetics of transient populations: research at the metapopulation level. Trends in Ecology and Evolution 5:207–210. Olwell, P., A. Cully, and P. Knight. 1990. The establishment of a new population of Pediocactus knowltonii: third year assessment. Ecosystem management: rare species and significant habitats. New York State Museum Bulletin 471:189–193. Ouborg, N. J., and R. Van Treuren. 1995. Variation in fitness-related characters among small and large populations of Salvia pratensis. Journal of Ecology 83:369–380. Parsons, L. S., and J. B. Zedler. 1997. Factors affecting reestablishment of an endangered annual plant at a California salt marsh. Ecological Applications 7:253. Pavlik, B. M., D. L. Nickrent, and A. M. Howald. 1993. The recovery of an endangered plant. I. Creating a new population of Amsinckia grandiflora. Conservation Biology 7:510–526. Pigott, C. D. 1988. The reintroduction of Cirsium tuberosum (L.) All. in Cambridgeshire. Watsonia 17:149–152. Parenti, R. L., and E. O. Guerrant Jr. 1990. Down but not out: reintroduction of the extirpated Malheur wirelettuce, Stephanomeria malheurensis. Endangered Species Update 8:62–63. Reichenbacher, F. W. 1990. Reintroduction brings Kearney’s blue-star from extinction’s edge. Plant Conservation 5:3. Reiling, K., and A. W. Davison. 1992. Spatial variation in ozone resistance of British populations of Plantago major L. New Phytologist 122:699–708. Rieseberg, L. H. 1991. Hybridization in rare plants: insights from case studies in Cercocarpus and Helianthus. Pages 171–181 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Robichaux, R. H., E. A. Friar, and D. W. Mount. 1997. Molecular genetic consequences of a population bottleneck associated with reintroduction of the Mauna Kea silversword (Argyroxiphium sandwicense ssp. sandwicense [Asteraceae]). Conservation Biology 11:1140–1146. Rossetto, M., and K. Dixon. 1993. Corrigin Grevillea in Western Australia. ReIntroduction News 6:8–9. Schemske, D. W., B. C. Husband, M. H. Ruckelshaus, C. Goodwillie, I. M. Parker, and J. G. Bishop. 1994. Evaluating approaches to the conservation of rare and endangered plants. Ecology 75:584–606. Schmidt, K. P., and D. A. Levin. 1985. The comparative demography of reciprocally sown populations of Phlox drummondii. I. Survivorships, fecundities, and finite rates of increase. Evolution 39:396–404. Schoen, D. L., and A. H. D. Brown. 1995. Maximizing genetic diversity in core collections of wild relatives of crop species. Pages 55–76 in T. Hodgkin, A. H. D. Brown,
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T. J. L. van Hintum, and E. A. Morales (eds.), Core Collections of Plant Genetic Resources. New York: Wiley. 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 Science 95:394–399. Shaw, S. C. 1984. Ecophysiological Studies on Heavy Metal Tolerance in Plants Colonizing Tideslow Rake, Derbyshire. Ph.D. thesis, University of Sheffield, Sheffield, UK. Silander, J. A. 1985. The genetic basis of the ecological amplitude of Spartina patens. II. Variance and correlation analysis. Evolution 39:1034–1052. Sork, V. L., K. A. Stowe, and C. Hochwender. 1993. Evidence for local adaptation in closely adjacent subpopulations of northern red oak (Quercus rubra L.) expressed as resistance to leaf herbivores. American Naturalist 142:928–936. Soulé, M., and D. Simberloff. 1986. What do genetics and ecology tell us about the design of nature preserves? Biological Conservation 35:18–40. Taylor, G. J., and A. A. Crowder. 1984. Copper and nickel tolerance in Typha latifolia clones from contaminated and uncontaminated environments. Canadian Journal of Botany 62:1304–1308. Templeton, A. R. 1991. Off-site breeding of animals and implications for plant conservation strategies. Pages 182–194 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Templeton, A. R., and B. Read. 1984. Factors eliminating inbreeding depression in a captive herd of Speke’s gazelle. Zoo Biology 3:177–199. Till-Bottraud, I., L. Wu, and J. Harding. 1990. Rapid evolution of life history traits in populations of Poa annua L. Journal of Evolutionary Biology 3:205–224. Turkington, R., and J. L. Harper. 1979. The growth, distribution and neighbour relationships of Trifolium repens in permanent pastures. IV. Fine-scale biotic differentiation. Journal of Ecology 67:245–254. van Jaarsveld, E. 1994. The establishment of Gasteria baylissiana in its habitat on the Zuurberg, E. Cape (South Africa). British Cactus and Succulent Journal 12:73–74. Via, S. 1994. The evolution of phenotypic plasticity: what do we really know? Pages 58–85 in L. Real (ed.), Ecological Genetics. Princeton, NJ: Princeton University Press. Wallace, S. 1992. Introduction of Conradina glabra: a pilot project for the conservation of an endangered Florida endemic. Botanic Gardens Conservation News 1:34–39. Warwick, S. I. 1991. Herbicide resistance in weedy plants: physiology and population biology. Annual Review of Ecology and Systematics 22:95–114. Warwick, S. I., and D. Briggs. 1980. The genecology of lawn weeds. V. The adaptive significance of different growth habits in lawn and roadside populations of Plantago major L. New Phytologist 85:289–300. Wu, L., and J. Antonovics. 1976. Experimental ecological genetics in Plantago. II. Lead tolerance in Plantago lanceolata and Cynodon dactylon from a roadside. Ecology 57:205–208. Wu, L., and A. L. Kruckeberg. 1985. Copper tolerance in two legume species from a copper mine habitat. New Phytologist 99:565–570. Zeng, Z. B., and C. C. Cockerham. 1991. Variance of neutral genetic variances within and between populations for a quantitative character. Genetics 129:535–553.
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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Ellstrand, N., and D. Elam. 1993. Population genetic consequences of small population size: implications for plant conservation. Annual Review of Ecology and Systematics 24:217–242. Ewens, W. J. 1979. Mathematical Population Genetics. Berlin: Springer-Verlag. Fenster, C., and M. Dudash. 1994. Genetic considerations for plant population restoration and conservation. Pages 34–62 in M. Bowles and C. Whelan (eds.), Restoration of Endangered Species: Conceptual Issues, Planning and Implementation. Cambridge, UK: Cambridge University Press. Frankel, O., A. Brown, and J. Burdon. 1995. The Conservation of Plant Biodiversity. Cambridge, UK: Cambridge University Press. Frankham, R. 1995. Inbreeding and extinction: a threshold effect. Conservation Biology 9:792–799. Frankham, R. 1999. Quantitative genetics in conservation biology. Genetical Research 74:237–244. Frankham, R., and K. Ralls. 1998. Inbreeding leads to extinction. Nature 392:441–442. Gliddon, C., and J. Goudet. 1994. The genetic structure of metapopulations and conservation biology. Pages 107–114 in V. Loeschcke, J. Tomiuk, and S. Jain (eds.), Conservation Genetics. Basel: Birkhauser. Hamrick, J., and M. Godt. 1990. Allozyme diversity in plant species. Pages 43–63 in A. Brown, M. Clegg, A. Kahler, and B. Weir (eds.), Plant Population Genetics, Breeding, and Genetic Resources. Sunderland, MA: Sinauer. Hamrick, J., M. Godt, D. Murawski, and M. Loveless. 1991. Correlations between species traits and allozyme diversity: implications for conservation biology. Pages 75–86 in D. Falk and K. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Hayworth, D., M. Bowles, B. Schaal, and K. Shingleton. 2001. Clonal population structure of the federal threatened Mead’s milkweed, as determined by RAPD analysis and its conservation implication. Pages 182–190 in N. P. Berstein and L. J. Ostrander (eds.), Proceedings of the 17th North American Prairie Conference: Seeds for the Future—Roots of the Past. Mason City: North Iowa Area Community College. Hedrick, P., and S. Kalinowski. 2000. Inbreeding depression and conservation biology. Annual Review of Ecology and Systematics 31:139–162. Helenurm, K., and B. A. Schaal. 1996. Reproductive ecology and breeding system of Lupinus texensis. American Journal of Botany 83:1585–1594. Holsinger, K., R. Mason-Gamer, and J. Whitton. 1999. Genes, demes, and plant conservation. Pages 23–46 in L. Landweber and A. Dobson (eds.), Genetics and the Extinction of Species. Princeton, NJ: Princeton University Press. Jain, S. 1994. Genetics and demography of rare plants and patchily distributed colonizing species. Pages 291–308 in V. Loeschcke, J. Tomiuk, and S. Jain (eds.), Conservation Genetics. Basel: Birkhauser. Lacerda, D., M. Acedo, J. Lemos Filho, and M. Lovato. 2001. Genetic diversity and structure of natural populations of Plathymenia reticulata (Mimosoideae), a tropical tree from the Brazilian Cerrado. Molecular Ecology 10:1143–1152. Lammi, A., P. Siikamaki, and K. Mustajarvi. 1999. Genetic diversity, population size, and fitness in central and peripheral populations of a rare plant, Lychnis viscaria. Conservation Biology 13:1069–1078. Lande, R. 1995. Mutation and conservation. Conservation Biology 9:782–791.
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Lande, R. 1999. Extinction risks from anthropogenic, ecological, and genetic factors. Pages 1–22 in L. Landweber and A. Dobson (eds.), Genetics and the Extinction of Species. Princeton, NJ: Princeton University Press. Leverich, W. J., and D. A. Levin. 1979. Age-specific survivorship and reproduction in Phlox drummondii. American Naturalist 113:881–903. Lynch, M., J. Conery, and R. Burger. 1995. Mutation accumulation and the extinction of small populations. American Naturalist 146:489–519. Matos, J., and B. Schaal. 2000. Chloroplast evolution in the Pinus montezumae complex: a coalescent approach to hybridization. Evolution 54:1218–1233. McGregor, R. 1977. Rare native vascular plants of Kansas. Technical Publication of the State Biological Survey of Kansas 5:1–44. McMillan, C. 1959. The role of ecotypic variation in the distribution of the central grassland of North America. Ecological Monographs 29:285–308. McMillan, C. 1964. Ecotypic differentiation within four North American prairie grasses. II. American Journal of Botany 51:1119–1128. McMillan, C. 1969. Survival patterns in four prairie grasses transplanted to central Texas. American Journal of Botany 56:108–115. Montalvo, A., S. Williams, K. Rice, S. Buchmann, C. Cory, S. Handel, G. Nabhan, R. Primack, and R. Robichaux. 1997. Restoration biology: a population biology perspective. Restoration Ecology 5:277–290. Morgan, J. 1999. Effects of population size on seed production and germinability in an endangered, fragmented grassland plant. Conservation Biology 13:266–273. Podolsky, R. 2001. Genetic variation for morphological and allozyme variation in relation to population size in Clarkia dudleyana, an endemic annual. Conservation Biology 15:412–423. Primack, R. 1996. Lessons from ecological theory: dispersal, establishment and population structure. Pages 209–233 in D. A. Falk, C. I. Millar, and M. Olwell (eds.), Restoring Diversity: Strategies for Reintroduction of Endangered Plants. Washington, DC: Island Press. Rabinowitz, D., J. K. Rapp, and P. M. Dixon. 1984. Competitive abilities of sparse grass species: means of persistence or cause of abundance. Ecology 65:1144–1154. Reed, D., and R. Frankham. 2001. How closely correlated are molecular and quantitative measures of genetic variation? A meta-analysis. Evolution 55:1095–1103. Reinartz, J., and D. Les. 1994. Bottleneck-induced dissolution of self-incompatibility and breeding system consequences in Aster furcatus (Asteraceae). American Journal of Botany 81:446–455. Schaal, B. 1989. The population biology of an annual Texas lupine. In C. H. Stirton and J. Zarucchi (eds.), Advances in Legume Biology. Monographs in Systematic Botany from the Missouri Botanical Garden. 29:283–292. St. Louis: Missouri Botanical Garden Press. Scheiner, S. 1993. Genetics and evolution of phenotypic plasticity. Annual Review of Ecology and Systematics 24:35–68. Schlichting, C. 1986. The evolution of phenotypic plasticity in plants. Annual Review of Ecology and Systematics 17:667–693. Soulé, M. 1986. Conservation Biology: The Science of Scarcity and Diversity. Sunderland, MA: Sinauer.
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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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Ashman, T. L. 1999. Quantitative genetics of floral traits in a gynodioecious wild strawberry Fragaria virginiana: implications for the independent evolution of female and hermaphrodite floral phenotypes. Heredity 83:733–741. Bennington, C. C., and J. B. McGraw. 1996. Environment-dependence of quantitative genetic parameters in Impatiens pallida. Evolution 50:1083–1097. Bonnin, I., J. M. Prosperi, and I. Olivieri. 1997. Comparison of quantitative genetic parameters between two natural populations of a selfing plant species, Medicago truncatula Gaertn. Theoretical Applications of Genetics 94:641–651. Bureau of Land Management. 2003. Technical protocol for the collection, study, and conservation of seeds from native plant species for Seeds of Success. Available at http://www.nps.gov/plants/sos/protocol/index.htm. Butlin, R. K., and T. Tregenza. 1998. Levels of genetic polymorphism: marker loci versus quantitative traits. Philosophical Transactions of the Royal Society of London 353:187–198. Colas, B., I. Olivieri, and M. Riba. 1997. Centaurea corymbosa, a cliff-dwelling species tottering on the brink of extinction: a demographic and genetic study. Proceedings of the National Academy of Sciences USA 94:3471–3476. Cole, C. T., and M. A. Kuchenreuther. 2001. Molecular markers reveal little genetic differentiation among Aconitum noveboracense and A. columbianum (Ranunculaceae) populations. American Journal of Botany 88:337–347. Darwin, C. 1859. On the Origin of Species by Means of Natural Selection. London: John Murray. DeMauro, M. M. 1993. Relationship of breeding system to rarity in the Lakeside daisy (Hymenoxys aucaulis var. glabra). Conservation Biology 7(3):542–550. Donohue, K., E. H. Pyle, D. Messiqua, M. S. Heschel, and J. Schmitt. 2001. Adaptive divergence in plasticity in natural populations of Impatiens capensis and its consequences for performance in novel habitats. Evolution 5:692–702. Elle, E. 1998. The quantitative genetics of sex allocation in the andromonoecious perennial, Solanum carolinense (L.). Heredity 80:481–488. Ennos, R. A., N. R. Cowie, C. J. Legg, and C. Sydes. 1997. Which measures of genetic variation are relevant in plant conservation? A case study of Primula scotica. Pages 73–79 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. Peterborough, UK: Joint Nature Conservation Committee. Falconer, D. S., and T. F. C. Mackay. 1996. Introduction to Quantitative Genetics. 4th edition. Essex, England: Addison Wesley Longman. Falk, D. A., and K. E. Holsinger. 1991. Genetics and Conservation of Rare Plants. New York: Oxford University Press. Farris, M. A. 1988. Quantitative genetic variation and natural selection in Cleome serrulata growing along a mild soil moisture gradient. Canadian Journal of Botany 66:1870–1876. Frankham, R. 1999. Quantitative genetics in conservation biology. Genetical Research 74:237–244. Guerrant, E. O., 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. M. Kareiva (eds.), Conservation Biology for the Coming Decade. New York: Chapman & Hall.
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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 M. J. W. Godt. 1989. Allozyme diversity in plant species. Pages 43–63 in A. H. D. Brown, M. T. Clegg, A. L. Kahler, and B. S. Weir (eds.), Plant Population Genetics, Breeding, and Genetic Resources. Sunderland, MA: Sinauer. Hamrick, J. L., M. J. W. Godt, D. A. Murawski, and M. D. Loveless. 1991. Correlations between species traits and allozyme diversity: implications for conservation biology. Pages 75–86 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Havens, K. 1994. Clonal repeatability of in vitro pollen tube growth rates in Oenothera organensis (Onagraceae). American Journal of Botany 81:161–165. Hedrick, P. W. 2001. Conservation genetics: where are we now? Trends in Ecology and Evolution 16:629–636. Hedrick, P. W., and O. Savolainen. 1996. Molecular and Adaptive Variation: A Perspective for Endangered Plants. Paper read at Southwestern Rare and Endangered Plant Conference, Fort Collins, CO. Holsinger, K. E. 1991. Conservation of genetic diversity in rare and endangered plants. Pages 626–633 in E. C. Dudley (ed.), The Unity of Evolutionary Biology: The Proceedings of the Fourth International Congress of Systematic and Evolutionary Biology. Portland, OR: Dioscorides Press. Holsinger, K. E., and P. Vitt. 1997. The future of conservation biology: what’s a geneticist to do? Pages 206–216 in S. T. A. Pickett, R. S. Ostfeld, M. Shachak, and G. E. Likens (eds.), The Ecological Basis of Conservation: Heterogeneity, Ecosystems, and Biodiversity. New York: Chapman & Hall. Jaramillo-Correa, J. P., J. Beaulieu, and J. Bousquet. 2001. Contrasting evolutionary forces driving population structure at expressed sequence tag polymorphisms, allozymes, and quantitative traits in white spruce. Molecular Ecology 10:2729– 2740. Kercher, S. M., and K. J. Sytsma. 2000. Genetic and morphological variation in populations of the rare prairie annual Agalinis skinneriana (Wood) Britton (Scrophulariaceae). Natural Areas Journal 20:166–175. Knapp, E. E., and A. R. Dyer. 1997. When do genetic considerations require special approaches to ecological restoration? Pages 345–363 in P. L. Fiedler and P. M. Kareiva (eds.), Conservation Biology: For the Coming Decade. New York: Chapman & Hall. Knapp, E. E., and K. J. Rice. 1996. Genetic structure and gene flow in Elymus glaucus (blue wildrye): implications for native grassland restoration. Restoration Ecology 4:1–10. Knapp, E. E., and K. J. Rice. 1998. Comparison of isozymes and quantitative traits for evaluating patterns of genetic variation in purple needlegrass (Nassella pulchra). Conservation Biology 12:1031–1041. Kuittinen, H., A. Mattila, and O. Savolainen. 1997. Genetic variation at marker loci and in quantitative traits in natural populations of Arabidopsis thaliana. Heredity 79:144–152. Lammi, A., P. Siikamaki, and K. Mustajarvi. 1999. Genetic diversity, population size, and fitness in central and peripheral populations of a rare plant Lychnis viscaria. Conservation Biology 13:1069–1078.
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Ledig, F. T., M. T. Conkle, B. Bermejo, T. Eguiluz, P. Hodgskiss, D. R. Johnson, and W. S. Dvorak. 1999. Evidence for an extreme bottleneck in a rare Mexican pinyon: genetic diversity, disequilibrium, and the mating system in Pinus maximartinezii. Evolution 53(1):91–99. Lynch, M. 1996. A quantitative-genetic perspective on conservation issues. Pages 471–501 in J. C. Avise and J. L. Hamrick (eds.), Conservation Genetics: Case Histories from Nature. New York: Chapman & Hall. Meagher, T. R. 1994. The quantitative genetics of sexual dimorphism in Silene latifolia (Caryophyllaceae). II. Response to sex-specific selection. Evolution 48:939–951. Meagher, T. R., J. Antonovics, and R. Primack. 1978. Experimental ecological genetics in Plantago. III. Genetic variation and demography in relation to survival of Plantago cordata, a rare species. Biological Conservation 14:243–257. Nilsson-Ehle, H. 1909. Kreuzungsuntersuchugen an Hafer und Weizen. Lunds Universitets arsskrift, Vol. 5, no. 2, Series 2. Lund, Sweden: Lunds Universitets. Ouborg, N. J., and R. Van Treuren. 1995. Variation in fitness-related characters among small and large populations of Salvia pratensis. Journal of Ecology 83:369–380. Podolsky, R. H. 2001. Genetic variation for morphological and allozyme variation in relation to population size in Clarkia dudleyana, an endemic annual. Conservation Biology 15:412–423. Prentice, H. C. 1992. The structure of morphometric and allozyme variation in relict populations of Gypsophila fastigiata (Caryophyllaceae) in Sweden. Biological Journal of the Linnean Society 47:197–216. Reed, D. H., and R. Frankham. 2001. How closely correlated are molecular and quantitative measures of genetic variation? A meta-analysis. Evolution 55:1095–1103. Schaal, B. A., W. J. Leverich, and S. H. Rogstad. 1991. Comparison of methods for assessing genetic variation in plant conservation biology. Pages 123–134 in D. A. Falk and K. E. Holsinger (eds.), Genetics and Conservation of Rare Plants. New York: Oxford University Press. Schemske, D. W., B. C. Husband, M. H. Ruckleshaus, C. Goodwillie, I. M. Parker, and J. G. Bishop. 1995. Evaluating approaches to the conservation of rare and endangered plants. Ecology 75:584–606. Schwaegerle, K. E., and D. A. Levin. 1991. Quantitative genetics of fitness traits in a wild population of phlox. Evolution 45:169–177. Schwaegerle, K. E., H. McIntyre, and C. Swingley. 2000. Quantitative genetics and the persistence of environmental effects in clonally propagated organisms. Evolution 54:452–461. Steen, S. W., L. Gielly, P. Taberlet, and C. Brochmann. 2000. Same parental species, but different taxa: molecular evidence for hybrid origins of the rare endemics Saxifraga opdalensis and S. svalbardensis (Saxifragaceae). Botanical Journal of the Linnean Society 132:153–164. Storfer, A. 1996. Quantitative genetics: a promising approach for the assessment of genetic variation in endangered species. Trends in Ecology and Evolution 11:343–348. Thiede, D. A. 1998. Maternal inheritance and its effects on adaptive evolution: a quantitative genetic analysis of maternal effects in a natural plant population. Evolution 52:998–1015.
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Waldmann, P., and S. Andersson. 1998. Comparison of quantitative genetic variation and allozyme diversity within and between populations of Scabiosa canescens. Heredity 81:79–86. Widen, B., and S. Andersson. 1993. Quantitative genetics of life-history and morphology in a rare plant, Senecio integrifolius. Heredity 70:503–514. Yang, R., F. C. Yeh, and A. D. Yanchuk. 1996. A comparison of isozyme and quantitative genetic variation in Pinus contorta ssp. latifolia by Fst. Genetics 142:1045–1052. Zhong, G. Y., and C. O. Qualset. 1995. Quantitative genetic diversity and conservation strategies for an allogamous annual species, Dasypyrum villosum (L.) Candargy (Poaceae). Theoretical Applications of Genetics 91:1064–1073.
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
15. Effects of Seed Collection on the Extinction Risk of Perennial Plants
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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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ecological/evolutionary context of ex situ conservation
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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313
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 ( 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 ( 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
317
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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ecological/evolutionary context of ex situ conservation
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-
15. Effects of Seed Collection on the Extinction Risk of Perennial Plants
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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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ecological/evolutionary context of ex situ conservation
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.
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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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ecological/evolutionary context of ex situ conservation
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.
Acknowledgments The authors would like to thank Jim Ault of Chicago Botanic Garden, Cliff Morden of the University of Hawaii, Chris Dalzell of Durban Botanical Garden, and Ed Guerrant of Berry Botanic Garden for sharing information and providing useful case studies. References Abbott, R. J. 1992. Plant invasions, interspecific hybridization and the evolution of new plant taxa. Trends in Ecology and Evolution 7:401–405. Abbott, R. J., D. J. Curnow, and J. A. Irwin. 1995. Molecular systematics of Senecio squalidus L. and its close diploid relatives. Pages 223–237 in D. J. N. Hind et al. (eds.), Advances in Composiae Systematics. London: Royal Botanic Gardens, Kew.
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Abbott, R. J., and R. J. Milne. 1995. Origins and evolutionary effects of invasive weeds. Pages 53–64 in C. H. Stirton (ed.), Weeds in a Changing World. London: British Crop Protection Council Symposium Proceedings 64. Allen, G. A., and C. L. Eccleston. 1998. Genetic resemblance of allotetraploid Aster ascendens to its diploid progenitors Aster falcatus and Aster occidentalis. Canadian Journal of Botany 76:338–344. Allendorf, F. W., R. F. Leary, P. Spruell, and J. K. Wenburg. 2001. The problems with hybrids: setting conservation guidelines. Trends in Ecology and Evolution 1611:613–622. Anonymous. 1989. 1988 Annual Report on the Status of California’s State Listed, Threatened, Endangered Plants and Animals. Sacramento: State of California, Department of Fish and Game. Anonymous. 1992. The New Royal Horticultural Society Dictionary of Gardening. London: Macmillan. Anonymous. 1997. Pennantia baylissiana: our rarest tree, its discovery and propagation. The New Zealand Garden Journal 2:12–13. Arnold, M. L. 1992. Natural hybridization as an evolutionary process. Annual Review of Ecology and Systematics 23:237–261. Arnold, M. L. 1997. Natural Hybridization and Evolution. Oxford, UK: Oxford University Press. Arnold, M. L., M. R. Bulger, J. M. Burke, A. L. Hempel, and J. H. Williams. 1999. Natural hybridization: how low can you go and still be important? Ecology 80:371–381. Ashton, P. A., and R. J. Abbott. 1996. Multiple origins and genetic diversity in the newly arisen allopolyploid species, Senecio cambrensis Rosser Compositae. Heredity 68:25–32. Baker, K., and G. A. Allen. 1977. Hybrid Hibiscadelphus Malvaceae in the Hawaiian Islands. Pacific Science 31:285–291. Bight, C. 1998. Life Out of Bounds: Bioinvasion in a Borderless World. Worldwatch Environmental Alert Series. New York: W.W. Norton. Bramwell, D. 1972. A revision of the genus Echium in Macaronesia. Lagascalia 21:37–115. Bray, R. A. 1986. Leucaena in northern Australia: a review. Forest Ecology and Management 16:345–354. Brochman, C. 1984. Hybridization and distribution of Argyranthemum coronopifolium Asteraceae-Anthemideae in the Canary Islands. Nordic Journal of Botany 4:729–736. Callaway, D. J. 1994. Magnolias. London: B.T. Batsford. Carney, S. E., D. E. Wolf, and L. H. Rieseberg. 2000. Hybridization and forest conservation. Pages 167–182 in A. Young, D. Boshier, and T. Boyle (eds.), Forest Conservation Genetics. Principles and Practice. Collingwood, Australia: CSIRO Publishing. Carr, G. D. 1995. A fully fertile intergeneric hybrid derivative from Argyroxiphium sandwicense subsp. macrocephalum x Dubautia menziesii Asteraceae and its relevance to plant evolution in the Hawaiian Islands. American Journal of Botany 82:1574–1581. Carr, G. D., and D. W. Kyhos. 1986. Adaptive radiation in the Hawaiian silversword alliance Compositae-Madiinae. II. Cytogenetics of artificial and natural hybrids. Evolution 40:959–976.
356
ecological/evolutionary context of ex situ conservation
Carr, G. D., and A. C. Medeiros. 1998. A remnant greensword population from Pu’u Alaea, Maui, with characteristics of Argyroxiphium virescens Asteraceae. Pacific Science 521:61–68. Considine, J. A. 1993. Progress in selection and cultivation of Australian native plants for floriculture. Acta Horticulturae 337:11–18. Crawford, D. J., S. Brauner, M. B. Cosner, and T. F. Stuessy. 1993. Use of RAPD markers to document the origin of the intergeneric hybrid x Margyacaena skottsbergii Rosaceae on the Juan Fernandez Islands. American Journal of Botany 80:89–92. Cremer, K., C. Van Kraayenoord, N. Parker, and S. Streatfield. 1995. Willows spreading by seed. Implications for Australian river management. Australian Journal of Soil and Water Conservation 8:18–27. Cronk, Q. C. B. 1995. A new species and hybrid in the St. Helena endemic genus Trochetiopsis. Edinburgh Journal of Botany 52:205–213. Cronk, Q. C. B., and J. L. Fuller. 1995. Plant Invaders. The Threat to Natural Ecosystems Worldwide. WWF People and Plants Conservation Manuals. London: Chapman & Hall. Cunneen, T. M. 1995. Breeding for improvement of the Marguerite Daisy Argyranthemum spp. Acta Horticulturae 420:101–103. Daehler, C. C., and D. R. Strong. 1997. Hybridization between introduced smooth cordgrass Spartina alterniflora and native Californian cordgrass S. foliosa in San Francisco Bay, California, USA. American Journal of Botany 84:607–611. Davis, M. A., J. P. Grime, and K. Thompson. 2000. Fluctuating resources in plant communities: a general theory of invasibility. Journal of Ecology 88:528–534. De Lamater, R., and W. Hodgson. 1987. Agave arizonica, an endangered species, a hybrid or does it matter? Pages 305–310 in T. S. Elias (ed.), Conservation and Management of Rare and Endangered Plants. Sacramento: California Native Plant Society. Eastwood, A., B. Bytebier, H. Tye, A. Tye, A. Robertson, and M. Maunder. 1998. The conservation status of Saintpaulia. Curtis’s Botanical Magazine 151:49–61. Ellstrand, N. C. 1992. Gene flow by pollen: implications for plant conservation genetics. Oikos 63:77–86. Ellstrand, N. C., H. C. Prentice, and J. F. Hancock. 1999. Gene flow and introgression from domesticated plants into their wild relatives. Annual Review of Ecology and Systematics 30:539–563. Ellstrand, N. C., and K. A. Schierenbeck. 2000. Hybridization as a stimulus for the evolution of invasiveness in plants. Proceedings of the National Academy of Sciences 97:7043–7050. Ellstrand, N. C., R. Whitkus, and L. H. Rieseberg. 1996. Distribution of spontaneous plant hybrids. Proceedings of the National Academy of Sciences 93:5090–5093. Ewel, J. J., D. J. O’Dowd, J. Bergelson, C. C. Daehler, C. M. D’Antonio, L. D. Gomez, D. R. Gordon, R. J. Hobbs, A. Holt, K. R. Hopper, C. E. Hughes, M. LaHart, R. R. B. Leakey, W. Lee, L. Loope, D. Lorence, S. M. Louda, A. E. Lugo, P. B. McEvoy, D. M. Richardson, and P. M. Vitousek. 1999. Deliberate introductions of species: research needs. BioScience 49:619–630. Federal Register. 1996. Endangered and threatened wildlife and plants; proposed policy and proposed rule on the treatment of intercrosses and intercross progeny (the issue of “hybridization”); request for public comment. Federal Register 61(26):4710–4713.
16. Hybridization in Ex Situ Plant Collections
357
Floate, K. D., and T. G. Whitham. 1993. The “hybrid bridge” hypothesis: host shifting via plant hybrid swarms. American Naturalist 141:651–662. Francisco-Ortega, J., A. Santos-Guerra, S. C. Kim, and D. J. Crawford. 2000. Plant genetic diversity in the Canary Islands: a conservation perspective. American Journal of Botany 87:909–919. Gallagher, K. G., K. A. Schierenbeck, and C. M. D’Antonio. 1997. Hybridization and introgression in Carpobrotus spp. Aizoaceae in California. I. Isozyme evidence. American Journal of Botany 84:905–911. 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. Grant, V. 1949. Pollination systems as isolating mechanisms in angiosperms. Evolution 7:51–64. Green, W. C. H., and A. Rothstein. 1998. Translocation, hybridization, and the endangered black-faced impala. Conservation Biology 12:475–480. Griffin, A. R., I. P. Burgess, and L. Wolf. 1988. Patterns of natural and manipulated hybridization in the genus Eucalyptus: a review. Australian Journal of Botany 36:41–46. Hammer, S., and S. Liede. 1990. Natural and artificial hybrids in Mesembryanthemaceae. South African Journal of Botany 56:356–362. Hardig, T. M., P. S. Soltis, and D. E. Soltis. 2000. Diversification of the North American shrub genus Ceanothus Rhamnaceae: conflicting phylogenies from nuclear ribosomal DNA and chloroplast DNA. American Journal of Botany 87:108–123. Harris, S. A., and R. Ingram. 1991. Molecular systematics of the genus Senecio L. I. Hybridization in a British polyploid complex. Heredity 69:1–10. Harrison, R. G. (ed.). 1993. Hybrid Zones and the Evolutionary Process. Oxford, UK: Oxford University Press. Hawkins, J. A., L. W. Olascoaga, C. E. Hughes, J. L. R. C. Jimenez, and P. M. Ruaro. 1999. Investigation and documentation of hybridization between Parkinsonia aculeata and Cercidium praecox Leguminosae: Caesalpinioideae. Plant Systematics and Evolution 216:1–2, 49–68. Hobbs, R. J., and S. E. Humphries. 1994. An integrated approach to the ecology and management of plant invasions. Conservation Biology 9:761–770. Holderegger, R. 1998. Localized hybridization between Saxifraga aizoides and S. mutata Saxifragaceae: reproductive ecology of S. x hausmannii and implications for conservation. Plant Systematics and Evolution 213:21–41. Hollingsworth, M. L., P. M. Hollingsworth, G. I. Jenkins, J. P. Bailey, and C. Ferris. 1998. The use of molecular markers to study patterns of genotypic diversity in some invasive alien Fallopia spp. Polygonaceae. Molecular Ecology 7:1681–1691. Hughes, C. E. 1998. Leucaena: a genetic resources handbook. Tropical Forestry Paper 37:1–274, Oxford Forestry Institute. Hughes, C. E., and S. A. Harris. 1994. The characterisation and identification of a naturally occurring hybrid in the genus Leucaena Leguminosae: Mimosoideae. Plant Systematics and Evolution 192:177–197. Hughes, C. E., and S. A. Harris. 1998. A second spontaneous hybrid in the genus Leucaena Leguminosae: Mimosoideae. Plant Systematics and Evolution 212:53–77. Hughes, C. E., and R. J. Jones. 1999. Environmental hazards of Leucaena. Pages 61–70 in H. M. Shelton (ed.), Leucaena: Adaptation, Quality and Farming Sys-
358
ecological/evolutionary context of ex situ conservation
tems. ACIAR Proceedings 86. Canberra: Australian Centre for International Agricultural Research. Huxel, G. R. 1999. Rapid displacement of native species by invasive species: effects of hybridization. Biological Conservation 89:143–152. Ihlenfeldt, H. D. 1994. Diversification in an arid world: the Mesembryanthemaceae. Annual Review of Ecology and Systematics 25:521–546. Imper, D. K. 1997. Ecology and conservation of Wolf’s evening primrose in northwestern California. Pages 34–40 in T. N. Kaye (ed.), Conservation and Management of Native Plants and Fungi. Corvallis: Native Plant Society of Oregon. Jackson, H. D., D. A. Steane, B. M. Potts, and R. E. Vaillancourt. 1999. Chloroplast DNA evidence for reticulate evolution in Eucalyptus Myrtaceae. Molecular Ecology 8:739–751. Jansen van Vuuren, P. J. 1995. New ornamental crops in South Africa. Acta Horticulturae 397:71–84. Jansen van Vuuren, P. J., J. Coetzee, and A. Coertse. 1991. South African flowering plants with a potential as future floricultural crops. Acta Horticulturae 337:65–71. Jefferson-Brown, M., and K. Pratt. 1997. The Gardener’s Guide to Growing Fritillaries. Newton Abbott, UK: David and Charles. Johnson, S. D. 1996. Pollination, adaptation and speciation models in the Cape flora of South Africa. Taxon 45:59–66. Johnson, S. D., and K. E. Steiner. 2000. Generalization versus specialization in plant pollination systems. Trends in Ecology and Evolution 15:140–143. Kalloo, G., and J. B. Chowdhury (eds.). 1992. Distant Hybridization in Crop Plants. Berlin: Springer-Verlag. Keller, M., J. Kollman, and P. J. Edwards. 2000. Genetic introgression from distant provenances reduces fitness in local weed populations. Journal of Applied Ecology 37:647–659. Klier, K., M. J. Leoschke, and J. F. Wendel. 1991. Hybridization and introgression in white and yellow ladyslipper orchids Cypripedium candidum and C. pubescens. Journal of Heredity 82:305–318. Klips, R. A. 1999. Pollen competition as a reproductive isolating mechanism between two sympatric Hibiscus species Malvaceae. American Journal of Botany 86:269–276. Knobloch, I. W. 1971. Intergeneric hybridization in flowering plants. Taxon 21:97–103. Kuehn, M. M., J. E. Minor, and B. N. White. 1999. An examination of hybridization between cat tail species Typha latifolia and Typha angustifolia using random amplified polymorphic DNA and chloroplast DNA markers. Molecular Ecology 8:1981–1990. Lammers, T. G. 1996. Phylogeny, biogeography and systematics of the Wahlenbergia fernandeziana complex Campanulaceae: Campanuloideae. Systematic Botany 21:397–415. Leapman, M. 2000. The Ingenious Mr. Fairchild. London: Headline Book Publishing. Leebens-Mack, J., and B. G. Milligan. 1998. Pollination biology in hybridizing Baptisia Fabaceae populations. American Journal of Botany 84:500–507. Lesica, P., and F. W. Allendorf. 1999. Ecological genetics and the restoration of plant communities: mix or match? Restoration Ecology 7:42–50.
16. Hybridization in Ex Situ Plant Collections
359
Levin, D. A., J. Francisco-Ortega, and R. K. Jansen. 1996. Hybridization and the extinction of rare plant species. Conservation Biology 10:10–16. Lindell, T. 1998. Breeding systems and crossing experiments in Anemone patens in the Anemone pulsatilla group Ranunculaceae. Nordic Journal of Botany 18:549–561. Lodge, D. M. 1993. Biological invasions: lessons for ecology. Trends in Ecology and Evolution 8:133–137. Lord, T. 1999. The Royal Horticultural Society Plant Finder. London: Dorling Kindersley. Low, T. 1999. Feral Future: The Untold Story of Australia’s Exotic Invaders. Victoria, Australia: Viking for Penguin. Lowe, A. J., and R. J. Abbott. 1996. Origins of the new allopolyploid species Senecio cambrensis Asteraceae and its relationship to the Canary Islands endemic Senecio teneriffae. American Journal of Botany 83:1365–1372. Marrero-Rodríguez, A. 1992. Evolución de la flora Canaria. Pages 1:55–92 in G. Kunkel (ed.), Flora y Vegetación del Archipélago Canario. Trotado Florístico. Canary Islands, Spain: EDIRCA, Las Palmas de Gran Canaria. Martinsen, G. D., and T. G. Whitham. 1994. More birds nest in hybrid cottonwood trees. Wilson Bulletin 106:474–481. Maunder, M. 1992. Plant reintroduction: an overview. Biodiversity and Conservation 1:21–62. Maunder, M. 1992. 1995. Endemic Plants: Options for an Integrated Conservation Strategy. An unpublished report submitted to the ODA British Government Overseas Development Administration and Government of St. Helena. Conservation Projects Development Unit, Royal Botanic Gardens, Kew. 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 145:1341–1350. Maunder, M., S. Higgens, and A. Culham. 2001a. The effectiveness of botanic garden collections in supporting plant conservation: a European case study. Biodiversity and Conservation 103:383–401. Maunder, M., B. Lyte, B. Baker, and J. Dransfield. 2001b. The conservation status and utility of botanic garden palm collections. Biological Conservation 98:259–271. Maunder, M., W. Page, J. Mauremootoo, R. Payendee, Y. Mungroo, A. Maljkovic, C. Vericel, and B. Lyte. 2002. The decline and conservation management of the threatened endemic palms of the Mascarene Islands. Oryx 361:56–65. McAllister, H. A. 1999. The importance of living collections for taxonomy. Pages 3–10 in S. Andrews et al. (eds.), Taxonomy of Cultivated Plants: Third International Symposium. London: Royal Botanic Gardens, Kew. McDade, L. A. 1990. Hybrids and phylogenetic systematics. I. Patterns of character expression in hybrids and their implications for cladistic analysis. Evolution 44:1685–1700. McGregor, M. 1995. Saxifrages: The Complete Cultivars and Hybrids. The Saxifrage Society International Register of Saxifrages. London: Castle Printers. Meikle, R. D. 1973. The development of garden plants from wild species. Pages 114–123 in P. S. Green (ed.), Plants Wild and Cultivated. London: Botanical Society of the British Isles.
360
ecological/evolutionary context of ex situ conservation
Milne, R. I., and R. J. Abbott. 2000. Origin and evolution of invasive naturalized material of Rhododendron ponticum L. in the British Isles. Molecular Ecology 9:541–556. Milne, R. I., R. J. Abbott, K. Wolff, and D. F. Chamberlain. 1999. Hybridization among sympatric species of Rhododendron Ericaceae in Turkey: morphological and molecular evidence. American Journal of Botany 86:1776–1785. Morgan, E. R., G. K. Burge, J. F. Seelye, M. E. Hopping, and J. E. Grant. 1999. Production of inter-specific hybrids between Limonium perezii StapfHubb. and Limonium sinuatum L. Mill. Euphytica 102:109–115. Mosseler, A. 1990. Hybrid performance and species crossability relationships in willow Salix. Canadian Journal of Botany 68:2329–2338. Motley, T. J., and G. D. Carr. 1998. Artificial hybridization in the Hawaiian endemic genus Labordia Loganiaceae. American Journal of Botany 85:654–660. Myers, N., R. A. Mittermeier, C. G. Mittermeier, G. A. B. Da Foncesca, and J. Kent. 2000. Biodiversity hotspots for conservation priorities. Nature 403:853–858. NCCPG (National Council for the Conservation of Plants and Gardens). The National Plant Collections Directory. London: National Council for the Conservation of Plants and Gardens, Royal Horticultural Society. Neill, D. A. 1988. Experimental studies on species relationships in Erythrina Leguminosae: Papilionoideae. Annals Missouri Botanical Garden 75:886–969. Nelson, E. C. 1994. Ergasiophygophytes in the British Isles: plants that jump the garden fence. Pages 17–30 in A. R. Perry and R. Gwynn Ellis (eds.), The Common Ground of Wild and Cultivated Plants: Introductions, Invasions, Control and Conservation. Botanical Society of the British Isles Conference Report 22. London: Botanical Society of the British Isles. Neuffer, B., H. Auge, H. Mesch, U. Amarell, and R. Brandl. 1999. Spread of violets in polluted pine forest: morphological and molecular evidence for the ecological importance of interspecific hybridization. Molecular Ecology 8:365–377. Newman, D., and D. A. Tallmon. 2001. Experimental evidence for beneficial fitness effect of gene flow in recently isolated populations. Conservation Biology 154:1054–1063. Newton, A. C., T. R. Allnutt, A. C. M. Gillies, A. J. Lowe, and R. A. Ennos. 1999. Molecular phylogeography, intraspecific variation and the conservation of tree species. Trends in Ecology and Evolution 14:140–145. Newton, W. C. F., and C. Pellew. 1929. Primula kewensis and its derivatives. Journal of Genetics 20:405–466. Nielsen, L. R. 2000. Natural hybridization between Vanilla claviculata W.Wright Sw. and V. barbellata Rchb.f. Orchidaceae: genetic, morphological and pollination experimental data. Botanical Journal of the Linnean Society 133:285–302. O’Brien, S. J., and E. Mayr. 1991. Bureaucratic mischief: recognizing endangered species and subspecies. Science 251:1187–1188. O’Hanlon, P. C., R. Peakall, and D. T. Briese. 1999. Amplified fragment length polymorphism AFLP reveals introgression in weedy Onopordum thistles: hybridization and invasion. Molecular Ecology 8:1239–1246. Orton, E. R. 1977. Successful hybridization of Gordonia lasianthus L. Ellis x Franklinia alatamaha Marshall. American Association of Arboreta and Botanic Gardens Bulletin 11:81–84.
16. Hybridization in Ex Situ Plant Collections
361
Pacheco, P., T. F. Stuessy, and D. J. Crawford. 1991. Natural interspecific hybridization in Gunnera Gunneraceae of the Juan Fernandez Islands, Chile. Pacific Science 45:389–399. Peakall, R., C. C. Bower, A. E. Logan, and H. I. Nicol. 1997. Confirmation of the hybrid origin of Chiloglottis x prescottiana Orchidaceae: Diurideae. 1. Genetic and morphometric evidence. Australian Journal of Botany 45:839–855. Pickersgill, B. 1993. Interspecific hybridization by sexual means. Pages 63–78 in M. D. Hayward et al. (eds.), Plant Breeding: Principles and Prospects. London: Chapman & Hall. Poppendieck, H. H. 1976. Mesembryanthemums and the problems of their cultivation. Pages 55–60 in J. B. Simmons et al. (eds.), Conservation of Threatened Plants. London: Plenum. Pringle, J. S. 1981. A review of attempted and reported interseries and intersubgeneric hybridization in Syringa Oleaceae. Baileya 21:101–123. Reichard, S. H., and C. W. Hamilton. 1997. Predicting invasions of woody plants introduced into North America. Conservation Biology 11:193–203. Rhymer, J. M., and D. Simberloff. 1996. Extinction by hybridization and introgression. Annual Review of Ecology and Systematics 27:83–109. Ricci, M., and L. Eaton. 1994. The rescue of Wahlenbergia larrainii in Robinson Crusoe Island, Chile. Biological Conservation 68:89–93. Richards, A. J. 1993. Primula. London: Batsford. Rieseberg, L. H. 1991. Hybridization in rare plants: insights from case studies in Helianthus and Cercocarpus. Pages 171–181 in D. A. Falk and K. E. Holsinger (eds.), Conservation of Rare Plants: Biology and Genetics. New York: Oxford University Press. Rieseberg, L. H. 1995. The role of hybridization in evolution: old wine in new skins. American Journal of Botany 82:944–953. Rieseberg, L. H. 1997. Hybrid origins of plant species. Annual Review of Ecology and Systematics 28:359–389. Rieseberg, L. H., and S. J. Brunsfeld. 1991. Molecular evidence and plant introgression. Pages 151–176 in P. S. Soltis et al. (eds.), Molecular Systematics. New York: Chapman & Hall. Rieseberg, L. H., A. Desrochers, and S. J. Youn. 1995. Interspecific pollen competition as a reproductive barrier between sympatric species of Helianthus (Asteraceae). American Journal of Botany 82:515–519. Rieseberg, L. H., and N. C. Ellstrand. 1993. What can molecular and morphological markers tell us about plant hybridization? Critical Reviews in Plant Science 12:213–241. Rieseberg, L. H., and D. Gerber. 1995. Hybridization in the Catalina Island mountain mahogany Cercocarpus traskiae: RAPD evidence. Conservation Biology 9:199–203. Rieseberg, L. H., and R. D. Noyes. 1998. Genetic map–based studies of reticulate evolution in plants. Trends in Plant Science 3:254–259. Rieseberg, L. H., and J. Wendel. 1993. Introgression and its consequences in plants. Pages 70–114 in R. Harrison (ed.), Hybrid Zones and the Evolutionary Process. New York: Oxford University Press. Rob, C. M. 1973. Garden escapes and naturalized plants. Pages 146–151 in P. S. Green (ed.), Plants Wild and Cultivated. London: Botanical Society of the British Isles.
362
ecological/evolutionary context of ex situ conservation
Roberts, H. F. 1965. Plant Hybridization before Mendel. New York: Hafner. Rossetto, M., F. Lucarotti, S. D. Hopper, and K. Dixon. 1997. DNA fingerprinting of a critically endangered relict species or a rare hybrid? Heredity 79:310–318. Rowley, G. D. 1982. Intergeneric hybrids in succulents. The National Cactus and Succulent Journal UK 37:1–6, 76–80. 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. Scott, J. K., and F. D. Panetta. 1993. Predicting the Australian weed status of southern African plants. Journal of Biogeography 20:87–93. Sedgley, M. 1995. Breeding biology of Banksia species for floriculture. Acta Horticulturae 397:155–162. Shapcott, A. 1998. The genetics of Ptychosperma bleeseri, a rare palm from the Northern Territory, Australia. Biological Conservation 85:203–209. Small, E. 1984. Hybridization in the domesticated-weed-wild complex. Pages 195–210 in W. F. Grant (ed.), Plant Biosystematics. Toronto: Academic Press. Smith, J. F., C. C. Burke, and W. L. Wagner. 1996. Interspecific hybridization in natural populations of Cyrtandra Gesneriaceae on the Hawaiian Islands: evidence from RAPD markers. Plant Systematics and Evolution 200:61–77. Snogerup, S. 1979. Cultivation and continued holding of Aegean endemics in an artificial environment. Pages 85–90 in H. Synge and H. Townsend (eds.), Survival or Extinction: The Practical Role of Botanic Gardens in the Conservation of Rare and Threatened Plants. London: The Bentham-Moxon Trust, Royal Botanic Gardens, Kew. Soltis, P. S., and M. A. Gitzendanner. 1999. Molecular systematics and the conservation of rare species. Conservation Biology 13:471–483. Sorensson, C. T., and J. L. Brewbaker. 1994. Interspecific compatibility among 15 species of Leucaena Leguminosae: Mimosoideae via artificial hybridization. American Journal of Botany 81:240–247. Stace, C. A. 1975. Hybridization and the Flora of the British Isles. London: Academic Press. Stace, C. A. 1989. Plant Taxonomy and Biosystematics. London: Edward Arnold. Stebbins, G. L. 1959. The role of hybridization in evolution. Proceedings of the American Philosophical Society 103:231–251. Steen, S. W., L. Gielly, P. Taberlet, and C. Bronchmann. 2000. Same parental species, but different taxa: molecular evidence for hybrid origins of the rare endemics Saxifraga apdalensis and S. svalbardensis Saxifragaceae. Botanical Journal of the Linnean Society 132:153–164. Stirton, C. H. 1978. Plant Invaders: Beautiful but Dangerous. Cape Town, South Africa: Department of Nature and Environmental Conservation. Stirton, C. H. 1999. The naturalised Lantana camara L. Lataneae-Verbenaceae complex in KwaZulu-Natal, South Africa: a dilemma for the culton concept. Pages 311–324 in S. Andrews et al. (eds.), Taxonomy of Cultivated Plants. London: Royal Botanic Gardens, Kew. Strahm, W. A. 1994. The Conservation and Restoration of the Flora of Mauritius and Rodrigues. Ph.D. thesis, the University of Reading, UK.
16. Hybridization in Ex Situ Plant Collections
363
Strand, C. 1998. Asian witch hazels and their hybrids: a history of Hamamelis in cultivation. The New Plantsman 5:231–245. Stuessy, T. F., D. J. Crawford, C. Marticorena, and O. Mario Silva. 1998. Isolating mechanisms and modes of speciation in endemic angiosperms of the Juan Fernandez Islands. Pages 79–96 in T. F. Stuessy and M. Ono (eds.), Evolution and Speciation of Island Plants. Cambridge, UK: Cambridge University Press. Synge, P. M. 1980. Lilies: A Revision of Elwes’ Monograph of the Genus Lilium and Its Supplements. London: B.T. Batsford. Thompson, J. D. 1991. The biology of an invasive plant. BioScience 41(6):393–401. Tobutt, K. R. 1992. Breeding woody ornamentals at East Malling, with particular reference to Sambucus nigra. Acta Horticulturae 320:63–68. Turner, R. 1998. Euphorbias: A Gardeners Guide. London: B.T. Batsford. Uosukainen, M. 1992. Rhododendron brachycarpum subsp. tigerstedtii Nitz.: a transmitter of extreme frosthardiness. Acta Horticulturae 320:77–81. Urbanska, K. M., and E. Landolt. 1998. Patterns and processes of man-influenced hybridization in Cardamine L. Pages 29–48 in L. W. D. Van Raamsdonk and J. C. M. den Nijs (eds.), Plant Evolution in Man-Made Habitats. Proceedings of the VIIth International Symposium of the International Organisation of Plant Biosystematists, Amsterdam. van Gaal, T. M., S. M. Galatowitsch, and M. S. Strefeler. 1998. Ecological consequences of hybridization between a wild species Echinacea purpurea and related cultivar E. purpurea “White Swan.” Scientia Horticulturae 76:73–88. Van Raamsdonk, L. W. D., J. P. Vaneijk, and W. Eikelbloom. 1995. Crossability analysis in subgenus Tulipa of the genus Tulipa L. Botanical Journal of the Linnean Society 117:147–158. Vilà, M., and C. M. D’Antonio. 1998. Fitness of invasive Carpobrotus Aizoaceae hybrids in coastal California. Ecoscience 5:191–199. Vilà, M., E. Weber, and C. M. D’Antonio. 2000. Conservation implications of invasion by plant hybridization. Biological Invasions 2:207–217. Villani, F., A. Sansotta, M. Cherubini, D. Cesaroni, and V. Sbordoni. 1999. Genetic structure of natural populations of Castanea sativa in Turkey: evidence of a hybrid zone. Journal of Evolutionary Biology 12:233–244. Vitousek, P. M., C. M. D’Antonio, L. L. Loope, and R. Westbrooks. 1996. Biological invasions as global environmental change. American Scientist 84:468–478. Walters, K. S., and H. J. Gillett (eds.). 1998. 1997 IUCN Red List of Threatened Plants. Compiled by the World Conservation Monitoring Centre. Gland, Switzerland: World Conservation Union (IUCN). Wang, J., and M. B. Cruzan. 1998. Interspecific mating in the Piriqueta caroliniana Turneraceae complex: effects of pollen load and composition. American Journal of Botany 85:1172–1179. Waser, N. M., L. Chittka, M. V. Price, N. M. Williams, and J. Ollerton. 1996. Generalization in pollination systems, and why it matters. Ecology 77:1043–1060. Waser, N. M., and M. V. Price. 1994. Crossing distance effects in Delphinium nelsonii: outbreeding and inbreeding depression in progeny fitness. Evolution 48:842–852. Webb, D. A., and R. J. Gornall. 1989. Saxifrages of Europe. Bromley, UK: Christopher Helm. Weller, S. G. 1994. Plant rarity and reproductive biology. Pages 90–117 in M. L. Bowles and C. J. Whelan (eds.), Restoration of Endangered Species: Conceptual
364
ecological/evolutionary context of ex situ conservation
Issues, Planning and Implementation. Cambridge, UK: Cambridge University Press. Wendel, J. F., and J. J. Doyle. 1998. Phylogenetic incongruence: window into genome history and molecular evolution. Pages 265–296 in D. S. Soltis et al. (eds.), Molecular Systematics of Plants II. DNA Sequencing. Norwell, MA: Kluwer. Whitham, T. G., G. D. Martinsen, K. D. Floate, H. S. Dungey, B. M. Potts, and P. Keim. 1999. Plant hybrid zones affect biodiversity: tools for a genetic based understanding of community structure. Ecology 80:416–428. Whitham, T. G., P. A. Morroe, and B. M. Potts. 1994. Plant hybrid zones as centers of biodiversity: the herbivore community of two endemic Tasmanian eucalypts. Oecologia 97:481–490. Wigginton, M. J. 1998. British Red Data Book. 1. Vascular Plants. Peterborough, UK: Joint Nature Conservation Committee. Wilks, W. (ed.). 1900. International Conference on Hybridization, the Cross Breeding of Species and the Cross Breeding of Varieties. Journal of the Royal Horticultural Society 24. Wilks, W. (ed.). 1907. Report on the Third International Conference 1906 on Genetics, Hybridization, the Cross Breeding of Varieties and General Plant Breeding. London: Royal Horticultural Society. Williams, E. G., J. L. Rouse, B. F. Palser, and R. B. Knox. 1990. Reproductive biology of Rhododendron. Horticultural Reviews 12:1–67. Williamson, M. 1996. Biological Invasions. London: Chapman & Hall. Winter, J. H. S., and D. J. Botha. 1994. The release of endangered plants into the horticultural trade: conservation or exploitation? Biodiversity and Conservation 32:142–147. Wolf, D. E., N. Takebayashi, and L. H. Rieseberg. 2001. Predicting the risk of extinction through hybridization. Conservation Biology 154:1039–1053. Wolfe, A. D., and W. J. Elisens. 1994. Nuclear ribosomal DNA restriction-site variation in Penstemon section Peltanthera Scrophulariaceae: an evaluation of diploid hybrid speciation and evidence for introgression. American Journal of Botany 81:1627–1635. Yeo, P. F. 1985. Hardy Geraniums. London: Timber Press. Zirckle, C. 1968. An overlooked eighteenth century contribution to plant breeding and plant selection. Journal of Heredity 59:195–198.
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 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 (
